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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
Morten Foldager Pedersen
Department of Life and Sciences & Chemistry,
Roskilde University, Roskilde, Denmark
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 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
Email: [email protected]
Phone: 919 + 843-4916
Dr. Anna Besse, Publishing Editor
Aquatic and Biogeosciences
Van Godewijckstraat 30
P. O. Box 17
3300 AA Dordrecht
The Netherlands
Email: [email protected]
Phone: 31 (0) 651 33 86 01
Southampton Oceanographic centre,
University of Southampton,
Southampton SO14 3ZH, Hants,
Department of Life Sciences and Chemistry,
Roskilde University. P.O. Box 260,
DK-4000 Roskilde,
CNR, Ist. Sperimentale Talassograf,
Spinata S. Raineri, I-98122 Messina,
Institute of Biology,
University of Southern Denmark,
Campusvej 55, DK-5230
Odense M,
Wisconsin Department of Natural Resources,
Department of Natural Resources Research Center,
1350 Femrite Drive, Monona, WI 53761,
Institute of Marine Research,
Department of Zoology,
University of Coimbra,
P-3004517 Coimbra,
Department of Life Sciences and Chemistry,
Roskilde University. P.O. Box 260,
DK-4000 Roskilde,
Institute of Marine Research,
Department of Zoology,
University of Coimbra,
P-3004517 Coimbra,
Department of Life Sciences and Chemistry,
Roskilde University. P.O. Box 260,
DK-4000 Roskilde,
Freshwater Biological Laboratory,
University of Copenhagen,
Helsingørsgade 51, DK-3400 Hillerød,
Boston University, Marine Program,
Marine Biology Laboratory, Woods Hole,
MA 02543,
Freshwater Biological Laboratory,
University of Copenhagen, Helsingørsgade 51,
DK-3400 Hillerød,
Institute of Biology,
University of Southern Denmark,
Campusvej 55, DK-5230 Odense M,
Department of Marine Ecology,
National Environmental Research Institute,
P.O. Box 358, DK-4000 Roskilde,
Dauphin Island Sea Laboratory,
Department of Marine Science,
University of Southern Alabama,
101 Bienville Boulevard, Dauphin Island,
AL 36528,
Netherlands Institute of Ecology,
Centre for Estuarine and Coastal Ecology,
P.O. Box 140, NL-4400 AC Yerseke,
Netherlands Institute of Ecology,
Centre for Estuarine and Coastal Ecology,
P.O. Box 140, NL-4400 AC Yerseke,
Netherlands Institute of Ecology,
Centre for Estuarine and Coastal Ecology,
P.O. Box 140, NL-4400 AC Yerseke,
Netherlands Institute of Ecology,
Centre for Estuarine and Coastal Ecology,
P.O. Box 140, NL-4400 AC Yerseke,
Netherlands Institute of Ecology,
Centre for Estuarine and Coastal Ecology,
P.O. Box 140, NL-4400 AC Yerseke,
Department of Environmental Sciences,
University of Virginia,
P.O. Box 400123
Charlottesville, VI 22904,
KRISTINA S. Gothenburg University,
Department of Marine Ecology,
P.O. Box 461, S-40530 Gothenburg,
School of Marine Science,
Virginia Institute of Marine Science,
College of William and Mary,
Gloucester Point, VI 23602,
Department of Marine Ecology,
National Environmental Research Institute,
Vejlsøvej 25, DK-8600 Silkeborg,
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.
by M.F. Pedersen, S.L. Nielsen and G.T. Banta
1. Introduction
2. Coastal Plant Communities and Nutrient Dynamics
3. Direct Effects of Plants on Nutrient Dynamics
4. Indirect Effects of Plants
5. Hypothesis
6. References
7. Affiliations
by K. Sand-Jensen and S.L. Nielsen
1. Introduction
2. How can estuarine plant communities be characterized?
3. What are the geometric consequences of variations in organism size
and shape?
4. What is the importance of organism size and suspended/attached life
form for water motion and solute transport?
5. What is the significance of plant size and shape for light utilization,
nutrient uptake and growth?
6. What are the basic functional properties and interrelations of
different photosynthetic communities?
7. Summary
8. Acknowledgements
9. References
10. Affiliations
by J. Hauxwell and I. Valiela
1. Introduction
2. Nutrients: background, nutrient limitation, and increased loading
to coastal zones
3. Components of seagrass ecosystems
4. Establishing patterns: effect of nutrient loading on assemblages of
coastal primary producers
5. Understanding processes: how are seagrasses lost?
6. Notes
7. References
8. Affiliations
by M.R. Flindt, J. Neto, C.L. Amos, M.A. Pardal, A. Bergamasco,
C.B. Pedersen and F.Ø. Andersen
1. Introduction
2. Plant transport patterns measured in the field
3. Macroalgae erosion thresholds and settling rates
4. Attached macroalgae sloughing rates
5. Macroalgae settling rates
6. Relating plant transport to growth and loss rates
7. Field studies on plant bound nutrient transport
8. Conclusion
9. Acknowledgements
10. References
11. Affiliations
by J.K. Petersen
1. Introduction
2. Physical forcing
3. Eutrophication and benthic suspension feeding
4. Where to go from now?
5. References
6. Affiliations
by J. Cebrián
1. Introduction
2. Grazing on marine benthic producers under pristine conditions:
extent, control and consequences
3. Towards an understanding of how eutrophication-induced changes
in benthic producers assemblages may affect herbivory
4. Conclusions and suggestions for future research
5. References
6. Affiliation
by G.T. Banta, M.F. Pedersen and S.L. Nielsen
1. Introduction
2. Decomposition
3. Comparing patterns of decomposition and nutrient release during
decomposition among marine primary producers
4. Changes in C, N and P composition during decomposition
5. Nutrient ratios – implications for mineralization and immobilization
6. Decomposition patterns – implications for nutrient retention
7. Case study – detritus dynamics in two small estuaries dominated by
different primary producers
8. Conclusion
9. References
10. Affiliations
by J.J. Middelburg, K. Soetaert, P.M.J. Herman, H.T.S. Boschker
and C.R. Heip
1. Introduction
2. Burial defined
3. Sediment accumulation
4. Nutrient concentrations
5. Importance of primary producers on organic matter burial
6. Conclusions
7. References
8. Affiliations
by K.J. McGlathery, K. Sundbäck and I.C. Anderson
1. Introduction
2. Benthic nutrient cycling processes
3. Sediment – water column fluxes
4. Competition between primary producers
5. The role of primary producers in the estuarine ‘filter’ – fate and
retention of assimilated nutrients
6. Changes in nutrient cycling with shifts in primary producer dominance
7. Conclusion
8. References
9. Affiliations
by N. Risgaard-Petersen
1. Introduction
2. Nitrification and denitrification in sediments
3. Denitrification of nitrate from the water column
4. Denitrification based on NO3 produced by nitrification
5. Alternative NO3 reduction pathways: DNRA
6. Denitrification in sediments colonized by benthic microalgae
7. Seagrasses and the nitrogen cycle
8. Conclusions
9. References
10. Affiliation
by S.L. Nielsen, M.F. Pedersen and G.T. Banta
1. Primary effects of eutrophication
2. Eutrophication the other way around – effects of plants on
nutrient dynamics
3. Conclusion
4. References
5. Affiliations
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
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
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).
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
Total autotrophic biomass (g C m-2)
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
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
Total NPP (g C m year )
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
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.
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
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,
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
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).
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
P uptake (g P m year )
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).
N loading (g N m year )
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
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
M.F. Pedersen, S.L. Nielsen and G.T. Banta
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
Interactions between vegetation and nutrient dynamics
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
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
M.F. Pedersen, S.L. Nielsen and G.T. Banta
microalgae may thus serve to conserve nutrients within the sediment and may even,
under certain circumstances, stimulate a permanent loss of nitrogen through
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.
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
Interactions between vegetation and nutrient dynamics
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.
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.
M.F. Pedersen, S.L. Nielsen and G.T. Banta
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,
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,
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,
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.
Interactions between vegetation and nutrient dynamics
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,
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.
M.F. Pedersen, S.L. Nielsen and G.T. Banta
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,
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.
Interactions between vegetation and nutrient dynamics
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.
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.
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
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.
K. Sand-Jensen and S.L. Nielsen
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
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,
Estuarine primary producers
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
− 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?
2.1 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
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
All aquatic
and terrestrial
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
Estuarine primary producers
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.
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
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
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
Estuarine primary producers
Table 2. Schematic categories of aquatic photosynthetic communities according to growth
habitat, taxonomy (algae vs. vascular plants) and organism size (microalgae vs. macroalgae).
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).
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
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).
A/V-ratio for different volumes
1 µm3
106 µm3
1012 µm3
6 T-1
4.84 x 10-2
4.84 x 10-4
Cylinder, short
∼ 4 T-1
7.95 x 10-2
7.95 x 10-4
Cylinder, long
∼ 4 T-1
17.14 x 10-2
17.14 x 10-4
Flat thallus
∼2 T-1
9.30 x 10-2
9.30 x 10-4
Very flat thallus
∼ 2 T-1
43.06 x 10-2
43.06 x 10-4
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
Estuarine primary producers
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).
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,
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
(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
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.
Estuarine primary producers
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
(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
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).
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
K. Sand-Jensen and S.L. Nielsen
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
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 (%)
Chlorophyll density (mg
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
Photon absorbance (mg -1 DW)
Plant thickness (mm)
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
K. Sand-Jensen and S.L. Nielsen
Chlorophyll concentration (mg g-1 DW)
Tissue thickness (mm)
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).
K. Sand-Jensen and S.L. Nielsen
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
Estuarine primary producers
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
K-selected algae
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).
K. Sand-Jensen and S.L. Nielsen
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
Incident light
(µmol m -2 s -1 )
Inorganic N and P
Increasing growth rate
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
Estuarine primary producers
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.).
K. Sand-Jensen and S.L. Nielsen
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
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
Estuarine primary producers
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)
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
K. Sand-Jensen and S.L. Nielsen
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).
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
Estuarine primary producers
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).
K. Sand-Jensen and S.L. Nielsen
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
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
Estuarine primary producers
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
Light compensation
(µmol m-2 s-1)
Photosynthetic capacity
(mg C g C-1 h-1)
Tissue C/N ratio
Tissue C/P ratio
Nitrogen uptake capacity
(µmol g DW-1 d-1)
Phosphorus uptake
(µmol g DW-1 d-1)
Growth rate
(doublings day-1)
Grazing rate
(% biomass grazed day-1)
Decomposition rate
(ln units day-1)
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
K. Sand-Jensen and S.L. Nielsen
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.
Estuarine primary producers
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
K. Sand-Jensen and S.L. Nielsen
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.
Estuarine primary producers
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
K. Sand-Jensen and S.L. Nielsen
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.
Estuarine primary producers
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
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
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.
The authors would like to thank Drs. Jens Borum and Valery Forbes for important
comments and suggestions on the content of this chapter and for linguistic
Ackerman, J.D. and Okubo, A. (1993). Reduced mixing in a marine macrophyte community. Functional
Ecology, 7, 305-309.
Admiraal, W. (1984). The ecology of estuarine diatoms. In: Round, F.E. and Chapman, D.J. (eds.)
Progress in phycological research, vol. 3: 269-323. Biopress, Bristol.
Aerts, R. (1996). Nutrient resorption from senescing leaves of perennials: are there several patterns.
Journal of Ecology ,84, 597-608.
Agustí S., Duarte C.M. & Kalff J. (1987) Algal cell size and the maximum density and biomass of
phytoplankton. Limnology and Oceanography, 32, 983-986.
Agusti, S., Enriques, S., Frost-Christensen, S., Sand-Jensen, K. and Duarte, C.M. (1994). Light harvesting
among photosynthetic organisms. Functional Ecology , 8, 273-279.
Agusti, S. and Kalff, J. (1989). The influence of growth conditions on the the size dependence of maximal
algal density and biomass. Limnology and Oceanography, 34, 1104-1108.
Aller, R.C. (2001). Transport and reactions in the bioirrigated zone. In: Boudreau, B.P. and Jørgensen, B.B.
(eds.), The benthic boundary layer, 169-301. Oxford University Press, Oxford.
Banse, K. (1982). Cell volumes, maximal growth rates of unicellular algae and ciliates, and the role of
ciliates in the marine pelagial. Limnology and Oceanography, 27, 1059-1071.
Bochardt, M.A. 1996. Nutrients. In R.J. Stevenson, M.L. Bothwell and R.L. Lowe, Algal ecology:
Freshwater benthic ecosystems, 184-227. Academic Press, San Diego.
Borum, J. (1985). 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. (1996). Shallow water and land/sea boundaries. In: Jørgensen, B.B. and Richardson, K. (eds.),
Eutrophication in coastal marine ecosystems, 179-203. American Geophysical Union, Washington
Borum, J. and Sand-Jensen, K. (1996). Is total primary production in shallow coastal waters stimulated by
nitrogen loading? Oikos, 76, 406-410.
Borum, J., Sand-Jensen, K., Binzer, T., Pedersen, O. and Greve, T.M. (in press). Oxygen movement in
seagrasses. In: T. Larkum, R. Orth and C.M. Duarte, Biology of seagrasses.
Estuarine primary producers
Boston, H.L. and Hill, W.R. (1991). Photosynthesis-light relations of stream periphyton communities.
Limnology and Oceanography, 36, 644-656.
Boudreau, B.P. (2001). Solute transport above the sediment-water interface. In: Boudreau, B.P. and
Jørgensen, B.B. (eds.), The benthic boundary layer, 104-126. Oxford University Press.
Carlton, R.G. and Wetzel, R.G. (1988). Phosphorus flux from lake sediments: effect of epipelic algal
oxygen production. Limnology and Oceanography, 24, 419-427.
Cebrian, J. and Duarte, C.M. (1994). The dependence of herbivory on growth rate in natural plant
communities. Functional Ecology, 8, 518-525.
Cebrian, J. and Duarte, C.M. (1995). Plant growth-rate dependence of detrital carbon storage in
ecosystems. Science, 268, 1606-1608.
Chapman, A.R.O. and Cragie, J.S. (1977). Seasonal growth in Laminaria longicruris: relations with
dissolved inorganic nutrients of nitrogen. Marine Biology, 40, 197-205.
Cloern, J.E. (1999). The relative importance of light and nutrient limitation of phytoplankton growth: a
simple index of coastal ecosystem sensitivity to nutrient enrichment. Aquatic Ecology 33: 3-15.
Cloern, J.E. (2001). Our evolving conceptual model of the coastal eutrophication problem. Marine Ecology
Progress Series 210: 223-253.
Dade, W.B., Hogg, A.J. and Boudreau, B.P. (2001). Physics of flow above the sediment-water interface. In
Boudreau, B.P. and Jørgensen, B.B. (eds.), The bentic boundary layer, 6-43. Oxford University Press,
Denny, M.W. (1993). Air and water. The biology and physics of life’s media. Princeton University Press,
Princeton NJ.
Dillon P.J. & Rigler F.H. (1974) A Test of a Simple Nutrient Budget Model Predicting the Phosphorus
Concentration in Lake Water. Journal of the Fisheries Research Board of Canada, 31, 1771-1778.
Duarte C.M. (1991a) Allometric scaling of seagrass form and productivity. Marine Ecology Progress
Series, 77, 289-300.
Duarte C.M. (1991b) Seagrass depth limits. Aquatic Botany, 40, 363-377.
Duarte, C.M. (1992). Nutrient concentrations of aquatic plants. Limnology and Oceanography 37: 882-889.
Duarte, C.M. (1995). Submerged aquatic vegetation in relation to different nutrient regimes. Ophelia 41:
Duarte, C.M. and Cebrian, J. (1996). The fate of marine autotrophic production. Limnology and
Oceanography 8: 1758-1766.
Duarte, C.M., Sand-Jensen, K., Nielsen, S.L., Enriques, S. and Agusti, S. (1995). Comparative functional
ecology. Trends in Ecology and Evolution 10: 418-421.
Enquist, B.J., Brown, J.H. and West, G.B (1998). Allometric scaling of plant energetics and population
density. Nature 395: 433-436.
Enriques, S., Duarte, C.M., Sand-Jensen, K. and Nielsen, S.L. (1996). Broad-scale comparison of
photosynthetic rates of photosynthetic organisms. Oecologia 108: 197-206.
Fairchild, G.W. and Everett, A.C. (1988). Effect of nutrient (N, P, C) enrichment upon standing crop,
species composition and primary production in an oligotrophic lake. Freshwater Biology 19: 57-70.
K. Sand-Jensen and S.L. Nielsen
Falkowski, P.G. and Raven, J.A. (1997). Aquatic photosynthesis. Blackwell Science, Malden,
Fenchel, T. (1974). Intrinsic rate of natural increase: The relationship with body size. Oecologia 14: 317326.
Fenchel, T. (1993). There are more small than large species? Oikos 68: 375-378.
Fenchel, T., Esteban, G.F. and Finlay, B.J. (1997). Local versus global diversity of microorganisms: cryptic
diversity of ciliate protozoa. Oikos 80: 220-225.
Fenchel, T., King, G.M. and Blackburn, T.H. (1998). Bacterial biogeochemistry. The ecophysiology of
mineral cycling, 2. ed. Academic Press, New York.
Fenchel, T. and Straarup, B.J. (1971). Vertical distribution photosynthetic pigments and the penetration of
light in marine sediments. Oikos 22: 171-182.
Foreman, R.E. (1977). Benthic community modification and recovery following intensive grazing by
Strongylocentrotus droebachiensis. Helgoländer wissenschaftliche Meeresuntersuchungen 30, 468484.
Frost-Christensen, H. and Sand-Jensen, K. (1992). The quantum efficiency of photosynthetsis in
macroalgae and submerged angiosperms. Oecologia 91: 377-384.
Gambi, M.C. Newell, A.R.M. and Jumars, P.A. (1990). Flume observations on flow dynamics in Zostera
marina (eelgrass) beds. Marine Ecology Progress Series 61: 159-169.
Geider, R.J., Osborne, B.A. and Raven, J.A. (1985). Growth, photosynthesis and maintenance metabolic
cost in the diatom Phaeodactylum tricornutum at very low light levels. Journal of Phycology 22: 3948.
Geider, R.J., Platt, T. and Raven, J.A. (1986). Size dependence of growth and photosynthesis in diatoms: a
synthesis. Marine Ecology Progress Series 30: 93-104.
Gerard, V.A. (1987). In situ water motion and nutrient uptake by the giant kelp Macrocystis pyrifera.
Marine Biology 69: 51-54.
Geertz-Hansen, O., Sand-Jensen, K., Hansen, D.F. and Christiansen, A. (1993). Growth and grazing
control of abundance of the marine macroalga, Ulva lactuca L. in a eutrophic estuary. Aquatic Botany
46: 101-109.
Gordon, N.D, McMahon T.A. and Finlayson, B.L. (1992). Stream hydrology. An introduction for
ecologists. John Wiley and Sons, New York.
Greve T.M. (2004). Regulation of eelgrass distribution: Stability, mortality and recolonisation. PhD-thesis.
Freshwater Biological Laboratory, University of Copenhagen.
Grime, J.P. (1979). Plant strategies and vegetation processes. Wiley and Sons, Chichester.
Gundersen, J.K. Jørgensen, B.B. (1990). Microstructure of diffusive boundary layers and the oxygen
uptake of the sea floor. Nature 345: 604-607.
Hanisak M.D. (1979). Nitrogen limitation of Codium fragile ssp. tormentosoides as dertermined by tissue
analysis. Marine Biology 50: 333-337.
Hansson, L.A. (1988). Effects of competitive interactions on the benthic biomass development of
planktonic and periphytic algae in lakes. Limnology and Oceanography 33: 121-128.
Hansson, L.A. (1990). Quantifying the impact of periphytic algae on nutrient availability for
phytoplankton. Freshwater Biology 24: 265-273.
Estuarine primary producers
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
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.
K. Sand-Jensen and S.L. Nielsen
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.
Estuarine primary producers
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,
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.
K. Sand-Jensen and S.L. Nielsen
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,
Smith, R.E.H. and Kalff, J. (1982). Size dependent phosphorus uptake kinetics and cell quota in
phytoplankton. Journal of Phycology 18: 275-284.
Estuarine primary producers
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.
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.
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.
S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling:
The IInfluence of Primary Producers, 59-92.
© 2004 Kluwer Academic Publishers. Printed in the Netherlands.
J. Hauxwell and I. Valiela
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
Effects of nutrient loading
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).
Bacterial fixation
Bacterial fixation
Cultivation of N-fixing
Fertilizer production
Burning fossil fuels
Burning/clearing existing
New bioavailable nitrogen
(million metric tons y-1)
2.3. Addition of nutrients to coastal zones
Increased nutrient loads into estuarine waters have resulted from increases in human
populations along coastlines and associated transformation of natural land into urban
development, agricultural land, and recreational facilities (i.e. golf courses, Nixon
1995, Cloern 2001). In addition to the sources of new bioavailable nitrogen resulting
from human activities (Table 1), it has been estimated that 37% of the world’s
population lives within 100 km of the shoreline (Cohen et al. 1997) and that by the
year 2010, 75% of the U.S. population will live within 75 km of the coastline
(Williams et al. 1991). Not only are concentrated populations of people contributing
large quantities of nitrogen to coastal waters via wastewater and fertilizers - but these
nutrients are often being imported to coastal watersheds from inland regions. In the
United States, for example, a vast majority of crop production occurring inland is
imported to support human populations concentrated along coastlines. Therefore,
coastal zones receive a disproportionately high load of anthropogenic nutrients.
In 1913, the process by which artificial fertilizer could be manufactured was discovered in
Germany by Fritz Haber, who synthesized ammonia by combining N2 and hydrogen gas at high
temperature and pressure. This discovery earned Haber the Nobel prize - ecologists at the
time had predicted that the earth had reached its carrying capacity for food production and
that mass starvation was imminent unless nitrogen fertilizer could be manufactured.
J. Hauxwell and I. Valiela
Atmospheric deposition, wastewater disposal, and fertilizer use are the primary
sources of anthropogenic nitrogen delivered to coastal watersheds (Fig. 1). The
nitrogen from these sources is in turn, after substantial biogeochemical
transformations—conveyed to receiving estuaries via surface or groundwater flow.
Atmospheric deposition of nitrogen and sewage outfalls may also deliver nitrogen
directly into estuaries (Fig. 1).
Nutrient loading and alterations in nutrient cycling vary across global landscapes as a
result of population density (Cole et al. 1993, Seitzinger et al. 2002) and agricultural
production (Howarth et al. 2002, Seitzinger et al. 2002). Overall, it has been
estimated that the global flux of phosphorus has increased almost 3-fold since the
increase in human agricultural and industrial activity over the past century (Howarth
et al. 2002). Increase in the pattern of nitrogen use is even greater than that of
phosphorus, and, at the high end, nitrogen inputs have been estimated to have
increased 6-fold to 8-fold (e.g., the north-eastern U.S. and Chesapeake Bay)
(Howarth et al. 2002).
Figure. 1. Sources of anthropogenically-derived nitrogen (atmospheric, fertilizers, and
wastewater) and mechanisms by which nitrogen may be delivered to coastal waters
(precipitation, run-off, groundwater) (prepared by S. Mazzilli).
Effects of nutrient loading
Seagrasses, algae, and certain bacteria make up the assemblage of primary producers
in coastal areas. Under pristine conditions, meadow biomass [but not necessarily
primary production (Montcreiff et al. 1992, Kaldy et al. 2002, Hauxwell et al. 2003;
Table 2)], is dominated by seagrasses. There are approximately 49 species of
seagrasses in the world represented within 12 genera (den Hartog 1970) (with some
debate, estimates of number of species range from 47 to 57), and seagrass meadows
occupy the coastlines of every continent except Antarctica. Seagrasses exhibit both
sexual (monoecious) and vegetative reproductive strategies. In most cases, vegetative
reproduction, via elongation of rhizomes and growth of new clonal shoots,
contributes the majority to overall recruitment within established meadows.
Seagrasses are primarily inter- and subtidal species, with the upper limit of
distribution controlled by physical factors, including desiccation, wave action, and ice
scour (Robertson and Mann 1984), and the lower limit primarily controlled by light
availability (Backman and Barilotti 1976, Dennison 1987, Duarte 1991, Koch and
Beer 1996). Nutrient uptake can occur via roots or leaves. Because seagrass roots
are typically buried in nutrient-rich sediments, growth is often light-limited and not
Seagrass meadows are highly productive, fixing, on average across the globe and
for all species, 1012 g DW m-2 y-1 (Duarte and Chiscano 1999). Seagrass meadows
serve additional ecological roles, including: (1) providing food and shelter for many
marine organisms (including commercially important fish and shellfish species), and
(2) preventing coastal erosion (binding sediments with root/rhizomes, and by
reducing current velocities with aboveground leaves). Seagrass meadows may also
be an important carbon sink for the oceans; despite contributing only 1% to total
oceanic production, seagrasses may account for 12% of total ocean carbon storage
due to their refractory composition (Duarte and Cebrian 1996).
Within seagrass meadows, algae may be visually less conspicuous, but
ecologically important components. From Table 2, it is apparent that even in
relatively healthy seagrass meadows, total primary production is dominated by the
algal components. Algae differ from plants in that they lack vascular structures such
as roots and stems. Algae may be microscopic, such as single-celled phytoplankton,
or macroscopic, and may be free-living (like many fast-growing estuarine species of
filamentous macroalgae) or attached to a substrate (epiphytes on seagrass leaves,
benthic microalgae, and certain species of macroalgae). Because physiological
characteristics of algae result in lower light requirements than seagrasses (Valiela et
al. 1997b), algal growth is typically nutrient-limited.
30 (119)
31 (122)
13 (256)
Montcreiff et al. 1992
values, not
extrapolated to areal
Kinney and Roman
Kaldy et al. 2002
(Budget 3)
69 (518)
53 (180)
0 (0)
Hauxwell et al. 2003
Bass Harbor Marsh, ME,
Lower Laguna Madre, TX,
20 (67)
Hauxwell et al. 2003
Low N estuary, Waquoit
Bay, MA, USA
High N estuary, Waquoit
Bay, MA, USA
Mississippi Sound, USA
46 (905)
14 (55)
24 (468)
31 (233)
27 (94)
25 (96)
17 (339)
Table 2. Relative contribution by different producers in shallow water and seagrass meadows, from published reports for various locations, expressed
as percentage of production by the different producer types calculated after summing available data. In parentheses, magnitude of production is
provided in g C m-2 y-1. ND = not determined.
J. Hauxwell and I. Valiela
Effects of nutrient loading
As nutrients are increasingly delivered to coastal waters, response by assemblages of
primary producers in shallow waters follows a general pattern: nutrient-limited
algal producers respond by growing at faster rates, algal standing stocks increase,
and seagrasses decline (Sand-Jensen 1977, Borum and Wium-Andersen 1980,
Borum 1985, Silberstein et al. 1986, Valiela et al. 1992, Neckles et al. 1994,
Hauxwell et al. 2001).
The associations between increased nitrogen loading, algal production, and
seagrass decline have been established through field observations (Cowper 1978,
Harlin and Thorne-Miller 1981, Neckles et al. 1994, Duarte 1995, Kinney and
Roman 1998) and within experimental mesocosms (Twilley et al. 1985, Burkholder
et al. 1994; Neckles et al. 1994, Short et al. 1995, Taylor et al. 1995a and 1995b,
Moore and Wetzel 2000). Recent and future research aimed at understanding the
numerical relationships between nitrogen loading and shifts in community structure
of estuarine producers at the watershed-estuary scale is invaluable in answering
several questions.
For example, do the patterns observed in small-scale
experimental settings hold for natural systems? At what level of urban development
do ecologically significant changes occur? More specifically, at what loading rate
might we expect seagrasses to disappear? How do patterns vary among estuaries
that differ in key physical features (e.g., water residence time, depth, latitude, etc.)?
In the following sections we will describe the observed changes that occur among
different taxa of primary producers as nutrient loads increase and, where data are
available, quantify the patterns at the watershed-estuary scale.
4.1. Loss of seagrasses and blooms of algae - observations from around the world
and experimental links
Over the past 3 decades, seagrass habitat has disappeared from coastal ecosystems
worldwide. Habitat can be lost due to natural disturbances, including biological
(herbivory, disease, bioturbation), geological (earthquakes), or meteorological
(hurricanes, ice scour) events (Short and Wyllie-Echeverria 1996). Human activities
may also cause seagrass decline, as a result of practices occurring either (1) directly
within the estuary (e.g., dredging, addition of docks, mooring of boats, harvesting of
shellfish using rakes or trawls, use of motorboats) (Orth et al. 2002 and references
therein) or, (2) as a result of urban development and agricultural activities within
adjoining land parcels (e.g., sediment transport, herbicide runoff, increased nutrient
loading (Kemp et al. 1983)). Short and Wyllie-Escheverria (1996) published a
comprehensive list of reports describing large-scale seagrass loss. In Table 3, we
summarize even more recent reports of seagrass habitat loss and the various causes.
We present the combined dataset in aggregate (by decade) in Fig. 2.
The majority of loss occurring at relatively large geographic scales, for various
seagrass species and from sites around the world has been due to anthropogenic
influences within adjoining land parcels (Fig. 2, top). In recent decades, the number
of documented cases of seagrass decline has increased and ∼90% of loss was
J. Hauxwell and I. Valiela
specifically attributed to human-induced disturbances. In particular, loss resulting
from increased nutrient loads or loss that was likely mediated by nutrient loading
(e.g., increased turbidity, degraded water quality, eutrophication, pollution)
accounted for over 60% of the decline attributed specifically to human activities.
Documented cases of seagrass expansion are few (Fig. 2, bottom; Table 4).
These few cases, however, have occurred relatively recently, with the majority
resulting from successful restoration projects. In a handful of situations (e.g.,
Tampa Bay, FL, USA; Johansson and Greening 2000), reductions in nutrient loads
have been achieved so that seagrasses have returned to systems in which they had
been lost.
Along with loss of seagrasses, there have often been concomitant reports of
increased standing stocks of algae worldwide. Phytoplankton (Valiela et al. 1992,
Duarte 1995) and periphyton (Sand-Jensen 1977, Borum and Wium-Andersen 1980,
Borum 1985, Silberstein et al. 1986, Neckles et al. 1994, Frankovich and
Fourqurean 1997; but see also Lin et al. 1996) often increase with nutrient supply,
incurring detrimental effects on seagrasses. Fletcher (1996) listed 50 sites around
the world where macroalgal accumulations (typically comprised of Chlorophytes
such as Chaetomorpha, Cladophora, Enteromorpha, and Ulva spp.) washed up on
beaches were impressive enough to be described as "green tides." In Table 5 we list
several sites around the world in which accumulations of drift algae have been
reported to actually displace seagrasses. Experimental work has repeatedly
demonstrated that macroalgal canopies competitively exclude seagrasses (Holmquist
1997, Hauxwell et al. 2001, Nelson and Lee 2001; but see also Bell and Hall 1997
and Maciá 2000). Though our focus has been documenting the loss of seagrass
habitat, increased standing stocks of fast-growing nutrient limited algae may also
displace slower growing algal taxa (e.g., fucoids in the North Atlantic, Baltic,
Norway, Mediterranean) (Fletcher 1996 and references therein) and corals (LaPointe
and Clark 1992, McCook et al. 2001; but see also Miller et al. 1999 and references
4.2. The numerical relationship between nitrogen loading rates and shifts in
assemblages of primary producers - ecosystems
Given the observational and experimental evidence linking seagrass decline with
algal blooms and increased nutrient loads, there is an obvious need for ecosystemlevel studies that can be applied to guide watershed management. There is some
difficulty in quantifying the response of estuarine producers to nitrogen loading at
relatively large spatial scales (e.g., global or watershed-estuary scales) (Nixon et al.
2001), due largely to 1) acquiring long-term data (significant drain on time and
financial resources), 2) the effort and difficulty involved with measuring loads of
nitrogen delivered from land to sea (again, significant drain on time and financial
resources; Valiela et al. 1992, Tomasko et al. 2001, Howarth et al. 2002, Sigua and
Tweedale 2003), and 3) the ability to observe biological patterns, given potential
wide variation in background physical and other attributes of different coastal areas
(e.g., water residence time, temperature, depth, geology, etc.).
Zostera marina
Syringodium filiforme
Syringodium filiforme
Zostera capricorni
Zostera marina
Zostera marina
Posidonia oceanica
Posidonia oceanica
Halodule wrightii
Posidonia oceanica
Zostera spp.
Halodule wrightii
Airport construction
Increased turbidity
90% in past decade
77% lost over last
Botany Bay, NSW Australia
Barnegat Bay, NJ, USA
Waquoit Bay, MA, USA
(Hamblin and Jehu Ponds)
Niedersachsen, Wadden
18.75 ha
Chronic light reductions
Florida Bay, FL, USA
Florida Bay, FL, USA
Outer FL Bay, FL, USA
Florida Bay, FL, USA
Chronic light reductions
Increased turbidity from fish
Pollution (sewage)
Construction of marinas,
dredging, hydrological
Chronic light reductions
Sea urchins
Salinity and other reasons
Persistent brown tide
2.6 km2 (4%) in 3.5
Laguna Madre, TX, USA
Florida Bay, FL, USA
Harnillo Bay, Murcia Spain
(SE coast)
Gulf of Marseille, France
Corsican coast,
Natural, notably warm El Niño
Area lost
S. Australia
Kastler and Michaelis
Gibbs 1997
Lathrop et al. 2001
Hauxwell et al. 2003
Hall et al. 1999
Hall et al. 1999
Rose et al. 1999
Zieman et al. 1999
Jupp 1977
Pasqualini et al. 1999
Onuf 1996, Street et al.
Hall et al. 1999
Ruiz et al. 2001
Seddon et al. 2000
Table 3. Published reports in which loss of seagrass (on a relatively large (km2) geographic scale) was described including a description of the
disturbance attributed to the decline. These reports were used in addition to those reviewed by Short and Wyllie-Echeverria (1996) to generate Figure 2.
Effects of nutrient loading
Zostera noltii
Zostera spp.
Zostera noltii
Zostera marina
77% lost over last
Niedersachsen, Wadden
German Wadden Sea
S. Australia
New South Wales, Australia
Torres Strait, b/t Papua
New Guinea and Australia
60% loss between 1986
and 1993
70% since 1977
Seto Inland Sea, Japan
Natural, hot El Nino summer
Nutrient loading, drag nets,
reclamation and port
Kastler and Michaelis
Bester 2000
Seddon et al. 2000
Smith 1997
Long et al. 1997
Komatsu 1997
J. Hauxwell and I. Valiela
Effects of nutrient loading
Figure. 2. The number of published reports in which significant loss (top panel) or expansion
(bottom panel) of seagrass areal coverage was documented, shown in aggregate by recent
Zostera capricorni
Zostera capricorni
Zostera marina
Ruppia maritima
Zostera marina
Botany Bay, New
South Wales, Australia
Prince William Sound,
coastal bays of the
Delmarva Peninsula,
Moreton Bay, Australia
Recovery less than 1 year after
extensive dugong grazing
Successful transplantation following
airport construction
32 km increase (~145%) in areal
coverage since opening of the Gulf
Intracoastal Waterway which resulted
in reductions in hypersalinity
Recovery two years after oil spill
56% increase between 1972 and 1993
in holes created by seismic survey
Jervis Bay, New South
Wales, Australia
Success Bank, W.
Laguna Madre, TX,
Posidonia australis
Posidonia coriacea
Amphibolis griffithii
Thalassia testudinum
Recovery after specific disturbance?
87% increase in areal coverage
between 1982 and 1992, following
reductions in nutrient loading
Tampa Bay, FL, USA
Thalassia testudinum
Halodule wrightii
120% increase in areal
coverage between ~1986 and
50% between 1972 and 1993
General expansion
Peterken and Conacher
Gibbs 1997
Orth et al. 2002
Dean et al. 1998
Kaldy and Dunton 1999
Kendrick et al. 1999
Meehan and West 2000
Robbins 1997,
Johansson and Greening
Table 4. Documented cases of seagrass expansion including cases in which expansion was attributed to recovery after an initial disturbance and cases
in which general expansion was noted.
J. Hauxwell and I. Valiela
epiphytes and
Enteromorpha radiata
Enteromorpha linza
Ulva rigida
Gracilaria tikvahiae
Ulvaria obscura
Ulva lactuca
Caulerpa taxifolia
Valiela et al. 1992, Peckol
et al. 1994, Hauxwell et al.
Nelson and Lee 2001
Sfriso 1987, Sfriso et al.
Waquoit Bay, MA,
NW coast, USA
Venice Lagoon, Italy
Zostera marina
Zostera marina
Zostera noltii
Zostera marina
Zostera noltii
Halodule wrightii
Tampa Bay, FL
Johansson and Greening
den Hartog 1994
Zaitsev 1992, Vasiliu 1996
NW Black Sea
Hampshire, UK
Meinesz et al. 1993
N. Mediterranean
Posidonia oceanica
Zostera marina
Zostera noltii
Table 5. Reports in which drift algae replaced seagrasses.
Effects of nutrient loading
J. Hauxwell and I. Valiela
Global-scale regional assessments of loadings have been accomplished (Howarth
et al. 2002, Seitzinger et al. 2002) that reflect broad-scale patterns in population
density, agricultural practices, and industrial and meteorological factors affecting
atmospheric deposition, and may be particularly useful from a management
perspective in terms of realistic goal-setting within a broad geographic region particularly if reductions in nutrient loads cannot be achieved due to regional patterns
in atmospheric deposition (difficult to manage at the smaller watershed-estuary
geographic scale).
Assessment of nutrient loading and its ecological effects at the watershed-estuary
scale seems most appropriate for advancing scientific knowledge related to patterns
and processes associated with seagrass loss, as well as serving as the scale most
appropriate for management and restoration. Development of numerical relationships
among nutrient loading (or concentration), seagrass decline, and algal blooms at the
watershed-estuary scale has recently been accomplished in a few systems, including
Tampa Bay, FL, USA (Johansson and Greening 2000), Danish estuaries (Nielsen et
al. 1989, Sand-Jensen and Borum 1991), and Waquoit Bay, MA, USA (Valiela et al.
1992, Hauxwell et al. 2003). Based on relationships developed in nearby Tampa Bay,
projections of future nitrogen loads and accompanying changes in chlorophyll a
concentrations and seagrass depth limits have been made for Lemon Bay, FL, USA
(Tomasko et al. 2001). Throughout the remaining sections, we highlight our efforts
in Waquoit Bay as a case study to illustrate both patterns in changes to community
structure of primary producers and mechanisms that drive those patterns.
4.3. Case study - Waquoit Bay
Over the past decade, we have conducted research in the Waquoit Bay estuarine
complex located on Cape Cod, MA, USA, where different land use patterns within
the watersheds of seven estuaries, similar in physical characteristics (e.g., depth,
water residence times, temperature, latitude, etc.); have resulted in different annual
loads of nitrogen delivered to those estuaries. Because increased urbanization within
certain watersheds is accompanied by increases in delivery of land-derived nitrogen,
we were able to use the estuaries of Waquoit Bay in a space-for-time substitution
(Pickett 1989) to infer the time course of increased eutrophication created by
increasing urbanization of watersheds. Because the range of nitrogen loads delivered
to Waquoit Bay estuaries encompasses ~75% of the range of reported loads to
different estuaries around the world, the response by estuarine producers to increased
nitrogen loading in Waquoit Bay may be representative of all but the most eutrophic
conditions encountered worldwide.
In Figure 3, we plotted seagrass loss, macroalgae, and phytoplankton versus
nitrogen loading. Seagrass loss occurs rapidly as nitrogen loads increase (Fig. 3, top),
and the pattern observed in the Waquoit Bay estuaries holds at the larger scale; the
same logarithmic relationship was observed using a compilation of published reports
around the world (Valiela and Cole 2002). Seagrasses remain only in estuaries of the
lowest nitrogen loading rates (lowest 12% of the range of loads delivered to estuaries
worldwide). In the Waquoit Bay estuaries, macroalgal canopy heights (Fig. 3,
middle) and phytoplankton (Fig. 3, bottom) increased linearly with nitrogen loading
rate. These sorts of patterns have been observed in other systems; Nielsen et al.
(1989) compared 20 Danish estuaries that spanned a nitrogen loading gradient and
Effects of nutrient loading
demonstrated a positive linear relationship between phytoplankton and total nitrogen
concentration, as well as depth penetration of seagrasses and total nitrogen.
Figure. 3. Relationship between seagrass loss (Zostera marina for the Waquoit estuaries
between 1987 and 1997 adapted from Hauxwell et al. (2003); other reports from around the
world adapted from Valiela and Cole (2002)), maximum macroalgal canopy height (annual
maximum from monthly sampling, adapted from Hauxwell et al. (2003)), and phytoplankton
standing stocks (4-year annual average between 1990 and 1994 monthly sampling, Foreman et
al. (in preparation); lower panel figure prepared by G. Tomasky) and nitrogen loading rates
(modelled, Valiela et al. 1997a) to the Waquoit Bay estuaries.
J. Hauxwell and I. Valiela
4.4. The effect of physical variation among systems on ecosystem-scale patterns
It is important to note that variations in the physical backdrop will effect (and
potentially mask) the patterns we described above. In essence, the relationships
described above (specific to small, shallow, temperate estuaries with a Cape Cod
geology, with short water residence times….) can be viewed as one subset within a
suite of patterns, with each subset representing systems that differ in key physical
features important enough to disrupt ecosystem-scale patterns. In Figure 4, we
summarized the effect of land-derived nitrogen loads on the relative contribution to
estuarine primary production by 3 major components (seagrasses, macroalgae, and
phytoplankton) in estuaries with short or longer water residence times. In the upper
panel, we present data from systems with water residence times of duration similar to
that of the Waquoit estuaries (≤3 d). Seagrass production decreases sharply with
nitrogen load and is replaced by phytoplankton and macroalgal production. In the
lower panel, we present data from systems with significantly longer water residence
times (≥45 d). In these settings, seagrasses are also rapidly lost, but the percentage of
production carried out by phytoplankton and macroalgae follow a different pattern
from that described for systems of shorter water residence times. Presumably,
because phytoplankton has more time to grow, it is able to outcompete benthic
producers (via shading or nutrient competition) and dominate production across much
of the nitrogen loading range. In summary, we emphasize the importance of physical
factors in mediating the patterns observed with nitrogen loading, and suggest that we
may be most successful in observing patterns when we aggregate data within
physically similar systems.
Increasing evidence demonstrates at least 3 mechanisms (both direct and indirect) by
which increased nitrogen supply leads to seagrass decline: (1) direct nitrate toxicity
(Burkholder et al. 1992, 1994; Touchette et al. 2003), (2) light limitation (Short et al.
1993, Valiela et al. 1997b) via phytoplankton, epiphytes, and/or macroalgae
(Backman and Barilotti 1976, Harlin and Thorne-Miller 1981, Twilley et al. 1985,
Howard and Short 1986, Short et al. 1993, Duarte 1995, Short et al. 1995, Taylor et
al. 1995a and 1995b, Jernakoff et al. 1996, Moore et al. 1996, Kinney and Roman
1998), and (3) unfavourable alterations in the biogeochemical environment
surrounding seagrasses (van Katwijk et al. 1997, Greve et al. 2003) that may be
imposed by macroalgal canopies (Valiela et al. 1992, Bierzychudek et al. 1993,
D’Avanzo and Kremer 1994, Krause-Jensen et al. 1996, McGlathery et al. 1997).
5.1. Nitrate toxicity
Burkholder et al. (1992, 1994) measured decreased production by Zostera marina
(but increased production by Halodule wrightii and Ruppia maritima) in mesocosm
plots receiving 5-10 µM daily pulsed additions of nitrate. Touchette et al. (2003)
measured a 40% decrease in shoot density, and decreased leaf and root production by
Zostera marina in mesocosms receiving daily pulsed additions of nitrate (8 µM)
relative to controls.
Effects of nutrient loading
Figure. 4. Partition of total primary production in shallow estuaries into contributions by
phytoplankton, macroalgae, and seagrasses, in relation to measured annual land-derived
nitrogen loading rates (Valiela et al. 2000a) in estuaries of short (top panel) or longer (bottom
panel) water residence times. The dark shaded, light shaded and white areas represent the
percentage of total primary production contributed by seagrasses, macroalgae, and
phytoplankton, respectively. Adapted from Valiela et al. 2000b.
Because photosynthetically active radiation reaching eelgrass was not reduced in
nitrate-enriched mesocosms and epiphytes were routinely cleaned from leaves, the
discrepancy was attributed directly to nitrate variation. We are not aware of
published studies in which a link between nitrate toxicity and seagrass loss at the
estuary scale has been established, and further research in this area is warranted.
J. Hauxwell and I. Valiela
Nitrate concentrations in highly loaded systems can exceed critical levels
demonstrated in these experiments.
5.2. Light limitation
Seagrasses are generally light-limited, with intensity and photoperiod as important
factors. In numerous studies, when light has been experimentally reduced, seagrass
production decreased (Backman and Barilotti 1976, Dennison and Alberte 1985,
Longstaff et al. 1999, and many others). When light has been experimentally
augmented, seagrass production increased (Dennison and Alberte 1985). In
numerous cases, declines in seagrass coverage have been attributed to concomitant
increases in standing stocks of phytoplankton, epiphytes, and/or macroalgae.
Because these primary producers are also dynamic biological and chemical
components of systems, other processes must be considered as well, including their
ability to compete with seagrasses for nutrients. Below, we outline the evidence for
light limitation of seagrasses by the various types of algal producers, and attempt to
distinguish cases in which light limitation can be ruled as the probable mechanism.
5.2.1. Effect of phytoplankton
Calculating the contribution of phytoplankton to shading of seagrasses is relatively
straightforward, but it is important to recognize that phytoplankton is one component
of the entire water column affecting light penetration1. In a comparative study,
Krause-Jensen and Sand-Jensen (1998) found phytoplankton alone (minus associated
water column components) to be much less effective at reducing light than other taxa
of producers such as macrophytes and benthic microalgae. Since the relationship
between light intensity and growth has been experimentally established for several
species of seagrasses, it is possible to estimate reductions in growth rates, given depth
and background attenuation due to nonphotosynthetic particles, and water column
chlorophyll concentrations. For example, at a high extreme, standing stocks of
phytoplankton in a eutrophic estuary such as the Childs River, Waquoit Bay, MA,
may reduce up to 89% of light reaching the water surface during summer peaks (120
mg chlorophyll a m-3; Tomasky et al. 1999), which would decrease eelgrass (Zostera
marina L.) production by 80% (Short et al. 1995).
Proliferations of phytoplankton have been directly linked to seagrass loss. In the
Laguna Madre, Texas, a persistent brown tide resulted in a rapid loss of 2.6 km2 of
Halodule wrightii after 3.5 years (Onuf, 1996). In this case, light limitation as a
process by which phytoplankton affect seagrasses is corroborated by a similar
example in which sediment loading resulted in similar losses of seagrass habitat.
Beer’s Law can be used to determine the fraction of surface irradiance (Iz/Io) reaching depth
(z): Iz/Io = e-(KB+KP)(z), and the contribution to attenuation made by the various components of
the water column. The total water column light attenuation coefficient (KT , in units /m), is
equal to the sum of 1) the light extinction coefficient due to absorption by water and its
nonphytoplankton components (KB) and 2) the light extinction coefficient due to phytoplankton
(KP), which equals chlorophyll concentration (C, in units mg m-3) multiplied by the
chlorophyll-specific light attenuation coefficient (kc = 0.016 m2/mg chlorophyll; Bannister
1974, Kirk 1994).
Effects of nutrient loading
Light attenuation due to sediment resuspension (resulting from maintenance dredging
on the Gulf Intracostal Waterway) had same effect on Halodule in the lower Laguna
Madre, resulting in a 60% decrease between 1965 and 1988 (Onuf 1994, Pulich et al.
The relationship between depth limits for seagrasses and light availability, as
modified by phytoplankton or sediment, also supports light availability as a primary
factor by which phytoplankton may influence distributions (Duarte 1991). In a study
that included 27 Danish fjords, Nielsen et al. (2002a) demonstrated that (1) there was
a positive linear relationship between eelgrass depth limit and Secchi depth (which is
strongly dependent on phytoplankton biomass and decreased with nitrogen
concentration; Nielsen et al. 2002b), and (2) there was a negative relationship
between total water column nitrogen concentration and eelgrass depth limit (power
function after log-transformation). Similar relationships were observed between
seagrass depth limits in Moreton Bay, Australia, and water column light attenuation
coefficients (also related to chlorophyll a concentration) (Abal and Dennison 1996).
5.2.2. Effect of epiphytes
There are a number of ways increased epiphytes may negatively affect seagrasses
(Jernakoff et al. 1996). In high energy systems, increased epiphyte loads may
increase the drag on seagrasses, resulting in increased loss of leaves or whole shoots.
Epiphytes also create a boundary layer around the leaf, which may affect carbon and
nutrient uptake and oxygen diffusion. Architecturally, epiphytes may reduce the
amount of light reaching seagrass leaves. The most convincing evidence linking
epiphyte loads to decreased growth by seagrass plants comes from a handful of
experiments. Sand-Jensen (1977) demonstrated that epiphytes (at ambient loads in
Vellerup Vig, Denmark), reduced eelgrass photosynthesis by 31% at optimum light
conditions relative to leaves that had been cleaned. Howard and Short (1986)
documented a 22% decrease in Halodule wrightii leaf elongation in plots that
excluded epiphyte-consuming grazers.
The numerical relationship between epiphyte biomass and light attenuation is best
represented in Brush and Nixon (2002), where they directly measured light
transmission through unscraped and scraped seagrass blades across a wide range of
epiphyte densities, using natural plants and undisturbed epiphyte communities. They
found that attenuation of photosynthetically active radiation through epiphytes was
well-described by a negative hyperbolic equation, with variation in terms depending
on the type of epiphytic communities.2 Maximum epiphyte densities for marine
systems (as reviewed in Brush and Nixon (2002)), ranged between 2.1 and 99 mg cm2
Equations from Brush and Nixon (2002) are listed below for various types of epiphytic
communities (epiphyte biomass units in mg DW cm-2) (see Brush and Nixon 2002 for review of
all other published equations):
(1) for an epiphytic community dominated by Cladophora sp.: % transmission = 100 (89.1(biomass/(biomass + 17.3)),
(2) for an epiphytic community dominated by Polysiphonia sp.: % transmission = 100 (92.4(biomass/(biomass + 2.2)),
(3) for an epiphytic community dominated by an unidentified green filamentous (probably
Ulothrix sp.): % transmission = 100 - (95.7(biomass/(biomass + 3.3)).
J. Hauxwell and I. Valiela
of leaf, resulting in % attenuations approximately 10-45% and 76-90%, respectively
(depending on epiphyte architecture). These calculations demonstrate relatively
substantial losses of light at even relatively low epiphyte loads.
Though robust measurements of light attenuation are possible, the exercise of
applying them and understanding the effect of light limitation imposed by epiphytes
is more complex than that of phytoplankton or macroalgae, and inconsistent effects of
epiphytes on seagrass photosynthesis are represented in the literature (see review by
Jernakoff et al. 1996). The growth pattern of most seagrasses, in which old leaves are
sloughed as new leaves appear, is undoubtedly advantageous, in terms of minimizing
build-up of epiphytic material on photosynthetically-active surfaces (Borum 1987).
5.2.3. Effect of macroalgae
Accumulations of macroalgae may shade seagrasses. Peckol and Rivers (1996)
measured rapid attenuation of light through macroalgal canopies comprised of either
a fine green filamentous alga (Cladophora vagabunda) or a thicker red branched alga
(Gracilaria tikvahiae) (Figure 5, top). Regardless of algal taxon, the percentage of
light reaching the surface of the macroalgal canopy was reduced by > 95% after
penetrating only 6-8 cm into the canopy, and for both species, light attenuation was
best described by a negative logarithmic equation. A rapid decay of light was also
measured by McComb et al. (1981), in which 99% of light was attenuated within a 1cm thick algal mat.
Algal standing stocks are usually reported as biomass, so in Figure 5 (bottom), we
illustrate the relationship between biomass and canopy height for green filamentous
(top, Peckol and Rivers (1996)), red branched (middle, Peckol and Rivers (1996)),
and a green foliose (bottom, Coffaro and Bocci (1997)) algae. In all cases, a linear fit
best described the data, with variation in slopes resulting from different
morphologies. In Waquoit Bay, macroalgal canopy heights of 6-8 cm may be
sparsely distributed in estuaries with low nitrogen loads, and, under high nitrogen
loading conditions, evenly distributed macroalgal canopies of 75 cm may be attained
(Hauxwell et al. 2001).
Examples of standing stocks of this order and greater can be found around the
world. In Hog Island Bay, VA, USA, canopies comprised of Ulva sp., Gracilaria
tikvahiae, and Cladophora sp., have reached 650 g DW m-2 (Havens et al. 2001). In
Scotland, Coleman and Stewart (1979) reported standing stocks of Enteromorpha
prolifera that reached 1000 g DW m-2. In the Upper Newport Bay estuary, CA, USA,
Kamer et al. (2001) reported algal peaks >150 g DW m-2 (comprised of
Enteromorpha intestinalis, Ulva expansa, and Ceramium spp.). In Coos Bay, OR,
USA, biomass of Enteromorpha spp. and Ulva spp. have reached 750 g DW m-2
(Pregnall and Rudy (1985)), and in the Palmones River estuary, southern Spain,
standing stocks of Ulva spp. have reached 375 g DW m-2 (Hernández et al. 1997).
Effects of nutrient loading
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
J. Hauxwell and I. Valiela
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.
Effects of nutrient loading
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
J. Hauxwell and I. Valiela
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).
Effects of nutrient loading
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.
J. Hauxwell and I. Valiela
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.
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
Effects of nutrient loading
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).
J. Hauxwell and I. Valiela
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.
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.
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.
Effects of nutrient loading
Cohen, J. E., Small, C., Mellinger, A., Gallup, J., & Sachs, J. D. (1997). Estimates of coastal populations.
Science, 278, 1211.
Cole, J. J., Peierls, B. L., Caraco, N. F. , & Pace, M. L. (1993). Nitrogen loading of rivers and a humandriven process. In M. J. McDonnell & S. T. A. Pickett (Eds.), Humans as components of ecosystems:
the ecology of subtle human effects and populated areas (pp. 141-157). New York: Springer-Verlag.
Coleman, N. V., & Stewart, W. D. P. (1979). Enteromorpha prolifera in a polyeutrophic loch in Scotland.
British Phycological Journal, 14, 121.
Cowper, S.W. (1978). The drift algae community of seagrass beds in Redfish Bay, Texas. Contributions
in Marine Science, 21, 125-132.
D’Avanzo, C., & Kremer, J. N. (1994). Diel oxygen dynamics and anoxic events in an eutrophic estuary
of Waquoit Bay, Massachusetts. Estuaries, 17, 131-139.
Dean, T. A., Stekoll, M. S. , Jewett, S. C. , Smith, R. O. , & Hose, J. E. (1998). Eelgrass (Zostera marina)
in Prince William Sound, Alaska: effects of the Exxon Valdez oil spill. Marine Pollution Bulletin, 36,
den Hartog, C. (1970). The sea-grasses of the world. Amsterdam: North-Holland Publishing Company.
den Hartog, C. (1994). Suffocation of a littoral Zostera bed by Enteromorpha radiata. Aquatic Botany,
47, 21-28.
Dennison, W. C. (1987). Effects of light on seagrass photosynthesis, growth and depth distribution.
Aquatic Botany, 27, 15-26.
Dennison, W. C., & Alberte, R. S. (1982). Photosynthetic response of Zostera marina L. (eelgrass) to in
situ manipulations of light intensity. Oecologia, 55, 137-144.
Dennison, W. C., & Alberte, R. S. (1985). Role of daily light period in the depth distribution of Zostera
marina (eelgrass). Marine Ecology Progress Series, 25, 51-61.
Duarte, C. M. (1991). Seagrass depth limits. Aquatic Botany, 40, 363-377.
Duarte, C. M. (1995). Submerged aquatic vegetation in relation to different nutrient regimes. Ophelia, 41,
Duarte, C. M., & Cebrián, J. (1996). The fate of marine autotrophic production. Limnology and
Oceanography, 41, 1758-1766.
Duarte, C. M., & Chiscano, C. L. (1999). Seagrass biomass and production: a reassessment. Aquatic
Botany, 65, 159-174.
Duffy, K. C., & Baltz, D. M. (1998). Comparison of fish assemblages associated with native and exotic
submerged macrophytes in the Lake Ponchartrain estuary, USA. Journal of Experimental Marine
Biology and Ecology, 223, 199-221.
Fletcher, R. L. (1996). The occurrence of "green tides" - a review. In W. Schramm & P.H. Nienhuis
(Eds.), Marine benthic vegetation (pp. 7-43). Berlin: Springer-Verlag.
Frankovich, T. A., & Fourqurean, J. W. (1997). Seagrass epiphyte loads along a nutrient availability
gradient, Florida Bay, USA. Marine Ecology Progress Series, 159, 37-50.
GESAMP. (1990). State of the Marine Environment. Joint Group of Experts on the Scientific Aspects of
Marine Pollution. Reports and Studies Number 39. United Nations Environment Programme,
Gibbs, P. (1997). Botany Bay seagrass habitat restoration. Fisheries-NSW, 1, 17-18.
Goodman, J. L., Moore, K. A. , & Dennison, W. C. (1995). Photosynthetic responses of eelgrass (Zostera
marina L.) to light and sediment sulfide in a shallow barrier island lagoon. Aquatic Botany, 50, 37-47.
Greve, T. M., Borum, J. , & Pedersen, O. (2003). Meristematic oxygen variability in eelgrass (Zostera
marina). Limnology and Oceanography, 48, 210-216.
Hall, M. O., Durako, M. J., Fourqurean, J. W., & Zieman, J. C. (1999). Decadal changes in seagrass
distribution and abundance in Florida Bay. Estuaries, 22, 445-459.
Harlin, M. M., & Thorne-Miller, B. (1981). Nutrient enrichment of seagrass beds in a Rhode Island
Coastal Lagoon. Marine Biology, 65, 221-229.
Hauxwell, J., Cebrián, J. , Furlong, C., & Valiela, I. (2001). Macroalgal canopies contribute to eelgrass
(Zostera marina) decline in temperate estuarine ecosystems. Ecology, 82, 1007-1022.
Hauxwell, J., Cebrián, J., & Valiela, I. (2003). Eelgrass (Zostera marina L.) loss in temperate estuaries:
relationship to land-derived nitrogen loads and effect of light limitation imposed by algae. Marine
Ecology Progress Series, 247, 59-73.
Hauxwell, J., Osenberg, C. W., & Frazer, T. K. (2004). Conflicting management goals: manatees and
invasive competitors inhibit restoration of a native macrophyte. Ecological Applications, 14, 571-586.
Havens, K. E., Hauxwell, J., Tyler, A. C., Thomas, S., McGlathery, K. J., Cebrian, J., Valiela, I., Steinman,
A. D., & Hwang, S. J. (2001). Complex interactions between autotrophs in shallow marine and
freshwater ecosystems: implications for community responses to nutrient stress. Environmental
Pollution, 113, 95-107.
J. Hauxwell and I. Valiela
Hernández, I., Peralta, G., Pérez-Lloréns, J. L., Vergara, J. J., & Niell, F. X. (1997). Biomass and
dynamics of growth of Ulva species in Palmones River estuary. Journal of Phycology, 33, 764-772.
Holmer, M., & Bondgaard, E. J. (2001). Photosynthesis and growth response of eelgrass to low oxygen
and high sulfide concentrations during hypoxic events. Aquatic Botany, 70, 29-38.
Holmquist, J. G. (1997). Disturbance and gap formation in a marine benthic mosaic: influence of shifting
macroalgal patches on seagrass structure and mobile invertebrates. Marine Ecology Progress Series,
158, 121-130.
Howard , R. K., & Short, F. T. (1986). Seagrass growth and survivorship under the influence of epiphyte
grazers. Aquatic Botany, 24, 287-302.
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., Sharpley, A., & Walker, D. (2002). Sources of nutrient pollution to coastal waters in the
United States: implications for achieving coastal water quality goals. Estuaries, 25, 656-676.
Jaubert, J. M., Chisholm, J. R. M., Ducrot, D., Ripley, H. T., Roy, L., & Passeron-Seitre, G. (1999). No
deleterious alterations in Posidonia beds in the Bay of Menton (France) eight years after Caulerpa
taxifolia colonization. Journal of Phycology, 35, 1113-1119.
Jernakoff, P., Brearley, A., & Nielsen, J. (1996). Factors affecting grazer-epiphyte interactions in
temperate seagrass meadows. Oceanography and Marine Biology: An Annual Review, 34, 109-162.
Johansson, J. O. R., & Greening, H. S. (2000). Seagrass restoration in Tampa Bay: a resource-based
approach to estuarine management. In S. A. Bortone (ed.), Seagrasses: monitoring, ecology,
physiology, and management (pp. 279-293). New York: CRC Press.
Jupp, B. P. (1977). The effects of organic pollution on benthic organisms near Marseille. International
Journal of Environmental Studies, 10, 119-123.
Kaldy, J. E., & Dunton, K. H. (1999). Ontogenetic photosynthetic changes, dispersal and survival of
Thalassia testudinum (turtle grass) seedlings in a sub-tropical lagoon. Journal of Experimental Marine
Biology and Ecology, 240, 193-212.
Kaldy, J. E., Onuf, C. P., Eldridge, P. M., & Cifuentes, L. A. (2002). Carbon budget for a subtropical
seagrass dominated coastal lagoon: how important are seagrasses to total ecosystem net primary
production? Estuaries, 25, 528-539.
Kamer, K., Boyle, K. A., & Fong, P. (2001). Macroalgal bloom dynamics in a highly eutrophic southern
California estuary. Estuaries, 24, 622-634.
Kastler, T., & Michaelis, H. (1999). The decline of seagrasses, Zostera marina and Zostera noltii, in the
Wadden Sea of Lower Saxony. Senckenbergiana Maritima, 29, 77-80.
Kemp, W. M., R. R. Twilley, J. C. Stevenson, W. R. Boynton, & J. C. Means. (1983). The decline of
submerged vascular plants in upper Chesapeake Bay: summary of results concerning possible causes.
Marine Technology Society Journal, 17, 78-89.
Kendrick, G. A., Eckersley, J., & Walker, D. I. (1999). Landscape-scale changes in seagrass distribution
over time: a case study from Success Bank, Western Australia. Aquatic Botany, 65, 293-309.
Kinney, E. H., & Roman, C. T. (1998). Response of primary producers to nutrient enrichment in a
shallow estuary. Marine Ecology Progress Series, 163, 89-98.
Kirk, J. T. O. (1994). Light and photosynthesis in aquatic ecosystems. Cambridge: Cambridge University
Koch, E. W., & Beer, S. (1996). Tides, light and the distribution of Zostera marina in Long Island Sound,
USA. Aquatic Botany, 53, 97-107.
Koch, M.S., & Erskine, J.M. (2001). Sulfide as a phytotoxin to the tropical seagrass Thalassia testudinum:
interactions with light, salinity, and temperature. Journal of Experimental Marine Biology and
Ecology, 266, 81-95.
Koch, M. S., Mendelssohn, I. A., & McKee, K. L. (1990). Mechanism for the hydrogen sulfide-induced
growth limitation in wetland macrophytes. Limnology and Oceanography, 35, 399-408.
Komatsu, T. (1997). Long-term changes in the Zostera bed area in the Seto Inland Sea (Japan) especially
along the coast of the Okayama Prefecture. Oceanologica Acta, 20, 209-216.
Krause-Jensen, D., McGlathery, K., Rysgaard, S., & Christensen, P. B. (1996). Production within dense
mats of Chaetomorpha linum in relation to light and nutrient availability. Marine Ecology Progress
Series, 134, 207-216.
Krause-Jensen, D., & Sand-Jensen, K. (1998). Light attenuation and photosynthesis of aquatic plant
communities. Limnology and Oceanography, 43, 396-407.
Lapointe, B. E., & Clark, M. W. (1992). Nutrient inputs from the watershed and coastal eutrophication in
the Florida Keys. Estuaries, 15, 465-476.
Lapointe, B. E., Littler, M. M., & Littler, D. S. (1992). Nutrient availability to marine macroalgae in
siliciclastic versus carbonate-rich coastal waters. Estuaries, 15, 75-82.
Effects of nutrient loading
Larned, S. T. (2003). Effects of the invasive, nonindigenous seagrass Zostera japonica on nutrient fluxes
between the water column and benthos in a NE Pacific estuary. Marine Ecology Progress Series, 254,
Lathrop, R. G., Styles, R. M., Seitzinger, S. P., & Bognar, J. A. (2001). Use of GIS mapping and
modeling approaches to examine the spatial distribution of seagrasses in Barnegat Bay, New Jersey.
Estuaries, 24, 904-916.
Lin, H. J., Nixon, S. W., Taylor, D. I., Granger, S. L., & Buckley, B. A. (1996). Responses of epiphytes
on eelgrass, Zostera marina L., to separate and combined nitrogen and phosphorus enrichment.
Aquatic Botany, 52, 243-258.
Long, B., Skewes, T., Taranto, T., Thomas, M., Isdale, P., Pitcher, R., & Poiner, I. (1997). Seagrass
dieback in north western Torres Strait. CSIRO MR-GIS-97/6, CSIRO Marine Research, Hobart,
Longstaff, B. J., Loneragan, N. R., O'Donohue, M. J., & Dennison, W. C. (1999). Effects of light
deprivation on the survival and recovery of the seagrass Halophila ovalis (R.Br.) Hook. Journal of
Experimental Marine Biology and Ecology, 234, 1-27.
Maciá, S. (2000). The effects of sea urchin grazing and drift algal blooms on a subtropical seagrass bed
community. Journal of Experimental Marine Biology and Ecology, 246, 53-67.
McComb, A. J., Arkins, R. P., Birch, P. B., Gordon, D. M., & Lukarelich, R. J. (1981). Eutrophication in
the Peel-Harvey estuarine system, western Australia. In B. J. Neilson & L. E. Cronin (Eds.), Estuaries
and nutrients (pp. 323-341). Clifton, NJ: Humana Press.
McCook, L. J., Jompa, J., & Diaz-Pulido, G. (2001). Competition between corals and algae on coral reefs:
a review of evidence and mechanisms. Coral Reefs, 19, 400-417.
McGlathery, K. J., Krause-Jensen, D., Rysgaard, S., & Christensen, P. B. (1997). Patterns of ammonium
uptake within dense mats of the filamentous macroalga Chaetomorpha linum. Aquatic Botany, 59, 99115.
Meehan, A. J., & West, R. J. (2000). Recovery times for a damaged Posidonia australis bed in south
eastern Australia. Aquatic Botany, 67, 161-167.
Meinesz, A. (1999). Killer algae. Chicago: University of Chicago Press.
Meinesz, A., de Vaugelas, J., Hesse, B. , & Mari, X. (1993). Spread of the introduced tropical green alga
Caulerpa taxifolia in northern Mediterranean waters. Journal of Applied Phycology, 5, 141-147.
Miller, M. W., Hay, M. E., Miller, S. L., Malone, D., Sotka, E. E., & Szmant, A. M. (1999). Effects of
nutrients versus herbivores on reef algae: a new method for manipulating nutrients on coral reefs.
Limnology and Oceanography, 44, 1847-1861.
Montcreiff, C. A., Sullivan, M. J., & Daehnick, A. E. (1992). Primary production dynamics in seagrass
beds of Mississippi Sound: the contributions of seagrass, epiphytic algae, sand microflora, and
phytoplankton. Marine Ecology Progress Series, 87, 161-171.
Moore, K. A., Neckles, H. A., & Moore, R. J. (1996). Zostera marina (eelgrass) growth and survival
along a gradient of nutrients and turbidity in the lower Chesapeake Bay. Marine Ecology Progress
Series, 142, 247-259.
Moore, K. A., & Wetzel, R. L. (2000). Seasonal variations in eelgrass (Zostera marina L.) response to
nutrient enrichment and reduced light availability in experimental ecosystems. Journal of
Experimental Marine Biology and Ecology, 244, 1-28.
National Research Council. (1994). Priorities for Coastal Science. National Academy Press. Washington,
Neckles, H. A., Koepfler, E. T., Haas, L. W., Wetzel, R. L., & Orth, R. J. (1994). Dynamics of epiphytic
photoautotrophs and heterotrophs in Zostera marina (eelgrass) microcosms: responses to nutrient
enrichment and grazing. Estuaries, 17, 597-605.
Nelson, T. A., & Lee, A. (2001). A manipulative experiment demonstrates that blooms of the macroalga
Ulvaria obscura can reduce eelgrass shoot density. Aquatic Botany, 71, 149-154.
Nielsen, S. L., Borum, J., Geertz-Hansen, O., & Sand-Jensen, K. (1989). Marine bundplanters
dybdegraense. Vand og Miljø, 5, 217-220.
Nielsen, S. L., Sand-Jensen, K., Borum, J., & Geertz-Hansen, O. (2002a). Depth colonization of eelgrass
(Zostera marina) and macroalgae as determined by water transparency in Danish coastal waters.
Estuaries, 25, 1025-1032.
Nielsen, S. L., Sand-Jensen, K., Borum, J., & Geertz-Hansen, O. (2002b). Phytoplankton, nutrients, and
transparency in Danish coastal waters. Estuaries, 25, 930-937.
Nixon, S. W. (1995). Coastal marine eutrophication: a definition, social causes, and future concerns.
Ophelia, 41, 199-219.
Nixon, S., Buckley, B., Granger, S., & Bintz, J. (2001). Responses of very shallow marine ecosystems to
nutrient enrichment. Human and Ecological Risk Assessment, 7, 1457-1481.
J. Hauxwell and I. Valiela
Onuf, C. P. (1994). Seagrasses, dredging and light in Laguna Madre, Texas, U.S.A. Estuarine, Coastal
and Shelf Science, 39, 75-91
Onuf, C. P. (1996). Seagrass responses to long-term light reduction by brown tide in upper Laguna
Madre, Texas: distribution and biomass patterns. Marine Ecology Progress Series, 138, 219-231.
Orth, R. J., Fishman, J. R., Wilcox, D. J., & Moore, K. A. (2002). Identification and management of
fishing gear impacts in a recovering seagrass system in the coastal bays of the Delmarva Peninsula,
USA. Journal of Coastal Research, 37, 111-129.
Pasquilini, V., Pergent-Martini, C., & Pergent, G. (1999). Environmental impact identification along the
Corsican coast (Mediterranean Sea) using image processing. Aquatic Botany, 65, 311-320.
Peckol, P., DeMeo-Anderson, B., Rivers, J., Valiela, I., Maldonado, M., & Yates, J. (1994). Growth,
nutrient uptake capacities and tissue constituents of the macroalgae Cladophora vagabunda and
Gracilaria tikvahiae related to site-specific nitrogen loading rates. Marine Biology, 121, 175-185.
Peckol, P., & Rivers, J. S. (1996). Contribution by macroalgal mats to primary production of a shallow
embayment under high and low nitrogen-loading rates. Estuarine, Coastal and Shelf Science, 43, 311325.
Pedersen, O., Borum, J., Duarte, C. M. & Fortes, M. D. (1998). Oxygen dynamics in the rhizosphere of
Cymodocea rotundata. Marine Ecology Progress Series, 169, 283-288.
Peterken, C. J., & Conacher, C. A. (1997). Seed germination and recolonization of Zostera capricorni
after grazing by dugongs. Aquatic Botany, 59, 333-340.
Pickett, S. T. A. (1989). Space-for-time substitution as an alternative to long-term studies. In G. E.
Likens (Ed.), Long-term studies in ecology: approaches and alternatives (pp. 110-135). New York:
Pregnall, A. M., & Rudy, P. P. (1985). Contribution of green macroalgal mats (Enteromorpha spp.) to
seasonal production in an estuary. Marine Ecology Progress Series, 24, 167-176.
Pregnall, A. M., Smith, R. D., Kursar, T. A., & Alberte, R. S. (1984). Metabolic adaptations of Zostera
marina (eelgrass) to diurnal periods of root anoxia. Marine Biology, 83, 141-147.
Pulich, Jr., W., Blair, C., & White, W. A. (1997). Current status and historical trends of seagrasses in the
Corpus Christi Bay National Estuary Program Study Areas. Publication number CCBNEP - 20. Texas
Natural Resource Conservation Commission, TX, USA. pp. 1-57.
Reusch, T. H. B., & Williams, S. L. (1999). Macrophyte canopy structure and the success of an invasive
marine bivalve. Oikos, 84. 398-416.
Robbins, B. D. (1997). Quantifying temporal change in seagrass areal coverage: the use of GIS and low
resolution aerial photography. Aquatic Botany, 58, 259-267.
Robertson, A. I., & Mann, K. H. (1984). Disturbance by ice and life-history adaptations of the seagrass
Zostera marina. Marine Biology, 80, 131-141.
Rose, C. D., Sharp, W. C., Kenworthy, W. J., Hunt, J. H., Lyons, W. G., Prager, E. J., Valentine, J. F.,
Hall, M. O., Whitfield, P. E., & Fourqurean, J. W. (1999). Overgrazing of a large seagrass bed by the
sea urchin Lytechinus variegatus in Outer Florida Bay. Marine Ecology Progress Series, 190, 211222.
Ruiz, J. M., Perez, M., & Romero, J. (2001). Effects of fish farm loadings on seagrass (Posidonia
oceanica) distribution, growth and photosynthesis. Marine Pollution Bulletin, 42, 749-760.
Rybicki, N. B., & Carter, V. (1986). Effect of sediment depth and sediment type on the survival of
Vallisneria americana Michx grown from tubers. Aquatic Botany, 24, 233-240.
Ryther, J. H. (1969). Photosynthesis and fish production in the sea. Science, 166, 72-76.
Sand-Jensen, K. (1977). Effects of epiphytes on eelgrass photosynthesis. Aquatic Botany, 3, 55-63.
Sand-Jensen, K., & Borum, J. (1991). Interactions among phytoplankton, periphyton, and macrophytes in
temperate freshwaters and estuaries. Aquatic Botany, 41, 137-175.
Seddon, S., Connolly, R. M., & Edyvane, K. S. (2000). Large-scale seagrass dieback in northern Spencer
Gulf, South Australia. Aquatic Botany, 66, 297-310.
Seitzinger, S. P., Kroeze, C., Bouwman, A. F., Caraco, N., Dentener, F., & Styles, R. V. (2002). Global
patterns of dissolved inorganic and particulate nitrogen inputs to coastal systems: recent conditions
and future projections. Estuaries, 25, 640-655.
Sfriso, A. (1987). Flora and vertical distribution of macroalgae in the lagoon of Venice: a comparison
with previous studies. Giornale Botanico Italiano, 121, 69-85.
Sfriso, A., Pavoni, B., Marcomini, A., & Orio, A. A. (1992). Macroalgae, nutrient cycles, and pollutants
in the lagoon of Venice. Estuaries, 15, 517-528.
Short, F. T. (1988). Effects of sediment nutrients on seagrasses: literature review and mesocosm
experiment. Aquatic Botany, 27, 41-57.
Short, F. T., Burdick, D. M., & Kaldy III, J. E. (1995). Mesocosm experiments quantify the effects of
eutrophication on eelgrass, Zostera marina. Limnology and Oceanography, 40, 740-749.
Effects of nutrient loading
Short, F. T., Burdick, D. M., Wolf, J. S., & Jones, G. E. (1993). Eelgrass in estuarine research reserves
along the East Coast, U.S.A., Part I: Declines from pollution and disease and Part II: Management of
eelgrass meadows. National Oceanic and Atmospheric Administration, Coastal Ocean Program
Publication, Rockville, Maryland. Pp. 1-83, M1-M24.
Short, F. T, & Wyllie-Echeverria, S. (1996). Natural and human-induced disturbance of seagrasses.
Environmental Conservation, 23, 17-27.
Sigua, G. C., & Tweedale, W. A. (2003). Watershed scale assessment of nitrogen and phosphorus
loadings in the Indian River Lagoon basin, Florida. Journal of Environmental Management, 67, 363372.
Silberstein, K., Chiffings, A. W., & McComb, A. J. (1986). The loss of seagrass in Cockburn Sound,
Western Australia. III. The effect of epiphytes on productivity of Posidonia australis Hook. F.
Aquatic Botany, 24, 355-371.
Smith, A. (1997). Seagrass protected in NSW estuaries. Fisheries-NSW, 1, 19-20.
Smith, R. D., Dennison, W. C., & Alberte, R. S. (1984). Role of seagrass photosynthesis in root aerobic
processes. Plant Physiology, 74, 1055-1058.
Smith, S. V. (1981). Marine macrophytes as a global carbon sink. Science, 211, 838-840.
Street, G. T., Montagna, P. A., & Parker, P. L. (1997). Incorporation of brown tide into an estuarine food
web. Marine Ecology Progress Series, 152, 67-78.
Taylor, D., Nixon, S., Granger, S., & Buckley, B. (1995a). Nutrient limitation and the eutrophication of
coastal lagoons. Marine Ecology Progress Series, 127, 235-344.
Taylor, D. I., Nixon, S. W., Granger, S. L., Buckley, B. A., McMahon, J. P., & Lin, H.-J. (1995b).
Responses of coastal lagoon plant communities to different forms of nutrient enrichment - a mesocosm
experiment. Aquatic Botany, 52, 19-34.
Tomasko, D. A., Bristol, D. L., & Ott, J. A. (2001). Assessment of present and future nitrogen loads,
water quality, and seagrass (Thalassia testudinum) depth distribution in Lemon Bay, Florida.
Estuaries, 24, 926-938.
Tomasky, G., Barak, J., Valiela, I., Behr, P., Soucy, L., & Foreman, K. (1999). Nutrient limitation of
phytoplankton growth in Waquoit Bay, Massachussetts, USA: a nutrient enrichment study. Aquatic
Ecology, 33, 147-155.
Touchette, B. W., Burkholder, J. M., & Glasgow, Jr, H. B. (2003). Variations in eelgrass (Zostera marina)
morphology and internal nutrient composition as influenced by increased temperature and water
column nitrate. Estuaries, 26, 142-155.
Twilley, R. R., Kemp, W. M., Stave, K. W., Stevenson, J. C., & Boynton, W. R. (1985). Nutrient
enrichment of estuarine communities. 1. Algal growth and effects on production of plants and
associated communities. Marine Ecology Progress Series, 23, 179-191.
U.S. Geological Survey. (1999). The quality of our nation’s waters - nutrients and pesticides. U.S.
Geological Survey Circular 1225. Pp. 1-82.
Valiela, I., & Cole, M. L. (2002). Comparative evidence that salt marshes and mangroves may protect
seagrass meadows from land-derived nitrogen loads. Ecosystems, 5, 92-102.
Valiela, I., Collins, G., Kremer, J., Lathja, K., Geist, M., Seely, B., Brawley, J., & Sham, C.H. (1997a).
Nitrogen loading from coastal watersheds to receiving estuaries: new methods and application.
Ecological Applications, 7, 358-380.
Valiela, I., Foreman, K., LaMontagne, M., Hersh, D., Costa, J., Peckol, P., DeMeo-Anderson, B.,
D'Avanzo, C., Babione, M., Sham, C., Brawley, J., & Lathja, K. (1992). Couplings of watersheds and
coastal waters: sources and consequences of nutrient enrichment in Waquoit Bay, Massachusetts.
Estuaries, 15, 443-457.
Valiela, I., Geist, M., McClelland, J., & Tomasky, G. (2000a). Nitrogen loading from watersheds to
estuaries: verification of Waquoit Bay Nitrogen Loading Model. Biogeochemistry, 49, 277-293.
Valiela, I., McClelland, J., Hauxwell, J., Behr, P. J., Hersh, D., & Foreman, K. (1997b). Macroalgal
blooms in shallow estuaries: Controls and ecophysiological and ecosystem consequences. Limnology
and Oceanography, 42, 1105-1118.
Valiela, I., Tomasky, G., Hauxwell, J., Cole, M. L., Cebrián, J., & Kroeger, K. D. (2000b).
Operationalizing sustainability: management and risk assessment of land-derived nitrogen loads to
estuaries. Ecological Applications, 10, 1006-1023.
van Katwijk, M. M., Vergeer, L. H. T., Schmitz, G. H. W., & Roelofs, J. G. M. (1997). Ammonium
toxicity in eelgrass Zostera marina. Marine Ecology Progress Series, 157, 159-173.
Vasiliu. F. (1996). The Black Sea. In W. Schramm & P. H. Nienhuis (Eds.), Marine benthic vegetation
(pp. 435-447). Berlin: Springer-Verlag.
J. Hauxwell and I. Valiela
Vitousek, P. M., Aber, J., Howarth, R. W., Likens, G. E., Matson, P. A., Schindler, D. W., Schlesinger, W.
H., & Tilman, G. D. (1997). Human alteration of the global nitrogen cycle: causes and consequences.
Ecological Applications, 7, 737-750
Williams, S. J., Dodd, D., & Gohm, K. K. (1991). Coasts in crisis. United States Geological Survey
Circulation 1075.
Williams, S. L., & Grosholz, E. D. (2002). Preliminary reports from the Caulerpa taxifolia invasion in
southern California. Marine Ecology Progress Series, 233, 307-310.
Zaitsev, Y. P. (1992). Recent changes in the trophic structure of the Black Sea. Fisheries Oceanography,
1, 180-189.
Zieman, J. C., Fourqurean, J. W., & Frankovich, T. A. (1999). Seagrass die-off in Florida Bay: long-term
trends in abundance and growth of turtle grass, Thalassia testudinum. Estuaries, 22, 460-470.
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.
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.
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.
M.R. Flindt et al.
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.
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
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, P<0.01) and Chaetomorpha
M.R. Flindt et al.
aerea. (r2= 0.82, P<0.01), while the transport of Zostera spp. (r2=0.082) was not
significantly related to the water current.
Figure 2. Correlation between current velocity and catchment of dominating plant material
Venice Lagoon, Malamocco site near the inlet, May 1995. (Flindt et al. 1997a, reprinted with
permission from Elsevier)
The correlation between drift of macroalgae and the current velocity indicated that
advective transport was a major mechanism for plant loss from the studied area. The
advective transport of rooted macrophytes like Zostera sp. did not show any clear
correlation to current velocity. This is most likely because the sloughed leaves from
Zostera, in contrast to the macroalgae, are floating at the water surface. Here the drift
direction is influenced by the wind. It may also be expected that the sloughing of
leaves from rooted macrophytes are not only controlled by current velocities but are
also dependent on the age of the leaves, the production, the physiological condition of
the plant, and various environmental conditions like anoxia and wind or wave created
Many field campaigns in different estuaries have been carried out over full
tidal cycles, so complete import-export transport patterns and net balances could be
established. Such a campaign is shown on Figure 3, where measurements were made
close to the Lido entrance in the Lagoon of Venice. The station was 1.5 meter deep.
The measurements started on a spring tide with an outgoing current where the tidal
amplitude was about 80 cm. As the current velocity increased, the plant transport
Plant bound nutrient transport
started, and the earlier established correlation between current speed and plant catch
was evident, because the highest current speed occurred between 2-4 h.
Plant transport (g dwt m -2)
red algae
rooted veg.
Time (hours)
Figure 3. An example of plant transport over a full tidal cycle. Venice Lagoon, Bacan site,
August 1998.
This example also underlines the general tendency in most of the campaigns, that the
loss of plant material across the boundary is very high. No more than 10-15% of the
exported plant matter returned with the subsequent flooding tide where the water
returned to the lagoon. Only 10% of the transported Ulva sp. returned, about 4% of
Chaetomorpha sp. and mixed red algae (Gracilaria sp, Ceramium sp., Polysiphonia
sp.) returned, while close to 25% of the leaves from rooted plants (Zostera marina,
Zostera noltii, Cymodocia sp.) returned to the lagoon. These differences can be
explained by the differences in settling rates among these macrophytes. In conclusion,
about 90 % of the plant mass was permanently lost to the coastal area – in this
example the Adriatic Sea. Although there may be variations in the import-export
balances of plant transport, the above discussed example may be considered as
representative. The variations in the import-export balances are primarily dependent
The bathymetric conditions around the outer boundary. If the transport is
from a shallow area to a deeper coastal area almost all transported plant
matter is lost. This is the case in Venice Lagoon, which has an average depth
of about 1 m, while the depth of the Adriatic Sea is about 20-50 m.
The current patterns outside the boundary, where the water may be more or
less stagnant – depending on local wind and coastal current dynamics.
M.R. Flindt et al.
The dominating kind of vegetation in the area. If the system is dominated by
rooted vegetation relatively more plant matter returns to the system.
To produce correct mass balances and to develop predictive models of macrophyte
advection and transport in estuaries, it is important to take the plant bound export and
dispersion of plant material into account. To do this, we need to know the threshold
for macroalgae mobilization, the mode of transport as a function of current velocity,
and the settling rates of the different dominating plant material.
The Southampton Oceanographic Lab Carousel (Amos et al 2000, Thomson & Amos
2002) was used in an experimental study on erosion and transport of macroalgal
material. The carousel was filled with prefiltered estuarine water and a 4 cm thick
sediment bed was created. Four species of macroalgae (Ulva lactuca, Cladophora sp.,
Chaetomorpha linum and Ceramium rubrum) were collected from the same site as
the sediment. Epiphytes and inorganic particulate matter were removed from the
surface of the algae to eliminate possible differences in densities and surface to
volume ratio between the specimens. Afterwards, the macroalgae were placed
unattached on the sediment surface. Air bubbles were removed. Mean current speed
was increased in steps of 0.5 cm s-1 at 10 minutes intervals to 11.0 cm s-1. The current
speed at the point where the plants started to move was termed the traction threshold.
In all experiments the plants started to move as bedload after which they
lifted into suspension at higher current velocities. The current speed at which the
plants lifted into suspension was the suspension threshold. Two pieces of Ulva sp.,
differing in weight by a factor of 10, were monitored at increasing current speeds (to
9 cm s-1). The smallest piece had a traction threshold at a current speed of 2.7 cm s-1
and a suspension threshold of 7.0 cm s-1. The largest piece of Ulva sp. had a slightly
lower traction threshold of 2.0 cm s-1, the speed of this algae increased with the
current speed in a similar fashion (Fig. 4A). However, the differences in threshold
velocity and transport rate between the two sizes were not significant. The traction
threshold for Chaetomorpha sp. (Fig. 4B) varied between 1.7 and 3.0 cm s-1 and was
thus similar to Ulva. None of the three Chaetomorpha pieces passed into suspension
but continued to tumble along the bed at all current speeds. All three data sets fitted a
logarithmic curve (r2= 0.963 to 0.986). The filamentous red algae Ceramium (Fig.
4C) showed a transport pattern and traction thresholds from 1.7 to 3.0 cm s-1. Two
pieces of Ceramium showed a suspension threshold of 9 cm s-1. The three data sets
also fitted well to a logarithmic curve (r2= 0.934 to 0.986). The three pieces of similar
size of Cladophora (Fig. 4D) started to move along the bottom at a current speed
between 1.7 and 3.0 cm s-1 similar to the red algae Ceramium. The first piece of
Cladophora came into suspension at 5.0 cm s-1, the other two at 9 cm s-1. All three
data sets fitted well a logarithmic curve (r2= 0.947 to 0.992).
When these experiments were initiated we expected a broad range of traction
thresholds. The hypothesis was that different morphological types of macroalgae:
sheet-formed (eg. Ulva sp.), unbranched filamentous (eg. Chaetomorpha sp. and
branched filamentous (eg. Ceramium sp.) would need hydraulically different traction
due to their differences in surface area. The results did not support these expectations.
Plant bound nutrient transport
Algae speed (m s )
A. Ulva lactuca
B. Chaetomorpha sp.
Algae speed (m s )
C. Ceramium sp.
D. Cladophora sp.
Current speed (m s )
Current speed (m s )
Figure 4. Correlations between algae speed and current speed for 4 algae species. (A) Ulva
lactuca. Correlations are shown for two pieces of algae differing in weight by a factor of 10 (r2
= 0.993). (B) Chaetomorpha sp. (r2 ≥ 0.963, n=3). (C) Ceramium sp. (r2 ≥ 0.934, n=3). (D)
Cladophora sp. (r2 ≥ 0.947, n=3).
A summary of the erosion experiments is presented in Figure 5. All species of
macroalgae began to move at very low current speeds (2 to 4 cm s-1). Once in motion,
the algae moved slowly along the sediment bed as traction. With increasing current
speed, the algae began to tumble along the bed and at a current speed from 6 to 10 cm
s-1 three of the algae species were in suspension, only Chaetomorpha sp. did not go
into suspension during the experiment. When in suspension the velocity of the algae
was the same as the current speed.
A major fraction of submerged estuarine plants is found either unattached at the bed
as for example Ulva sp. and Chaetomorpha sp. or loosely attached as Enteromorpha
sp., Ceramium sp. and Polysiphonia sp. All these types of algae are observed drifting
in the water column or moving as bedload transport within Venice Lagoon and at The
Solent estuary near Southampton. When these drifting macroalgae were observed at
new areas they were earlier mistakenly considered as algae produced at the site. The
plants release oxygen from their tissue, which sometimes becomes trapped as
attached bubbles thereby changing the density of the macroalgae, causing them to
float. Buoyancy and lift is thus responsible for part of the erosion of macroalgae.
M.R. Flindt et al.
Current speed (m s-1)
bed load
Figure 5. Transport pattern (bed load or suspended) for four algae species: Ulva,
Chaetomorpha, Ceramium, and Cladophora. Results from all carousel experiments (see text).
Error lines indicate standard deviation of mean.
The present experiments were undertaken without associated oxygen bubbles.
Nevertheless, the macroalgae started moving at very low current speeds and at 9 cm s1
all, regardless size and weight, moved into suspension. It thus appears that the
unattached macroalgae are highly mobile and respond to even low steady flows. This
fact can have a great impact on how we interpret the biology of unattached
macroalgae in shallow estuaries, especially in estuaries influenced by tide or
meteorological forcing. Furthermore, the plant transport can play an important role in
the nutrient balance of these estuaries through the export and/or import of large
amounts of plant material by tides.
In Venice Lagoon a significant correlation was observed between current
velocity and the catch of Ulva rigida and Chaetomorpha aeraea. By extrapolating the
regression lines to zero catch, it was possible to find the current velocity for the
erosion threshold. The field erosion thresholds were 16 cm s-1 for Ulva rigida and 6
cm s-1 for Chaetomorpha aeraea were a little above the measured traction thresholds
of the laboratory study, where both macroalgae started to move at 2-4 cm s-1. The
difference in threshold velocity could be caused by differences in bed roughness in
the field and laboratory. No doubt that bed roughness creates variations in threshold.
The sediment in Venice was silty sand, while the laboratory bed was muddy silt. In
nature, macrophytes tumbling along the bottom initially may be trapped by pools,
stones or in the rooted vegetation when the current velocities are low. The differences
may also be due to more imprecise measurement in the field.
Plant bound nutrient transport
It is obvious that unattached macroalgae may be transported by the currents, but a
major part of macroalgae grows fixed to shells or stones. In this section we describe
results for attached macroalgae like Enteromorpha intestinalis, Porphyra sp., Ulva sp.
and Fucus sp. on sloughing rates and later these are compared with other loss rates
like sporulation and grazing.
Sloughing rates were measured at controlled conditions in a laboratory
carousel, where current velocities could be varied. Ten specimens of each macroalgal
species attached to shells or stones were placed evenly distributed in the lab-carousel.
The current speed was increased in steps from still water until the macroalgae
including the attachment started to move. Between each speed increment the
sloughed macroalgae material was collected manually. The biomasses of the sloughed
macroalgal material as well as the remaining attached parts left at the end of each
experiment were measured. Figure 6 shows the results from the sloughing experiment
for four species of macroalgae.
The first attached macroalgal species to loose biomass were Porphyra and
Ulva which at a current speed of 12 cm s-1 had lost 3-4% of their biomass, increasing
to 4-7% at 15 cm s-1. At a current speed of 40 cm s-1 the loss was 40 - 50%. The
sloughing of Enteromorpha was slightly less affected by the current speed. For this
macroalgae the sloughing was 1% of the biomass at 20 cm s-1, increasing to 20% at
35 cm s-1. Above this speed Enteromorpha and shells started to move. The most
resistant species to sloughing was Fucus that started to lose biomass at a speed of 30
cm s-1, increasing to 1.5% at a speed of 35 cm s-1. The maximum loss measured on
Fucus was 4% at a relatively high current speed of 50 cm s-1. Furthermore, Fucus was
attached to large stones that did not move with the highest current speeds.
Linear regression between current speed and accumulated sloughing rates were made
for each species and presented in Table 1. The statistics show that macroalgae
sloughing rates have a high dependency on current speed and that the differences in
thresholds are much bigger than the measured variation in erosion thresholds for
unattached macroalgae.
Table 1. Sloughing threshold and linear regression between sloughing and current speed for
four species of macroalgae
Porphyra sp.
Threshold (cm
Ulva sp.
Enteromorpha sp.
Fucus sp.
y = 0.0128x –
y = 0.0150x –
y = 0.0089x –
y = 0.0016x –
M.R. Flindt et al.
Accumulated sloughing
(% of biomass)
Current speed (cm s )
Figure 6. Relationship between accumulated sloughed macroalgae biomass (% of total
biomass) and current speed. The study was made in the laboratory at Southampton
Oceanography Centre, May 2001
Assuming that the force created by the current on the macroalgae increases
linearly with current speed, it becomes possible to predict when the attached species
will start to lose mass, and at which current speed all macroalgal matter becomes lost
(Table 2). Compared to the erosion threshold for unattached macroalgae, the loss by
sloughing requires a much higher current speed making the attached macroalgae less
sensitive to transport losses. However, the ephemeral macroalgae started to loose
biomass at current speeds of 10-15 cm s-1 and they had a considerable loss at 30-50
cm s-1. For Fucus sp. and comparable macroalgae with more firm structural mass and
a stronger holdfast, the sloughing rate will never become a major loss rate, because an
unrealistic high current speed of 655 cm s-1 would be required to make loss rates by
sloughing important.
Table 2. Current speeds for beginning and maximum sloughing rates for four macroalgae
Porphyra sp.
Ulva sp.
Enteromorpha sp.
Fucus sp.
Current speed (cm s-1) where sloughing rate
is maximal
Plant bound nutrient transport
These results are based on conservative assumptions because all
measurements were carried out on healthy plants without ghost tissue, which may not
be representative for an average macroalgal community. The experiments were also
carried out so no thalli contained air bubbles.
There may be many explanations for the presence or absence of macroalgae
in different areas at certain times of the growth season, but the above results suggest
that the presence of e.g. Enteromorpha sp. in the intertidal becomes limited to period
where they can realize nearly maximum growth rates. In sub optimal periods, the
biomass loss due to sloughing will result in decreases of biomass and extinction. This
seems the case in the Mondego Estuary where Enteromorpha is absent in the channel
with the highest current velocities of 1 m s-1, but once in a while is able to manage to
grow in the shallow hydrodynamically protected intertidal areas.
The experimental studies on sloughing were verified by qualitative field
studies at different tidal currents. Here we observed that Ulva, Enteromorpha and
Porphyra were transported in the estuary attached to shells and small stones, while
we never encountered Fucus sp. transported. Enteromorpha sp. was the algae that had
the highest transportation frequency as attached and alive. This may be explained as a
combined morphological and physiological phenomenon where photosynthetic
produced oxygen easily becomes trapped in the tube-formed thallus. When this
happens, the buoyancy of the algae increases resulting in the ability to compensate for
the extra mass the shells or small stones exert. Thus, much lower current velocities
are needed to bring the plant into the water column. Fucus, attached to stones, may
only be transported at very high current speeds. This occurs as bed load transport
which is difficult to observe in turbid estuaries.
To model transportation of macrophytes, both a parameterization of the erosion
process as well as the settling process are necessary. This section describes the
diversity among different morphological types of macrophytes with respect to settling
The dominant macrophytes living in Venice Lagoon were collected for
settling studies in the laboratory. Epiphytes and sediment particles were gently
removed by washing in estuarine water. All settling rates were measured on living
plant material except the leaves from the rooted vegetation, which only settle when
the aerenchym is filled with water. This occurs at senescence, and therefore dead
leaves were used for settling measurements on rooted plants. The settling rate was
measured for various sizes (1 – 10 cm2) of circular and square discs of Ulva lactuca.
The settling experiments were carried out in a plexiglass tube with a height of 2.2 m
and 20 cm in diameter. The tube was filled with prefiltered estuarine water and kept at
a temperature of 24 oC. Plant material touching the walls of the cylinder was
excluded. The settling mode of circular and square sheets of algae was either on the
edge or in a horizontal position. The experiments were repeated with other species of
macroalgae and leaves from rooted vegetation in order to determine the variation in
settling rates. The results are presented in Fig. 8.
M.R. Flindt et al.
Settling rate (m h )
Ulva area (cm )
Figure 8. Settling rates for different sizes and forms of Ulva lactuca. The size is indicated by
the area. Circles and squares indicate circular and square discs of the algae, respectively. No
significant differences between forms or size classes were observed.
It was clear that the settling rates of Ulva were independent of size even over one
order of magnitude. Square pieces of Ulva sp. settled with rates of 17-70 m h-1 and
circular pieces with rates of 17-75 m h-1, thus there seemed to be no significant
differences in the settling rate between the two shapes. Ulva began settling by random
motion and continued to settle in two modes. Settling with the thallus in a horizontal
position was slower (35 m h-1) than settling with the thallus on its edge (60 m h-1).
Settling rate (m h-1)
Plant mass (g wwt)
Figure 9. Settling rates for four macrophyte species with different morphology. Size classes are
indicated by weight (g wwt). There was no significant change in settling rates within the range
of one decade of wet weight for any of the macrophytes. Studies were performed in Venice,
August, 1998.
Plant bound nutrient transport
Our results showed that settling rates of different sized pieces of macroalgae
and a seagrass species was relatively independent of shape or size (Fig. 8, 9). The
comparison of settling rates for different species of macroalgae and rooted plants
showed more variation, although all plants settled relatively fast (Fig. 10).
The red algae, Chondrus sp. and Gracilaria sp. had the highest settling rates (80-85 m
h-1). The red algae Polysiphonia sp. settled much slower (30 m h-1), and the rooted
macrophyte Cymodocea sp. was intermediate. The experiment also showed that there
was no difference in settling rates among Ulva pieces in different conditions (light
and dark thallus); the major difference was in the chaotic settling pattern.
We attempted to apply our experimental findings to observations of current
velocities in Venice Lagoon to get an idea about the erosion and deposition areas
(Bergamasco et al. 2003). The results were that almost all areas in the Lagoon showed
current velocities higher than the measured traction thresholds. Thus, 90% of Venice
Lagoon was subject to macrophyte mobilization.
In some areas this resulted in biomass losses through erosion which were
higher than the plant growth rate. The plants therefore disappeared from these areas.
Settling rate (m h -1)
Figure 10. Average settling rates for all studied plants. The Ulva class is subdivided into light
and dark plants which are settling either with the thallus horizontally (h) or vertically (v)
during the settling.
Despite some variation in the measured settling rates of macrophytes, the
experiments showed that resuspended plant material settle fast. This mechanism is
important in explaining the general picture of the export/import balance at the open
boundary of Venice Lagoon (Fig. 3). It appears that about 90% of the exported plant
biomass did not return with the following flooding tide. The reason is that, that once
in the Adriatic Sea, the plant material within 15-40 minutes will settle to 20 m depth,
long before the tide turns and the strong in-going current eventually could carry the
plant material back into the Lagoon.
M.R. Flindt et al.
It is complicated to evaluate the relative importance of growth, grazing, sporulation
and transport processes for the standing stock of non-attached or loosely attached
macroalgae in a shallow system like the Lagoon of Venice because the dynamics of
the physical, chemical and biological parameters are acting on these processes with
both positive and negative feed-back mechanisms. The plant material caught at the
outer boundary may have different origins which may be difficult to identify. Most
often the caught material is a mixture of plants transported from distant inner areas
and more locally grown. How can the local contribution to the transport loss be
estimated and what is the contribution of this process to the dynamic biomass budget?
Complex and time consuming campaigns are necessary. We tried to make these
estimation based on direct field measurements in Roskilde Fjord and Venice Lagoon.
The growth and grazing rates of Ulva sp. were measured by an enclosure/exclosure
technique described by Geertz-Hansen et al. (1993). Cages were built of 20 cm long
transparent Perspex tubes (Ө = 20 cm) cut longitudinally in halves. The resulting
Perspex arches constituted the roof of the cages and the sides and bottom were
covered with nylon net. The cages were fixed to the sediment surface by drilling
coiled steel pegs into the sediment. In the growth measurements, a mesh size of 0.5
mm was used, preventing grazers from entering the cages. In the grazing
measurement a mesh size of 4 mm was used, allowing the invertebrate grazers to
enter the cage. In each chamber the specific growth rate µ was calculated (eq. 1) from
the increase in surface area of 5 Ulva discs, cut from free-floating thalli with a
sharpened Perspex tube (Ө = 2.5 cm).
µ = (ln (biomasst) – ln (biomasst0)) d-1
where µ is specific growth rate (d-1), biomasst0 and biomasst are the tissue areas at the
start and at the end of the incubation, respectively, and t is the number of days.
The specific grazing rate g (d-1) was calculated as
g = (ln (biomasst,0.5) – ln (biomasst,4.0)) t-1
where biomasst,0.5 and biomasst,4.0 are the final surface areas of Ulva discs in 0.5 mm
and 4.0 mm cages, respectively.
The specific sporulating rate s (d-1) was calculated as
s = (ln (ghost_tissuet) – ln (ghost_tisuet0)) t-1
where ghost_tissuet and ghost_tisuet0 are the surface area of transparent ghost tissue
of Ulva discs area at time t and t0. At the same time plant loss due to transportation
was measured as net catches with the procedure described above.
Plant bound nutrient transport
Figure 11. Growth, grazing and erosion/sedimentation for three transects in Møllekrogen,
southern part of Roskilde Fjord, Denmark, 1996. Biomass losses are negative while gains are
positive. From Salomonsen et al. (2000). The transects East and West represent stations with
an average depth of 1m, while the Deep transect has an average depth of 3m.
The first example (Fig. 11) is from the shallow protected Møllekrogen in the inner
part of Roskilde Fjord, where Ulva lactuca was the dominating macroalgae with a
biomass of 2-35 g dwt m-2. During this study, 10 representative stations were placed
inside an area of 0.42 km2. The stations were placed on three transects, an eastern
and a western with average depths of 1 meter, and a middle deep transect at 3 meters
depth. The tidal amplitude in this area is only 20 cm, resulting in the lowest
hydrodynamic forces among the studied sites (Mondego Estuary, Portugal, Venice
Lagoon, Italy and Roskilde Fjord, Denmark). Measurements of area specific biomass,
growth and loss rates were based on a sampling frequency of twice a week.
It is evident that the changes in rates were more dynamic at the shallow stations
where gain and loss processes constituted a higher fraction of the biomass than at the
deeper stations. This was also reflected by the fact that the deeper area had a higher
biomass than the shallow area.
M.R. Flindt et al.
Although the growth rates were much lower (Salomonsen et al. 1999). In some
periods, the biomass was transported from the shallow area to the deeper areas.
During the study there were no observations of sporulation, while the loss rate by
isopod grazing at the western transect varied between 0.0 and 0.4 d-1 of the biomass.
The maximal loss rate by transportation was about 1.0 d-1, while the biomass import
rate was about 2.0 d-1. However, the transportation loss rate cannot be higher than 1.0
d-1 plus the new biomass produced by growth, altogether about 1.2 d-1, while gains as
import may be several times the existing biomass.
Figure 12. Growth, grazing and erosion/sedimentation for the whole study area in
Møllekrogen, southern part of Roskilde Fjord, Denmark, 1996. Positive and negative bars
indicate growth and loss of biomass, respectively. Rising hatching indicates growth, falling
hatching indicate grazing, and cross-hatching indicate export. The solid line is biomass
abundance. From Salomonsen et al. (1999).
The plant dynamics for the study area is shown in Fig. 12. It is evident that the major
loss process is transportation of plant matter across the open boundary. This process
accounted for 75% of the biomass loss, while grazing only accounted for about 25%,
and sporulation was not observed. The current velocity in this study area was
simulated by a well-calibrated 2D-hydrodynamic model. This model showed that the
current velocity peaked at about 10 cm s-1, which happened between days 15-20,
where the highest plant export was observed (Fig. 12). For the rest of the study
period, the current velocity was below 5 cm s-1.
The second example is from Venice Lagoon, where we measured plant
losses and gains at stations representative for huge subtidal areas and in intertidal
Plant bound nutrient transport
Rate constants (d )
flats. The field techniques and methods applied for measurements of plant loss and
growth rates were the same as described for the previous study. The results from the
summer 1998 are shown in Fig. 13, where the average rates for the whole area were:
Growth 0.075 d-1, sporulation 0.014 d-1, grazing 0.027 d-1, transport 0.064 d-1. It was
verified that transport was the dominating loss process resulting in a slightly negative
biomass balance.
Figure 13. Field measurements of growth, grazing, sporulation and transport of benthic
macrophytes at different stations representing important habitats in Venice Lagoon, August,
1998. The dominating species at the stations is indicated. Positive values indicate gain, and
negative values indicate loss of biomass for the stations.
Looking for patterns in the measured rates at the different locations, it seems
that the grazing rates are independent of whether the locations were subtidal or
intertidal. The grazing rates varied between 0.01 d-1 and 0.04 d-1. The range of
growth rates varied from 0.02 d-1 to 0.16 d-1, with the lowest rates at the high elevated
intertidal flats (Zostera noltii stations). However, these growth rates were not high
compared to maximal rates of about 0.67 d-1 found in Roskilde Fjord (Geertz-Hansen
et al. 1993). Sporulation was only present at the high elevated stations of the intertidal
zone, where it varied from 0.03 d-1 to 0.06 d-1. The sporulation at these stations was
explained by high water temperatures of about 30°C in the stagnant shallow intertidal
pools, which together with a very high and long lasting insolation of about 1800-2000
µmol m-2 s-1, resulted in unfavourable growth conditions for Ulva sp. The balance of
the measured rates was estimated according to the equation:
Balance = growth – (sporulation + grazing + transport)
M.R. Flindt et al.
The negative result was in agreement with the very low biomass of macroalgae at
these stations. The net transport was negative at all stations in this campaign, with the
highest exports in the subtidal areas, where the current velocities were highest and
where the plants were exposed to the current for the longest periods. At the subtidal
Station 10 close to the Lido entrance, where a mixed vegetation was found
(Cymodocea sp. (the dominating species), Chaetomorpha sp., Ulva sp., and Zostera
sp.) the plant population managed to sustain a loss by transport although the loss rate
were 0.13 d-1. The biomass did not decrease significantly during the study.
In May-June 1995 we made a similar study on loss and growth processes at the
station Malamocco (subtidal, mixed vegetation). Here the growth rate was only 0.043
d-1, the grazing rate was 0.01 d-1, the current velocity dependent transport loss ranged
from 0.03 and 0.17 d-1, and no sporulation was observed. Further, we mapped
biomasses of the ephemeral macroalgae and did nutrient mass balances across the
large subtidal flat. The initial biomass of Ulva sp. around the station decreased from
0.82 to 0.10 kg wwt m-2 during the one month study period, while the biomass of
Chaetomorpha sp. decreased from 0.45 to 0.03 kg wwt m-2. Assuming that these
biomasses, growth and loss rates are representative for the complete 4 km2 subtidal
flat, the initial biomass of macroalgae was 600 ton dwt, but decreasing to 61 ton dwt
during one month. By construction of a small dynamic model for the area (Fig. 14),
we reach an understanding of how big a transport loss a community of ephemeral
macroalgae is able to compensate. Furthermore, we have the possibility to study the
importance of the variability of the initial biomass, growth and loss rates.
The simulation shown in Fig 15 A+B was based on the measured initial
biomass and rates from the Malamocco station in 1995. However, it appears that the
simulated biomass did not decrease to the measured final biomass of 61 tons dwt; it
only decreased to 170s ton dwt. Therefore, the loss rates have to be higher in the
model, although the accumulated transported plant mass is 810 and the total loss
about 910 tons dwt. This scenario is very conservative, because the area also receives
a major import from the inner part of the lagoon. The simulation thus verifies that our
measurements of plant bound nutrient transport were realistic. The plant bound
nutrient export made up about 90-95% of the total nutrient export from Venice
Lagoon in this part of the growth season (Flindt et al. 1997a).
We also simulated the influence on plant transport of a macroalgal growth rate which
was increased sufficiently to keep the biomass in steady state (Fig. 15 C+D).
To obtain a steady state for the macroalgae biomass, the growth rate constant has to
be increased to 0.1 d-1, which still is a very realistic rate and far from maximal
growth. The simulation shows that the growth rate and the plant transport rate
alternate. During neap tides the growth rate compensates for the losses during spring
Plant bound nutrient transport
growt_rate_const graz_rate_const
Figure 14. Powersim conceptual diagram of a macroalgae model with three state variables
(boxes): macroalgae biomass accumulated grazing and accumulated transport. Rates (circles):
growth rate, grazing rate and transport rate. The rate constants were the measured rates, while
the transport_rate_constant was made dependent on current speed using the tidal cycle and
introducing neap and spring tide cycles.
The major consequence is that the plant transport increases from 810 ton dwt
to 1857 ton dwt, 2.3 times the measured transport. This shows that high growth rates
support high losses, which is the case in Venice Lagoon. The nutrient loading is very
high in Venice Lagoon, and not limiting the primary producers. Because the average
depth is only 1 meter, light never becomes a limiting factor, apart from when self
shading occur. Deeper channels exist but they only cover a few percentage of the total
lagoon area.
A simulation with a higher growth rate of 0.16 d-1 combined with a grazing rate of
0.03 d-1 is shown in Figure 15 (E+F). This simulation results in a large increase of
macroalgal biomass, increasing from 600 tons dwt to a final amount of about 1800
tons dwt. A tripling within a month, which is not unrealistic. The accumulated
simulated grazing becomes almost twice the initial macroalgae biomass, while the
accumulated plant transport becomes 5 times that measured in 1995.
M.R. Flindt et al.
Plant mass (kg dwt d )
Plant mass (kg dwt)
plant mass (kg dwt d )
Plant mass (kg dwt)
Growth rate
Grazing rate
Sum (grazing)
Sum (transport)
Plant mass (kg dwt d )
plant mass (kg dwt)
Time (days)
Time (days)
Figure 15. The left panels (A+C+E) show the simulated development in the standing biomass
of macroalgae, accumulated plant transport and accumulated grazing. The right panels
(B+D+F) show the realized rates per day for growth, transport and grazing. Scenario (A+B):
Rate constants as measured in the summer 1995: grazing 0.01 d-1, growth 0.043 d-1, and
average plant transport 0.08 d-1. Scenario (C+D): Rate constants as measured during the
summer 1995: Grazing 0.01 d-1 and average plant transport 0.08 d-1, while the growth
constant was increased to 0.1 d-1. Scenario (E+F): Rate constants were measured during the
summer 1998: grazing 0.03 d-1, growth 0.16 d-1, and average plant transport 0.08 d-1.
Plant bound nutrient transport
During several field campaigns in Venice Lagoon (May-June 1995, August 1998,
March 1999, June 1999), Roskilde Fjord (May 1996, August 1996, May-July 1999,
September 1999) and Mondego Estuary (monthly sampling in 2000-2001), growth,
grazing, and transport of macroalgae and rooted macrophytes were measured during
different tidal excursions, while water samples were analysed for dissolved inorganic
nutrients and total suspended solids. Water samples were often taken from platforms
by an ISCO Manning water sampler. Samples were taken with a high frequency
ensuring a fine temporal resolution over the tidal cycles. The samples were filtrated
and analysed for suspended solids, loss on ignition of the particulate organic matter,
total phosphorus and total carbon and nitrogen. The filtrated water was analysed for
dissolved reactive phosphate, ammonia, and nitrite + nitrate, and sometimes also for
dissolved organic nitrogen.
Nets were fixed to wooden poles at the sides and anchored to the
bottom to ensure that plant material did not escape at the bottom. The measured
horizontal transport was normalized to effective exposure of the net, by using the
hydrodynamic simulation to calculate the areal means of the current direction from
the six nearest grid points. This mean value was then used to calculate the effective
width of the net. Other times the current velocity and direction was measured
manually. The net was harvested every 0.5-1 hour, and after sorting, the wwt. of each
species was recorded. Each macroalgal species were analysed for dry/wet weight
ratio, total carbon, total nitrogen and total phosphorus.
The current velocity and direction was either logged or it was calculated by a
local well-calibrated and validated hydrodynamic model (Venice Lagoon: 2D models
and measurements (Flindt et al. (1997)), Umgisser & Zampato (2001) and loggers,
Mondego Estuary: 2D modelling (Duarte et al. (2001)) and measurements, Roskilde
Fjord: 1D model (Flindt & Kamp-Nielsen (1997)), 2D Lagrangian model
(Salomonsen et al. (1999)), 3D finite element model (Umgisser & Flindt (2001)) and
For Venice Lagoon and Roskilde Fjord, the measurements were repeated so
many times that it was possible to calculate yearly mass balances at the outer
boundary. Such mass balances has never before included all nutrient fractions, so here
we are able to verify to what extent plant bound nutrient export is essential at
different tidal regimes, where Venice represents a meso-tidal system and Roskilde
Fjord a micro-tidal system.
Venice Lagoon
Venice Lagoon is 50 km long and the average width is about 10 km covering an area
of 540 km2. It communicates with the Adriatic Sea by three outer boundaries: Lido,
Malamocco and Pellestrina. The width of these boundaries varies from 400 m to 900
m. The average depth is only 1 m, while the maximum tidal excursion is about 1.1 m.
About 20% of the area is inter-tidal flats, which are only periodically submerged,
while the major part of the lagoon is very shallow subtidal areas.
M.R. Flindt et al.
The combination of the shallowness and the relatively large tidal range ensure a short
residence time of 1-10 days depending on the location. Because of the bathymetry
light never becomes growth limiting for benthic plants, which are responsible for 9599% of the primary production. The nutrient loading is about 6000 tons nitrogen and
1000 tons phosphorus, which together with the good light conditions makes the
lagoon one of the most productive aquatic systems (Sfriso, 1992, 1995).
Fig. 16. Venice Lagoon with the active stations during the field campaigns in the summer 1998
and winter 1999.
Nutrient transport was measured at the three outer boundaries (station 10, 60 and 90)
during the field campaigns in summer 1998 and winter 1999. The northern Station 10
covered the drainage from a mixture of Spartina sp. salt marsh areas and intertidal
flats dominated by ephemeral macroalgae and Zostera noltii. The middle and
southern part of the lagoon were covered by measurements at Station 60 and Station
90. These areas were dominated by large subtidal areas where seagrasses and
ephemeral macroalgae dominated. The measured export balances during summer and
winter are presented in Figure 17.
Plant bound nutrient transport
Mass flux (ton P d )
Mass flux (ton N d )
Mass flux (ton P d )
Mass flux (ton N d )
Figure 17. Nitrogen and phosphorus mass balance for Venice Lagoon based on transport
measurements during summer 1998 (A+B) and winter 1999 (C+D). DIN-N is dissolved
inorganic nitrogen, Veg-X is the summed plant bound nutrient transport (nitrogen or
phosphorus), SPM-X is the suspended particulate nutrient, Fyto-X is based on chlorophyll
a.measurements converted by the Redfield-ratio to phytoplankton bound nutrient, and PO4-P is
The mass balances at the three outer boundaries are based on transport measurements
at Stations 10, 60 and 90 (Fig. 16). The results are extrapolated to cover the full
hydraulic cross section area for each of the boundaries. This may introduces an
underestimation of the mass balance because we were only able to carry out the
campaigns at shallow stations. It was impossible to place the nets at the bottom in the
deeper channels where most of the transport occurs.
During the growth season plant bound nutrient transport dominated. About 75% of
the nitrogen export appears plant bound, while about 55% of the phosphorus is
exported through plant movements. Both concentrations of dissolved inorganic
nitrogen and phosphate are low, although higher than during the measurements in
1995. Phytoplankton did not contribute to the mass balances. Its contribution was less
than 0.5% of the mass transport for both phosphorus and nitrogen. Due to the
shallowness and extended boat activities in the area resuspension occurs frequently.
This results in relatively high concentrations of suspended particulate matter, which
for nitrogen accounted for about 20% of the nutrient export during the summer and
M.R. Flindt et al.
about 25% in the winter. The suspended particulate phosphorus transport was about
40% of the mass balance in the summer and 33% in the winter. During the field
campaigns in 1995, 1998 and 1999 the N/P ratio of the suspended particulate matter
was relatively low (1.5-5), indicating that suspended particles act as a phosphate sink
for the freshwater discharge in the winter and spring, but also as a sink for the internal
phosphate loading from the sediment. During the growth season the N/P-ratio of the
transported plants was high (about 15) which may indicate phosphorus limited
During winter export was dominated by dissolved nutrients, where DIN accounted for
about 60%, while dissolved phosphate contributed with 40% of the export. Plant
bound nutrient transport accounted for 23% of the nitrogen export and about 17% of
the phosphorus export.
Nitrogen (ton y )
Phosphorus (ton Y )
Dis. Inorg.
Figure 18. Yearly phosphorus and nitrogen mass balance for Venice Lagoon.
The export measurements were integrated for each season to get the annual mass
balance (Fig. 18). Here it is assumed that the growth season ranges from March until
November (270 days) and the winter covers December-February (90 days). For the
whole year, plant bound nitrogen export accounted for about 63% of the total
calculated export, while plant bound phosphorus export accounted for 42%. The
official nutrient loading is 6000 tons N and 1000 tons P (Flindt et al. 1997b). Our
calculated export was 5100 tons nitrogen and 500 tons phosphorus. This means that
about 50% of the phosphorus is either missing or accumulated inside the lagoon. For
nitrogen, the calculated export accounts for 85% of the loading. Venice Lagoon is net
exporting suspended particulate matter. This is due to erosion of tidal flats. The
reason for this is that the major tributaries are cut off from the lagoon, so they
discharge directly to the Adriatic Sea. The missing sand and silt contribution has
Plant bound nutrient transport
resulted in degradation of salt marsh habitats in extended areas of the northern part of
the lagoon. For this reason we assume that a major part of the missing 50%
phosphorus was transported undetected as bedload transport. A partly reintroduction
of major tributaries will be necessary to keep the inorganic particle budget balanced
in the lagoon; otherwise the salt marshes will degrade. A reintroduction of the major
tributaries will also include iron import that will act as a sink so the internal
phosphate loading from the sediment not becomes realized as new macroalgae
biomass. This will be even more effective if the tributary particle mass contribution
happens in spring where the melting water from the mountains carry huge amounts of
inorganic particles. Here the import of iron and particle mass will be relatively higher,
while the nutrient contribution will be exported before the vegetation starts growing.
Roskilde Fjord
Roskilde Fjord is a shallow estuary connected to Kattegat on the north coast of
Zealand, Denmark (Fig. 19). The length from the inner southern part to the outer
northern part is about 40 km and the maximum width is about 10 km. The surface
area is 122 km2. The average depth is 3 m and the drainage area is 1127 km2. The
tidal excursion is about 0.2 m, so Roskilde Fjord represents a micro-tidal system. The
hydraulics are more dynamic than indicated by the tidal range, because wind forcing
is able to cause changes in water levels of up to 1.6 m above mean sea water level.
Together with the freshwater discharge of about 350 x 106 m3 y-1 from tributaries and
diffuse runoff, these mechanisms create gradients of salinity from about 10 ‰ in the
southern part to about 20 ‰ at the outer boundary. The residence time in the southern
part of Roskilde Fjord is above 1 year and about 2-4 weeks in the northern part close
to the open boundary. This long residence time has huge ecological impacts on light
climate and on the benthic pelagic nutrient coupling. About half of the primary
production is created by benthic primary producers, while phytoplankton is
responsible for the other fraction. The nutrient loading is about 1500 tons nitrogen
and about 60 tons phosphorus (Hedal & Hansen 2001).
For Roskilde Fjord we have grouped the measured integrated nutrient transport across
the outer boundary into three seasons. Most of the nutrient transport took place during
the winter to early spring (Jan-Apr). This period was dominated by high
concentration of nitrate due to freshwater runoff. About 50% of the export was
dissolved inorganic nutrient, while suspended particulate nitrogen contributed with
about 40%. The phytoplankton nitrogen was about 5%, primarily created during the
spring bloom.
The nutrient transport was lowest during the growth season from May to
September. Here the export was dominated by plant material transport that accounted
for about 70% of the nitrogen export. During the autumn and early winter (Oct-Dec),
the suspended matter was dominating the export, but dissolved nutrient and plant
transport accounted for about 25% and 20%, respectively. Dissolved organic nitrogen
was included in these studies, but this fraction accounted at maximum during the
growth season for 0.5% of the nitrogen transport.
M.R. Flindt et al.
Figure 19. Roskilde Fjord. Left: Bathymetric map showing the depth conditions. Right:
Current velocity map presenting the range of current speeds in the fjord during a drop in water
level of 30 cm.
Plant bound nutrient transport
Export (ton N)
Export (ton P)
Figure 20. Nitrogen and phosphorus mass balance for Roskilde Fjord 2000. The upper panel
shows the export of the different N-fractions: NHx-N is ammonia, NOx-N is nitrate + nitrite,
DON-N is dissolved organic nitrogen, SPM-N is suspended particulate nitrogen, Veg-N is
nitrogen bound in vegetation and Fyto-N is phytoplankton nitrogen. The lower panel shows the
export of the phosphorus fractions: POP is phosphate, SPM-P is suspended particulate
phosphorus, Veg-P phosphorus bound in vegetation and Fyto-P is phytoplankton phosphorus.
For the phosphorus transport the tendencies were the same with high dissolved
inorganic phosphate export in January to April, but also with a huge contribution from
suspended particulate phosphorus of about 40%. Phytoplankton transport contributed
about 8% of the phosphorus transport in this period increasing to about 11% in the
growth season. In the growth season, the phosphorus export was dominated by plant
M.R. Flindt et al.
material (55%) and suspended particulate phosphorus. In this period there was a small
net import of phosphate. In October to December, phosphate was regenerated, which
together with an increasing freshwater discharge influence the transport pattern.
Mass transport (%)
Dis. Inorg.
Dis. Org.
Figure 21. Annual contribution of nitrogen and phosphorus fractions to the mass transport in
Roskilde Fjord for 2000. Dis.Inorg is dissolved inorganic nutrients (phosphorus or nitrogen).
Dis. Org. is dissolved organic nutrients. SPM is suspended particulate nutrients. Veg is plant
bound nutrients, and Fyto is phytoplankton nutrients.
An annual nitrogen and phosphorus balance for the different nutrient fractions
contributing to the mass balance at the outer boundary of Roskilde Fjord is shown in
Figure 21. Here it becomes evident that nitrogen and phosphorus mainly are exported
from Roskilde Fjord as dissolved inorganic nutrients and suspended particles (about
70-80% of the total export). But nutrients bound in plant material was also important,
the macrophytes accounted for 24% of the phosphorus export and 18% of the
nitrogen export, while phytoplankton bound transport accounted for about 8% and
3% of the phosphorus and the nitrogen export, respectively.
Mondego Estuary, Portugal.
All study sites were within the South arm of the Mondego Estuary. This location is 10
km long and the average width is about 0.3 km covering an area of 3.4 km2. It
communicates with the North Atlantic Ocean. The average depth is about 2 m, while
the maximal tidal excursion is about 3.3 m. Huge areas are intertidal salt marsh
covered downstream by Spartina maritima and upstream by Scirpus sp. The
Mondego Estuary represents a macro-tidal salt marsh system. The residence time is
about 1-4 days depending on the location. The system has a high turbidity, causing
light to be growth limiting for the submersed and lower intertidal plant communities.
Plant bound nutrient transport
The nutrient loading is about 126 tons nitrogen and about 10 tons phosphorus (Lillebø
et al. 2002).
Figueira da Foz Harbour
North arm
Gala bridge
Murraceira Island
South arm
1 Km
Armazéns channel
Intertidal areas
Pranto river
Figure 22. Map showing the Mondego Estuary with the South arm where the transport
measurements were made.
In Mondego, the nutrient loading and transport is complicated by several processes.
First of all, during rainy periods water drains from agricultural areas in the lower part
of the Mondego Valley through a sluice at the end of the Prante River. These events
are not electronically registered, so the exact discharge of water and nutrient is not
known. But when the freshwater discharge is high, the salinity drops to just above
zero in the inner part of the south arm. This results in collapse of Enteromorpha sp.,
because the macroalgae can not achieve a positive net growth rate at low salinities
(Martins et al. 1999, 2001). If the freshwater discharge is lower, causing salinities
about 5 ‰ in the inner part, huge Enteromorpha blooms are realized. These blooms
happens only in moderately wet years, while very rainy or dry years are free of
blooms, because the losses through sloughing and transportation and/or sporulation is
higher than the net growth rate. During the studied year 2000, there was no spring
bloom of Enteromorpha in the estuary. In Figure 23 representative measurements of
nitrogen and phosphorus transport at the outer boundary are shown for a summer
(Fig.23 A+B) and a winter campaign (Fig.23 C+D). In the summer period dissolved
inorganic nitrogen (DIN) dominated the transport in the system with 60% of the net
export. About 50% of the exported DIN returned during the following import
situation, so about 50% was permanently lost to the coast. With respect to suspended
M.R. Flindt et al.
particulate nitrogen (SPM-N) the net nitrogen export was about 35%, while the
macrophyte bound nitrogen only was about 5% of the total net N-export.
P-transport (kg d )
N-transport (kg d )
P-transport (kg d )
N-transport (kg d )
Figure 23. Nitrogen and phosphorus import, export and mass balance for the Mondego South
arm based on transport measurements during summer 2000 (A+B) and winter 2000 (C+D).
DIN-N is dissolved inorganic nitrogen, Veg-X is the summed plant bound nutrient transport
(nitrogen or phosphorus), SPM-X is the suspended particulate nutrient, Fyto-X is based on
chlorophyll a. measurements converted by the Redfield-ratio to phytoplankton nutrient and
PO4-P is phosphate.
A higher proportion of SPM-N was permanently lost by the transport to the sea due to
a low re-import. The reason for this is that the particles were transported as bedload
so the heavy fraction sedimentated before the water returned to the estuary again. The
phytoplankton bound nutrients only accounted for less than 1% of the mass transport.
The residence time was very short in the estuary, and therefore, if the phytoplankton
is not able to obtain growth rates close to the maximal they are washed out of the
system. The transport pattern for phosphorus during the summer was similar to that
for nitrogen. Here the net phosphorus export was 52%, 40% and 8%, respectively, for
phosphate, SMP-P and macrophyte bound phosphorus. During the early winter DIN
transport dominated the nitrogen export. But the net loss of DIN measured as the
difference between import and export was less, because about 2/3 of the exported
DIN was re-imported. During this winter there was a net import of both nitrogen and
phosphorus in the form of suspended particulate matter. The reason for this net
particle mass import was most probably that some of the huge discharge from the
Plant bound nutrient transport
north arms was forced into the south arm during rising water levels. Proportionally, a
higher amount of the phosphorus transport was particle bound as compared to the
nitrogen transport. Macrophyte and phytoplankton bound nutrient transport was
As earlier stated, unbalanced energy budgets for salt marsh ecosystems led Odum
(1968) to propose the “Outwelling Hypothesis”. This hypothesis has been inspiration
for many investigations and reviews (e.g. Nixon 1980, Dame 1989, 1994, Dame &
Allen 1996) of plant transportation from salt marshes to coastal waters. It seems
evident that North American and European salt marshes respond different as
exporters of the primary production. American salt marshes export a larger fraction of
the production than the European ones. A good explanation was given by Beeftinf
(1977) and Lefeuvre (1994), who stated that the American marshes topographically is
located lower and hereby more exposed to hydrodynamic forcing. Thus, the potential
period for hydrodynamic forced sloughing of plant leafs is much longer. Another
essential difference is that the vegetation coverage is much less in the lower part of
European salt marshes compared with American marshes at the same elevation
(Adam 1990, Packham & Willis 1997), and there is therefore less plant material to
Our study from Mondego is in agreement with other investigations of European salt
marshes (Hemminga et al. 1993, 1996, Lefeuvre et al 1994,), where the general
tendency is that most of the production is kept inside the systems (Bouchard &
Lefeuvre 2000).
Through laboratory and field measurements it has been verified that unattached living
macroalgae start to move as bedload transport at very low current velocities (2-3 cm
s-1), and at about 6-9 cm s-1 most of these are lifted from the bed and come into
suspension where they move with the same speed as the water. Even attached
macroalgae are often transported, and at a current velocity threshold of 12 cm s-1
some ephemeral species start to loose mass due to sloughing. This process increases
with current speed and at about 40 cm s-1, they loose a considerable amount of mass
ranging between 25% and 50% depending on the species. Other macroalgae like
Fucus sp. are relatively insensitive to sloughing, the tissue is simple too strong to be
torn apart. The effect of these low loss thresholds is a potentially high export of
macroalgae from estuaries to coastal waters. Measurements verify that most
submersed plants settle within a range of 35 m h-1 to 85 m h-1. Compared to
phytoplankton, macroalgae settling rates are about 1000-5000 times faster. This
mechanism is important in explaining the general picture of the export/import balance
at the open boundaries between estuaries and coastal waters, where only a small
fraction of the exported plant mass returns with the following flooding tide. The plant
matter simply settles before the tide changes direction. The only complication is the
newly sloughed material from seagrasses, which do not settle due to the
aerenchymatic tissue. Seagrass movements are therefore both affected by current
speed, wind speed and wind direction. Hereby their transport pattern becomes more
complex to predict.
M.R. Flindt et al.
The field measurements presented from Mondego Estuary, Venice Lagoon and
Roskilde Fjord, cover large variations in tidal amplitudes. Mondego Estuary
represents a macro-tidal salt marsh system, Venice Lagoon represents a meso-tidal
estuary, and Roskilde Fjord represents a micro-tidal system.
Table 4.Relative contribution of plant bound nutrient transport to the total mass transport
Mondego Estuary
Venice Lagoon
Roskilde Fjord
Contribution of plant transport to mass balances
Nitrogen (%)
Phosphorus (%)
There are large differences in the relative importance of plant bound nutrient transport
from turbid salt marshes like Mondego (or other European salt marsh systems) and
from estuaries with extensive shallow subtidal areas such as Venice Lagoon or
Roskilde Fjord (table 4). This difference is first of all linked to differences in
hydrodynamic forcing as well as differences in plant society. In European salt
marshes hydrodynamic forcing is so high that turbidity restricts plant growth in the
lower elevated salt marsh areas and in subtidal areas. Logically, a low production will
lead to lower losses. The production is higher at the high elevated marsh areas but the
hydrodynamic forcing is lower and less frequent because the water only reaches the
Spartina beds during the days around spring tide. There is no literature about
sloughing threshold for Spartina, but the threshold is most probably high. Plant
communities in macro-tidal salt marshes and meso- to micro-tidal estuaries differ.
Zostera marina is the dominant subtidal rooted plant in meso- and micro-tidal
estuaries. This species slough by nature because the oldest leaf is rejected when a new
leaf is formed,. This sloughing is non-dependent on current dynamics. Furthermore,
the leaves float in the first period after rejection, so no current energy is needed to lift
it from the bottom before transportation. Free living or attached ephemeral
macroalgae with high transportation potential are also much more abundant in mesoand micro-tidal estuaries, where the residence time is longer and the hydrodynamic
forcing is less. These mechanisms explain why estuaries like Venice Lagoon and
Roskilde Fjord export large quantities of plant matter, and thus also a large nutrient
mass associated with this export.
Monitoring programs around the world are set up to understand and evaluate the
effects of agricultural dominated nutrient runoff. Managers and scientists working
with applied problems like the reasons for oxygen depletion in coastal waters, need to
have reliable nutrient mass fluxes at the outer boundary of estuaries. Today, plant
bound nutrient transport is not included in these mass fluxes, which leads to nutrient
balances for the estuaries where the estimated nutrient retention is too high. The
consequence is underestimation of the nutrient loading to the open coastal areas. For
Venice Lagoon and Roskilde Fjord the yearly underestimation is 60% and 18% on the
nitrogen budget and 44% and 23% on the phosphorus budget, respectively.
Plant bound nutrient transport
The studies were financed by the European MAST III Program through activities
linked to the F-ECTS project (MAS3-CT97-CT97-0145), the European TMR
Program through activities linked to the WET project (FMRX-CT96-0051), and by
the Portuguese Praxis program through a PhD grant.
Adam, P. (1990). Saltmarsh ecology. Cambridge University Press, Cambridge. 461pp.
Amos, C.L., Cloutier, D., Cristante, S. & Cappucci, S. (2000). The Venice Lagoon study (F-ECTS) –
February, 1999. Geological Survey of Canada Open File Report 3904: 47p.
Beeftink, W.K. (1977). The coastal salt marshes of Western and Northern Europe: an ecological and
phytosocological approach. In: (Chapman, V.J. (Ed.), Wet Coastal Ecosystems (pp. 109-155).
Amsterdam: Elesvier.
Borum, J. (1996). Shallow waters and land/sea boundaries. In B.B. Jørgensen & K. Richardson (Eds.)
Eutrophication in Coastal Marine Ecosystems. (pp. 179-203) Washington DC, American Geophysical
Borum, J. (1997). Ecology of coastal waters and their response to eutrophication. In K. Sand-Jensen & O.
Pedersen (Eds.), Freshwater Biology (pp. 102-115). Copenhagen: Gad.
Bouchard, V. & Lefeuvre, J.-C. (2000). Primary production and macro-detritus dynamics in a European
salt marsh: carbon and nitrogen budgets. Aquatic Botany 67, 23-42.
Dame, R.F. & Stilwell, D. (1984). Environmental Factors Influencing Macrodetritus Flux in North Inlet
Estuary. Estuarine, Coastal and Shelf Science 18, 721-726
Dame, R.F. (1989). The importance of Spartina alterniflora to Atlantic coast estuaries. Criical. Review of
Aquatic Sciences, Vol. 1 (pp. 639-660).
Dame, R.F. (1994). The net flux of materials between marsh-estuarine systems and the sea: the Atlantic
coast of the United States. In W.J. Mitsch (Ed.), Global wetlands old world and new (pp. 295-302).
New York: Elsevier.
Dame, R.F., & Allen, D.M. (1996). Between estuaries and the sea. Journal of Experimental Marine
Biology 200, 169-185.
Duarte A.S., Pinho J.L., Pardal M.A., Neto J.M., Vieira J.P., Santos F.S. (2001). Effect of residence times
on River Mondego estuary eutrophication vulnerability. Water Science and Technology, 44 (2-3), 329336
Enriquez, S., Duarte, C. & Sand-Jensen, K. (1994). Patterns in decomposition rates among photosynthetic
organisms: The importance of detritus C:N:P. Oecologia 94, 457-471.
Flindt, M., Salomonsen, J., Carrer, M., Bocci, M., & Kamp-Nielsen, L., (1997a). Loss, growth and
transport dynamics of Chaetomorpha area and Ulva rigida in the Lagoon of Venice during an early
summer field campaign. Ecological modelling, 102, 133-141.
M.R. Flindt et al.
Flindt, M.R., Marques, J.C., Partial, M.R., Bocci, M., Bendoriccho, G., Kamp-Nielsen, L., Nielsen, S.N., &
Jørgensen, S.E. (1997b). Description of the three shallow estuaries: Mondego River (Portugal),
Roskilde Fjord (Denmark) and the Lagoon of Venice (Italy). Ecological Modelling, 102, 17-31.
Flindt, M.R. & Kamp-Nielsen, L. (1997). Modelling an estuarine eutrophication gradient. Ecological
Modelling. 102: 143-154
Flindt, M.R., & Kamp-Nielsen, L. (1998). The influence of sediment resuspension on nutrient metabolism
in the eutrophic Roskilde Fjord, Denmark. Verhandlungen Internationale Vereinigung Limnologie, 26,
Flindt, M.R., Pardal, M.A., Lillebø, A.I., Martins, I., & Marques, J.C. (1999). Nutrient cycling and plant
dynamics in estuaries: A brief review. Acta Oecologica, 20, 237-248.
Geertz-Hansen, O. Sand-Jensen, K., Hansen, D.F., & Christiansen, A. (1993). Growth and grazing control
of abundance of the marine macroalga, Ulva lactuca L. in a eutrophic Danish estuary. Aquatic Botany
46, 101-109.
Greenway, M. (1976). The grazing of Thalassia testudinum in Kingston Harbour, Jamaica. Aquatic Botany
2, 117-126.
Hedal, S. & Hansen, L.R. (2001). Roskilde Fjord 2000. Frederiksborg Amt.
Hemminga, M.A., Cattrijsse, A. & Wielemaker, A. (1996). Bedload and Nearbed Detritus Transport in a
Tidal Saltmarsh Creek. Estuarine, Coastal and Shelf Science 42, 55-62.
Hernández, I., Peralta, G., Pérez-Llorens, J. L., Vergara, J., & Niell, F. X. (1997). Biomass and dynamics
of growth of Ulva species in Palmones river estuary. Journal of Phycology 33, 764-772.
Kamp-Nielsen, L. (1992). Benthic pelagic coupling of nutrient metabolism along an estuarine
eutrophication gradient. Hydrobiologia, 235/236, 457-470.
Lefeuvre, J.C. (1994). Comparative studies on salt marsh processes in the baie du Mont Saint-Michel: a
multi-disciplinary study. In: Mitsch, W.J. (Ed.). Global Wetlands: Old World and New. (pp. 215-234).
Amsterdam. Elsevier.
Lillebø, A.I, Pardal, M.A., Flindt, M.R, Neto, J.M., Macedo, F., Martins, I. &
Marques, J.C. (2002). Nutrient dynamics in the intertidal pools of the Mondego Estuary. IV. Possible
contribution to dissolved inorganic phosphorus loading. In Aquatic Ecology of the Mondego River
Basin. Global importance of local experience. Ed. Pardal, Marques & Graca. (pp. 287-300). Coimbra
University Press.
Martins, I., Oliveira, J.M., Flindt, M.R. & Marques, J.C. (1999). The effect of salinity on the growth of
Enteromorpha intestinalis (Chlorophyta) in the Mondego estuary (west Portugal). Acta Oecologica. 20
(4) 259-265.
Martins, I, Pardal, M.A., Lillebø, A.I., Flindt, M.R. and Marques, J.C. (2001). Hydrodynamics as a major
factor controlling the occurrence of green macroalgal blooms on the influence of precipitation and
river management. Estuarine, Coastal and Shelf Science. 52: 165-177.
Menzies, R.J., Zaneveld, J.S., & Pratt, R.M. (1967). Transported turtle grass as a source of organic
enrichment of abyssal sediments off North Carolina. Deep-Sea Res. 14, 111-112.
Menzies, R.J., & Rowe, G.T. (1969). The distribution and significance of detrital turtle grass Thalassia
testudinum on the deep sea floor off North Carolina. Internationale Revue der Gesamten
Hydrobiologie 54, 217-222.
Nixon, S.W. (1980). Between coastal marshes and coastal waters – A review of twenty years of speculation
and research on the role of salt marshes in estuarine productivity. In P. Hamilton & K.B. MacDonald
(Eds.), Estuarine and wetlands processes (pp. 437-525). New York: Plenum Press.
Plant bound nutrient transport
Odum, E.P. (1968). A research challenge: evaluating the productivity of coastal and estuarine water,
Proceedings of the 2nd Sea Grant conference (pp. 63-64). Kingston: University of Rhode Island.
Odum, W.E., Fischer, J.S., & Pickral, J.C. (1979). Factors controlling the flux of particulate organic carbon
from estuarine wetlands. In R.J. Livingston (Ed.), Ecological processes in coastal and marine systems
(pp. 69-80). New York: Plenum Press.
Packham, J.R., & Willis, A.J. (1997). Ecology of Dunes, Salt marsh and Shingle. Chapman & Hall,
London. 335 pp.
Pedersen, M. F. (1995). Nitrogen limitation of photosynthesis and growth: comparison across aquatic plant
communities a Danish estuary (Roskilde Fjord). Ophelia, 41, 261-272.
Salomonsen, J., Flindt, M.R., & Geertz-Hansen, O. (1997). Significance of advective transport of Ulva
lactuca for a biomass budget on a shallow water location. Ecological Modelling, 102, 129-132.
Salomonsen, J., Flindt, M.R., & Geertz-Hansen, O. (2000). Modelling advective transport of Ulva lactuca
(L) in the sheltered bay, Møllekrogen, Roskilde Fjord, Denmark. Hydrobiologia, 397, 241-252.
Sfriso, A., Pavoni, B., Marcomini, A., & Orio, A. A. (1992). Macroalgae, nutrient cycles, and pollutants in
the Lagoon of Venice. Estuaries,15, 517-528.
Sfriso, A., (1995). Temporal and spatial responses of growth of Ulva rigida C. Ag. To environmental and
tissue concentrations of nutrients in the Lagoon of Venice. Botanica Marina, 38, 557-573.
Sfriso, A. & Marcomini, A. (1997). Macrophyte Production in a Shallow Coastal Lagoon. Part 1: Coupling
with Chemical-Physical Parameters and Nutrient Cencentration in Waters. Marine Environment
Research 44(4), 351-375.
Thomson, C.E.L. & Amos, C.L. (2002). The impact of mobile Disarticulated Shells of Cerastoderma edulis
on the Abrasion of a Cohesive Substrate. Estuaries 25, 204-214.
Umgiesser, G. & Zampato, L. (1997). Hydrodynamic and salinity modelling of the Venice channel network
with coupled 1-D-2-D mathematical models. Ecological modelling 138, 75-85.
Umgiesser, G. & Flindt, M.R., Amos, C.L. & Bergamasco, A. (2001). Exporting the phytobenthos-reaction
modeling approach: Roskilde Fjord case study. Proceedings from the ELOISE-congress, Bavena, Italy.
Vermat, J.E. & Sand-Jensen, K. (1987). Survival, Metabolism and growth of Ulva lactuca under winter
conditions: a laboratory study of bottlenecks in the life cycle. Marine Biology 95, 55-61.
Viaroli, P., Pugnetti, A., & Ferrari, I. (1991). Ulva rigida growth and decomposition processes and related
effects on nitrogen and phosphorus cycles in a coastal lagoon (Sacca di Goro, Po River Delta). 25th
Proc European Marine Biology, 77-83.
Viaroli, P., Bartoli, M., Bondavalli, C., Christian, R.R., Giordani,G., & Naldi, M. (1996). Macrophyte
communities and their impact on benthic fluxes of oxygen, sulphide and nutrients in shallow eutrophic
environments. Hydrobiologia, 329, 105-119.
Viaroli, P. Bartoli, M., Fumagalli. I. & Giordani G (1997). Relationship between benthic fluxes and
macrophyte cover in a shallow brackish lagoon. Water Air and Soil Pollution, 99 (1-4), 533-540.
Zieman, J.C., Thayer, G.W., Robblee, M.B., & Zieman, R.T. ( 1979). Production and export of sea grasses
from a tropical bay. In R.J. Livingston (Ed.), Ecological processes in coastal and marine systems (pp.
21-33). New York: Plenum Press.
M.R. Flindt et al.
M.R. Flindt & F.Ø. Andersen: Institute of Biology, University of Southern Denmark,
Campusvej 55, DK-5230 Odense M, Denmark.
J. Neto & M.A. Pardal: Institute of Marine Research, Department of Zoology,
University of Coimbra, Portugal.
C.L. Amos: Southampton Oceanographic centre, University of Southampton,
A. Bergamasco: CNR, Messina, Sicily, Italy.
C.B. Pedersen: Freshwater Biological Laboratory, University of Copenhagen,
Helsingørsgade 51, DK-3400 Hillerød, Denmark.
For more than two decades it has been recognized that benthic suspension feeders, by
filtering particles out of the water, may have a major impact on the overlying water
column. Benthic suspension feeders comprise a large group of both passive and active
filter feeders, but only the impact of actively pumping macro-zoobenthic organisms
will be dealt with. Of these are bivalves like mussels, oysters and clams the most
investigated benthic suspension feeders. Other important groups are ascidians,
sponges, polychaetes and bryozoans. By actively pumping water past a filtering
apparatus, benthic suspension feeders are able to capture particles from the size of big
bacteria of 1-2 µ (sponges and ascidians) or small phytoplankters or other particles of
about 4 µ (bivalves) 100% efficiently. The activity of benthic suspension feeders is
influenced by a number of environmental characters like temperature and particle
concentration. Within e.g. the temperature range of a given suspension feeder,
increase in temperature will result in increased clearance rate (e.g. Petersen &
Riisgård 1992). With increased particle load, clearance will on the other hand be
reduced above a certain threshold level. The different species of benthic suspension
feeders differ to some extent with regard to how the respond in relation to
environmental parameters. Some species, like most bivalves, have the ability to sort
particles and are thus well adapted to the turbid environment of many estuaries. These
eco-physiological aspects of benthic suspension feeding are, however, not the topic of
my presentation. It is rather the ecological impact of the grazing by benthic
suspension feeders on the biological structure of the water column.
Two different but related approaches have been used to document the importance of
benthic suspension feeders in shallow coastal waters: “Ask the animals” is an ecophysiological view of the potential impact of suspension feeding populations of
macro zoobenthos, whereas “ask the water” is an ecological approach focused on
studying the water column.
Ask the animals
By “asking the animals”, physiological parameters like filtration rate, growth or
energy budgets of individual benthic suspension feeders are extrapolated to
population level. By determining filtration rate of a dominant benthic suspension
feeder as a function of size, temperature and concentration of particulate organic
S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling:
The IInfluence of Primary Producers, 129-152.
© 2004 Kluwer Academic Publishers. Printed in the Netherlands.
J. K. Petersen
material (POM), a model of filtration in relation to environmental conditions can be
established. Combined with estimates of population size, geographical extension and
size structure of the population in consideration, an estimate of benthic grazing
potential in a given area can be calculated. Reported grazing potentials may be high,
often more than one time the volume of the investigated ecosystem per day (Jordan &
Valiela 1982, Loo & Rosenberg 1989, Petersen & Riisgård 1992, Petersen et al.
An example of this kind of approach is the study (Petersen & Riisgård 1992) of the
ascidian Ciona intestinalis in Kertinge Nor, a shallow Danish cove. The area of
extension and the density of the ascidian population were determined several times by
scuba diving and photogrammetrical analysis of stereophotographs. Clearance rates
of the ascidians were determined in the laboratory as a function of size and
temperature. From population density, size structure and water temperature estimates
of grazing capacity could be estimated. As seen in table 1, the potential grazing
impact of the ascidians in the cove was always high but variable due to both
individual capacities of the ascidians as well as their population density. In other
systems dominated by organisms with a more stable population structure, like
perennial mussel, the variation in benthic grazing potential is likely less than in this
highly dynamic annual species.
A more integrative approach is to extend the population description to include
secondary production and compare this with either grazing potential or primary
production. It has thus been shown that secondary production by infaunal suspension
feeding bivalves could be of the same magnitude as the primary production in a
shallow semi-exposed habitat (Möller et al. 1985, Loo & Rosenberg 1996). However,
in the more exposed Laholm Bay with deeper water, the same infaunal species did not
fully exploit their grazing potential and secondary production was about 50% of
primary production (Loo & Rosenberg 1989).
Using estimates of secondary production, as an indication of benthic control is in fact
one step beyond the classical “ask the animals” approach. The primary question for
benthic ecologists has often been to determine limiting factors for zoobenthos
populations (see e.g. Fréchette & Lefaivre 1990, Fréchette et al. 1992, Wildish &
Kristmanson 1997). From this point of view, food can be a limiting factor whether
seen on the small scale of a mussel bed (Frechette & Bourget 1985a, Fréchette &
Bourget 1985b) or on a larger ecosystem level (Beukema & Cadée 1986, Smaal &
van Stralen 1990, Smaal et al. 2001). For both levels the concept is to understand how
food supply has the potential to structure populations. The next natural step is to
apply the concept of food limitation to a manageable tool and estimate or model
carrying capacity of those coastal areas used for cultivation or fishery of edible
benthic suspension feeders like mussels, clams and oysters. It is, however, important
that the carrying capacity characteristic of a system involves both physical (mixing)
and biological (primary production) properties.
Ask the water
By asking the water main focus is put on water column parameters and it is the study
of these that lead to the conclusions regarding benthic control.
Grazing on pelagic primary producers
Table 1. Population clearance of the ascidian Ciona intestinalis in Kertinge Nor based on
individual clearance rate and population abundance. T1/2 is the mean residence time for an
algal cell over the ascidian population and Q is the quotient of the ascidian filtration capacity
to the water volume of the cove.
ml min-1
ind m-2
l h-1 m-2
J. K. Petersen
In this approach, phytoplankton biomass and productivity are thus compared with
available nutrients and various loss factors like zooplankton grazing or advection and
other limiting factors like light extinction. The result is then that the loss and limiting
factors cannot account for the difference between potential phytoplankton
biomass/production and actually experienced levels of e.g. phytoplankton biomass in
a given estuary. The only remaining explanation is that large populations of benthic
suspension feeders, most often bivalves, act as a phytoplankton sink. A variation over
the same theme is changes in time in phytoplankton concentration coincident with
invasion of a suspension feeder. A classic example is the studies of San Francisco
Bay (Cloern 1982, Nichols 1985, Alpine & Cloern 1992). Using a simple model
describing change of phytoplankton concentration where dispersive transport and
zooplankton grazing balance growth rate, calculated concentrations of phytoplankton
were much higher than actually observed concentrations in South San Francisco Bay
(Cloern 1982). This led to the conclusion that some other factor had to control
phytoplankton biomass. By analysing benthic fauna composition and estimating the
filtration capacity of the benthic suspension-feeders, primarily clams, it was shown
that these had the capacity to clear the water column more than once per day and
concluded that benthic infauna thus control phytoplankton in South San Francisco
Bay. It could further be shown that invasion of a non-indigeneous clam in the
northern part of the bay, resulted in persistently low levels of chlorophyll a (fig. 1,
(fig. 28b) and a 5-fold reduction of primary production (Alpine & Cloern 1992).
Similarly Hily (1991) stated that in the eutrophic Bay of Brest, phytoplankton
biomass remained low in summer and fall most likely due to huge populations of
benthic suspension feeders that have the potential of filtering 30% of the bay volume
per day. In a similar approach but covering a longer time period, Chauvaud et al
supported the view that benthic suspension feeders, and in particular the gastropod
Crepidula fornicata, mitigate the effects of eutrophication by keeping phytoplankton
biomass low despite increased nutrient input to the Bay of Brest. Møhlenberg (1995)
demonstrated in a shallow nutrient rich Danish cove how stratification stimulated
phytoplankton growth, whereas vertical mixing resulted in loss of phytoplankton
biomass. The decreases in phytoplankton biomass were unaccounted for by changes
in net community growth rate of the phytoplankters and were explained by grazing by
a large population of mussels.
Enclosure experiments
Another approach to the determination of potential benthic control of phytoplankton
has been the use of enclosures or mesocosms. The assumption is that in situ studies
are difficult to reduce to a series of conclusive measurements and that laboratory
experiments cannot necessarily be translated into natural ecological events. Hence by
enclosing natural estuarine water and manipulating with expected controlling factors
like nutrient supply and/or benthic grazing more specific but ecological relevant
answers could be obtained. A number of different mesocosm experiments have been
carried out with the aim studying the effect of benthic grazing on the phytoplankton
community. The systems encompass a huge range in enclosed volume, duration,
mixing and how suspension feeding is induced (Riemann et al. 1988, Sullivan et al.
1991, Olsson et al. 1992, Graneli et al. 1993, Prins et al. 1995). The experiments tend
Grazing on pelagic primary producers
to give variable results due to unforeseen incidences in individual enclosures, but the
general pattern is that benthic suspension feeding reduces overall levels of
phytoplankton biomass. Through excretion of waste products, the suspension feeders
can further enhance phytoplankton growth rates (Dame 1996) and lead to a
succession in the planktonic community towards small opportunistic species
(Riemann et al. 1988). Thus, suspension feeders, supporting the notion that they play
a key role in shallow estuarine habitats, can mediate both top-down and bottom-up
control of phytoplankton.
Figure 1. Time series of chlorophyll a concentration and mean abundance of the nonindigenous clam Pomatocorbula amurensis before and after its introduction to northern San
Francisco Bay (from Cloern 2001).
Freshwater experiences
In freshwater research, benthic grazing on phytoplankton had in general received less
attention until the invasion of zebra mussels (Dreissena polymorpha) in North
American freshwater lakes, rivers and streams. Since their first appearance in the mid
1980’s several studies have been carried out demonstrating a number of effects
(Rutherford et al. 1999): a) reduced levels of phytoplankton or chlorophyll a
concentration irrespective of eutrophication status (e.g. Fahnenstiel et al. 1995a); b)
increased water clarity stimulating growth of aquatic macrophytes (Fahnenstiel et al.
J. K. Petersen
1995b) and changing stratification patterns (Yu & Culver 2000); c) increase in the
frequency and duration of blue-green algal blooms (Fahnenstiel et al. 1995a, Nalepa
& Fahnenstiel 1995); d) a general shift in energetic pathways from the pelagic to the
benthic organisms (Rutherford et al. 1999) thereby e) affecting zooplankton biomass
and composition (Lavrentyev et al. 1995, Horgan & Mills 1999, Idrisi et al. 2001);
and f) altering the benthic communities (Lozano et al. 2001).
The studies of the impact of zebra mussels on ecosystems and accompanying studies
on other freshwater benthic suspension feeders have followed more or less the same
pattern as marine investigations with focus either on the bivalves and their grazing
capacity or on the water column parameters. Due partly to the enforcement of the
study objective to the scientific community by the invasion of the mussels and partly
to the blessings of access to pre- and post-invasion data, there has been a tendency
towards more integrated studies, that clearly demonstrates the ecosystem effects of
benthic grazing.
Figure 2. The concentration boundary layer which develops over a blue mussel reef in a
unidirectional flow, Z, water depth; X, mussel bed length; C0, initial seston concentration,
from Wildish & Kristmanson (1997).
Despite the differences in scientific approach it is evident that benthic suspension
feeders play a dynamic role in the structure and function of coastal ecosystems. In
addition, some species are of economic interest and the harvest or cultivation of these
species requires management. Structurally, benthic suspension feeders can influence
the bathymetry of creeks, provide new habitats in the form of dense epibenthic beds,
increase bottom roughness and foul all hard substrata including man made
installations. Functionally, they have the potential of mitigating effects of
eutrophication by limiting phytoplankton and seston concentrations in the water
column and thereby alter trophic structures. But some of these roles are speculative
due to the shortcomings of the scientific approaches. When asking the animals, it is
thus obvious that laboratory derived filtration rates are not necessarily applicable to
natural conditions. Hence, even in shallow systems the grazing potential is not
realized as evidenced by the presence of chlorophyll a in areas with a benthic grazing
Grazing on pelagic primary producers
potential exceeding the water volume of the system up to several times. The attempts
to duplicate the complexity of the estuarine ecosystem in mesocosms or enclosures
also suffer from weaknesses, notably the lack of sufficient replication, their
vulnerability to unforeseen and chaotic incidences in one treatment/enclosure and the
problems with physical mixing and edge effects due to the walls. And asking the
water has so far been done more rarely, giving partly correlative answers or been
local rather than global in significance. These shortcomings are very much related to
the physical dimension of the habitats of marine benthic suspension feeders, i.e. water
movement and stratification. For that reason, freshwater experiences cannot be
directly applied to coastal areas due to the absence of tidal currents and weaker and
less persistent (only thermal) stratification patterns in freshwater systems. If the
analogy of benthic suspension feeders in a eutrophic estuary is a vacuum-cleaner and
that these unlike a real vacuum-cleaner are sessile, it is not trivial how, how fast and
how much of the carpet is swept over the vacuum-cleaner if one wants a clean floor.
Benthic suspension feeding in tidally mixed areas
From the management and research of exploited bivalve stocks it has become evident,
that these animals are dependent on flow of water to supply them with food, i.e.
hydrodynamically limited (Wildish & Kristmanson 1997). It is intuitively obvious
that in still water, benthic suspension feeders will rapidly deplete the part of the water
column they can access and thus rapidly become food limited. A sustainable size of a
population of benthic suspension feeders is thus dependent on the rate of food
supplied by advection, besides from what can be obtained by bio-mixing resulting
from exhalent currents (Larsen & Riisgård 1997).
Assuming a balance between turbulent mass transfer of seston and consumption at the
boundary in a well developed boundary layer, Wildish & Kristmanson (1979)
proposed that bed consumption rate is dependent on benthic grazing potential:
where NA is the food flux in g m-2 h-1, B is the biomass of suspension feeders in g m, R is clearance capacity in m3 g-1 h-1 and α is a dimensionless arbitrary expression of
filtration efficiency, indicating the degree of re-filtration and/or irregular filtration
patterns. Flux of organic material to the bottom can be defined by free stream seston
concentration and current velocity:
NA = γufree (C0 – C’)
where ufree is free stream current velocity in m h-1, C0 and C’ respectively upstream
and effective food concentration in g m-3 and γ is a dimensionless hydrodynamic
parameter related to bottom roughness. Equations 1 and 2 can be combined:
(C0 – C’)/C’ = BRα/γufree
J. K. Petersen
The right-hand side of the equation is the ratio of filtration capacity to the turbulent
mixing intensity, or the seston depletion index (Wildish & Kristmanson 1997). When
it exceeds a certain threshold value, depletion effects can be assumed to occur
(Herman et al. 1999). Upstream or at the edges in bidirectional tidal flows, bed
consumption rate and flux of organic matter to the bottom may not be in balance and
the seston concentrations near the bottom will vary with depth and distance from the
leading edge. But a depletion effect will gradually build up resulting in the formation
of an increasing concentration boundary layer (fig. 3) that can be adequately
described by equation 3. Further model studies (Wildish & Kristmanson 1984,
Muschenheim 1987, Frechette et al. 1989) have shown that the concentration
boundary layer will depend on current velocity, mussel bed path length and water
depth. In general, the models predict that decreasing current speed and increasing
mussel bed length will lead to an increased concentration boundary layer, finally
reaching the surface, and to larger vertical gradients in seston concentration. Both
vertical and horizontal seston profiles over experimental mussel, found in flume
studies (Wildish & Kristmanson 1984, Butman et al. 1994), have more or less
accurately confirmed the model predictions.
Some evidence in favour of this general model has also been presented. A number of
studies have thus shown that mussels near the edges of mussel beds have higher
growth rate and higher meat contents than mussels in the centre of patches (Newell
1990). Similarly, Fréchette & Bourget (1985b) demonstrated that mussels raised 1 m
above the bottom grew significantly better than mussels on the bottom and explained
it by differences in food concentration (Frechette & Bourget 1985a). Few others have
been able to demonstrate vertical seston depletion (fig. 2) above dense beds of
benthic suspensions-feeders (Frechette et al. 1989, Muschenheim & Newell 1992),
but the results were ambiguous due to e.g. increased vertical mixing by wind
(Frechette & Bourget 1985) or varying current speeds (Frechette et al. 1989).
Measurements of near-bed vertical depletion of chlorophyll a rarely give uniform
results due to resuspension and biodeposition. Demonstration of vertical profiles is,
however, not limited by practical problems only. The model assumptions of e.g.
homogeneous roughness will not hold in natural mussel beds, not even in infaunal
suspensions-feeder beds (O'Riordan et al. 1995). Inhomogeneous roughness can thus
emerge from differences in 3D-structure or gaps in mussel beds (Frechette et al.
1989), but is also a result of exhalent jet currents that act as solid objects protruding
from the bed and resulting in an increase in bed roughness of up to 2 orders of
magnitude compared to a flat bed (Duren et al. 2002). Hence, these animals can
significantly affect their food supply by altering their physical environment (Herman
et al. 1999).
Horizontal depletion of phytoplankton or seston has been demonstrated in tunnels and
in situ flumes (e.g. Asmus et al. 1992). However, enclosing a body of water in a
tunnel or a flume may produce artificial and unrealistic hydrodynamic effects, unless
care is taken in the design of the flume or tunnel. Early field observations have also
demonstrated horizontal depletion over populations of bryozoans and sponges (Buss
& Jackson 1981) or in shallow coves (Carlson et al. 1984), but in situ studies are few
and rarely demonstrate effects on an ecosystem level. Noren et al. (1999) and Haamer
& Rodhe (2000) showed that water passing a mussel bed in the Sound between
Sweden and Denmark was depleted of chlorophyll a and phytoplankton cells in the
Grazing on pelagic primary producers
entire height of the fully mixed water column over a several km long more or less
continuous mussel bed, but experimental studies on a larger scale are in general rare.
Chl. a, µg/l
Figure 3. Depletion of chlorophyll a as a function of height (cm) over a mussel bed in macrotidal Oosterschelde. Relation between logarithm of depth and concentration of chlorophyll a is
indicated by a line. Redrawn from Møhlenberg & Petersen (1998).
Suspension feeding in micro-tidal areas
In micro-tidal areas with low current velocities of 5-10 cm s-1 or less many of the
same characteristic features, like concentration boundary layers or horizontal profiles,
of coastal areas with higher current velocities are seen. Vertical profiles of
chlorophyll a above benthic suspension feeder beds have thus been observed in a
number of studies (Riisgård et al. 1996a, Riisgård et al. 1996b, Vedel 1998, Dolmer
2000b, Dolmer 2000a, Petersen et al. 2000). In Kertinge Nor, Riisgård et al. (1998)
demonstrated the occurrence of density driven currents in periods with large
differences in density at the borders of the cove. In the cove proper this resulted in
stratification with relative well defined, but opposing, currents above and below the
pycnocline with current velocities around 1 cm s-1. Below the pycnocline, but just
above the population of Ciona intestinalis, depletion of chlorophyll a in the passing
water mass could be observed.
Advective transport generated by tide or density driven currents is, however, not the
only structuring factor in micro-tidal shallow areas. Wind generated mixing is thus an
important means of providing food to dense populations of benthic suspension
feeders. In a shallow semi-enclosed basin over mussel beds it could thus be shown
J. K. Petersen
that stratification lead to vertical gradients in chlorophyll a and build up of
phytoplankton biomass, whereas destratification leads to uniform chlorophyll a
concentrations in the entire water column and decreasing biomass (Møhlenberg
1995). It was further shown that vertical mixing was primarily caused by larger wind
events where mean wind of 4-6 m s-1 through 12 hours was sufficient to break down
stratification. In Kertinge Nor it could be demonstrated that filtration by the dense
population of ascidians resulted in vertical profiles of chlorophyll a in the water
column even during fully mixed conditions. It could further be demonstrated that the
steepness of the seston depletion was dependent on the degree of mixing and duration
of respectively stratified and mixed conditions (fig. 4). Stratification was found to be
dependent primarily on wind conditions and to a lesser extent on advective currents.
Further, the height of the concentration boundary layer was dependent on whether the
ascidians were attached to eelgrass leaves, and thus lifted into the water column, or
were lying on the bottom in clumps. Other studies have demonstrated the importance
of vertical mixing to phytoplankton blooms in shallow estuaries with huge biomass of
benthic suspension feeders (Cloern 1991, Koseff et al. 1993).
An important implication of the occurrence of dense populations of benthic
suspension feeders in micro-tidal and/or unexposed areas is that flow does not
necessarily restrict these populations in the classical sense. If sufficient food is
produced due to eutrophication and that the food can be made accessible to the
benthic community by mixing due to wind/waves, dense populations can be
maintained as seen in e.g. Danish estuaries (Conley et al. 2000).
Plant canopies
Plant canopies like seagrass meadows or kelp forests are generally believed to alter
the current pattern by decreasing current velocity within the canopy but increasing it
just above and on the sides of the canopy. Concurrent with the decrease in current
speed within the canopy or seagrass bed, turbulence is increased especially at the top
of the leaf canopy. The reduced flow inside the canopy may explain enhanced
sedimentation rates (Gacia et al. 2002) and reduced resuspension (Terrados & Duarte
2000). The microcosm of particles within a seagrass canopy has been shown to have
an increased biovolume of particles compared to the above canopy concentrations,
but with a larger fraction of detrital material of both angiosperm and planktonic origin
(Duarte et al. 1999). Plant canopies may thus act as both sources and sinks of
The effects of plant canopies on benthic suspension feeders are debated and previous
results are not unequivocal (Wildish & Kristmanson 1997). Some results show
increased growth, others decreased growth compared to outside canopies. The
contradicting results may have several explanations. Differences in growth response
can be a result of physiological differences in the ability of different suspension
feeders to digest and utilize seagrass or macro-algae material or differences between
plant species in nutritious value. Alternatively, attachment of the suspension feeders
to the leaves of the plants could have lifted the benthic suspension feeders differently
up from the bottom and thus differently away from the food-limited concentration
boundary layer in the different investigations.
Grazing on pelagic primary producers
Eutrophication may interact with benthic suspension feeders at several levels: 1)
increased nutrient input may lead to increased primary production and standing stock
of phytoplankton biomass, thereby potentially enhancing benthic secondary
production or carrying capacity; 2) benthic suspension feeders may mitigate adverse
effects of eutrophication like increased turbidity and oxygen depletion in bottom
waters by grazing on phytoplankton cells and thus increase water transparency and
diminish sedimentation; and 3) by suspension feeder removal of phytoplankton cells,
stored nutrients will to some extent become readily available to production through
excretion and may thus alter plankton composition.
Food limitation and carrying capacity
It can intuitively be appreciated that on the level of the patch of benthic suspension
feeders, food limitation is predominantly a result of physically mediated food supply.
If turbulence, vertical mixing and horizontal advection are sufficiently high,
competitive interactions will be of minor importance and the size of the patch will
depend on the physical forcing. On a system scale, benthic suspension feeder biomass
and production will, however, be limited of both physical constraints and system
(primary) productivity (Heip et al. 1995) or productivity of adjacent systems.
There are various sources of evidence for the tight coupling between system primary
productivity and secondary production of benthic suspension feeders. It has thus been
shown that biomass of both overall zoobenthos (Beukema & Cadée 1986, Beukema et
al. 2002) - and specifically suspension feeder (Beukema & Cadée 1997) - biomass
and production could be related to pelagic phytoplankton biomass and production. It
could further be shown that growth of a dominant suspension feeding bivalve was
related to plankton cell concentration during the growing season (Beukema & Cadée
1986, Beukema & Cadée 1991, Beukema et al. 2002). Similarly, a tight correlation
between fluctuations in annual primary production and mussel growth, indicated by
condition index, was demonstrated in one of the main mussel production sites in
Holland, the Oosterschelde (Smaal & van Stralen 1990, Smaal et al. 2001). Benthic
secondary production has also been correlated directly to nutrient supply in e.g. Ria
de Arosa, where growth and condition of cultivated mussels are closely related to the
annual average upwelling index (Blanton et al. 1987), indicating the degree of
nutrients transported from deep Atlantic waters to the Galician estuaries through
upwelling. Further, Josefson & Rasmussen (2000) demonstrated dependency of
benthic biomass on nutrient load from land run-off (fig. 5), with increased nutrient
loading resulting in increased biomass of particularly suspension feeders that
constituted >90% of the biomass. Above a nutrient load of approximately 125 g Ntot
m-2 yr-1 biomass decreased, probably due to increased occurrence of oxygen depletion
or persistent stratification.
The relation was derived for a number of different estuaries and was further
improved, if residence time of the estuaries were taken into consideration. On an
ecosystem level, carrying capacity can be defined as the population size that can be
supported by a given system and will depend on water residence time and primary
J. K. Petersen
production of the system in relation to the suspension feeder biomass and its
clearance capacity (Dame & Prins 1998).
Chl. a depletion, µg l -1 cm-1
Figure 4. Chlorophyll a depletion (slope of the linear relation between depth and
concentration of chl. a) as a function of hours of mixed conditions before sampling. Negative xvalues indicate hours of mixed conditions and positive x-values indicate hours of stratified
conditions prior to sampling (Petersen et al unpublished data)
A number of different coastal ecosystems have been compared with regard to these
parameters (Smaal & Prins 1993, Heip et al. 1995, Dame & Prins 1998). The systems
encompass a huge range in physical structure with regard to area, average depth,
volume, tidal range and residence time and primary production. In general the
systems fell in a range from small fast systems with short residence time supporting
denser populations to larger systems with longer residence times and lower biomass
(Dame & Prins 1998). A surprisingly good correlation between volume specific
biomass of benthic suspension feeders and water residence time of a number of
primarily tidal estuaries could thus be demonstrated (Smaal & Prins 1993, Heip et al.
1995). Including variables like primary production time, defined as annually averaged
phytoplankton biomass to primary production, and clearance time, defined as the time
needed for the suspension feeder population to clear the entire water column, some of
the systems were outliers (Dame & Prins 1998).
Grazing on pelagic primary producers
Figure 5. Plot of mean total biomass (AFDW) of benthic biomass against total area specific Nload for 14 Danish estuaries (From Josefson & Rasmussen 2000, printed with permission from
Outliers were systems with low primary production time due to e.g. high turbidity
like the Bay of Marennes-Oléron, or systems with low clearance time in relation to
primary production time due to e.g. large areas without benthic suspension feeders as
in (the present) Chesapeake Bay. When recalculating the data by Josefson &
Rasmussen (2000), it can further be seen that shallow Danish estuaries sustain a
higher volume specific biomass of benthic suspension feeders than what can be
inferred from the relation found by Heip et al. (1995). Both the outliers and the
Danish estuaries point to the same conclusion. Carrying capacity on the system level
will depend on both a biological component, i.e. the primary production, and a
physical component, that cannot be reduced to residence time, but must include some
measure of contact rate between the water column and the suspension feeding
populations. The physical contact rate will depend on mixing characteristics like
current speed and bottom roughness, but also on wind regime and bathymetry i.e. the
ratio of shallow shoals to deeper channels. In addition, contact rate will depend on the
physical placement of the suspension feeders, whether as a normal benthic
population, attached to macro-algae or seagrasses or cultivated on suspended culture
A special feature of some systems is that clearance time is shorter than primary
production time or water residence time. Thereby it is indicated that for the
populations to be sustained, food must be supplied either by advection or from
J. K. Petersen
unaccounted food sources. Resuspended material has been shown to be of importance
to benthic suspension feeder growth (Emerson 1990, Smaal & Zurburg 1997, Coma
et al. 2001). The resuspended material may be composed of epibenthic micro-algae
and detritus including material of macro-phytobenthic origin and may potentially
constitute a major part of the food supply to benthic suspension feeders (Héral 1993).
For budget calculations and in relation to primary production time and thus carrying
capacity of estuaries it is worthwhile noting, that only rarely is epibenthic primary
production included in estimates of system primary production. The contribution of
non-phytoplankton material to suspension feeder diet and how this relates to water
column mixing and primary production must be included in carrying capacity models
for specific areas that are commercially exploited.
Benthic suspension feeders as eutrophication control
In the comparison of different ecosystems mentioned above, some of the ecosystems
had lower clearance time compared to water residence time (Dame 1996). The
implication of this is not only that food for the benthic suspension feeders potentially
have to originate from outside the system, but also that the benthic suspension feeders
potentially regulate system phytoplankton biomass and production. If in addition,
clearance time is approximately equal to primary production time, then the systems
will be even more susceptible to benthic control (Dame & Prins 1998, Prins et al.
Evidence of top-down control by benthic suspension feeders under eutrophic
conditions has been demonstrated in various ways. In a series of experiments with
land-based, stirred mesocosms (Prins et al. 1995, Prins et al. 1999, Escaravage &
Prins 2002) it could thus be shown that even under conditions with high external
nutrient load, mussel grazing may exert an effective control on phytoplankton
biomass when mussel clearance capacity corresponds to 20-35% of the volume of the
mesocosms per day, but that structuring effects were negligible at a clearance
capacity of 15% or less (Prins et al. 1995). An important point of the experiments was
that with a strong top-down control, the ecosystem was independent of increased
nutrient loading and phytoplankton biomass remained low. A similar conclusion was
pointed out by Herman & Scholten (1990) in a modelling study, showing that
increasing nutrient loading under high suspension feeding pressure only marginally
changed phytoplankton concentration, implying that benthic grazing can act as
eutrophication control. As a consequence thereof, substantial changes in suspension
feeder biomass can be expected to have dramatic consequences for water column
parameters, notably chlorophyll a. In the absence of benthic suspension feeders,
grazing on phytoplankton will be performed by zooplankton. The developmental lagphase of zooplankton renders them, however, less efficient as phytoplankton control
(Herman & Scholten 1990, Prins et al. 1998). The lower metabolic rates of benthic
suspension feeders in contrast enable them to survive periods of low food, and they
can thus be present and ready to ingest when phytoplankton blooms.
On the scale of estuaries, studies are few but invasion of non-indigenous species has
been shown to reduce average phytoplankton concentrations dramatically (Nichols
1985, Alpine & Cloern 1992). In the northern San Francisco Bay mean chlorophyll a
Grazing on pelagic primary producers
concentration was normally dependent on mean river flow, but in 1976-77 and again
since 1986 that relation was no longer valid and chlorophyll a concentrations were
dramatically reduced (Alpine & Cloern 1992, Cloern 1996). In both cases the
reduction could be related to invasion of clams, in 1976-77 primarily Mya arenaria
and in 1986 by Potamocorbula amurensis, since neither nutrient nor light limitation
could explain the decrease. On a shorter time scale, breakdown of stratification has
been shown to lead to decrease in phytoplankton biomass (Cloern 1991, Koseff et al.
1993, Møhlenberg 1995). Similarly, decoupling of benthic grazing due to die off of a
large part of the benthic suspension feeders (Riisgård et al. 1995) or to stratification
(Cloern 1991, Koseff et al. 1993, Møhlenberg 1995) result in increase in
phytoplankton biomass. And it has been shown that biomass of mussels is a better
predictor of chlorophyll a concentration in Danish estuaries than any quantity related
to nutrient fluxes and concentrations (Kaas et al. 1996). Chlorophyll a was thus
inversely related to mussel biomass in these estuaries, which also could be
demonstrated in 15 Canadian estuaries (Meeuwig 1999).
Data on suspension feeder control of phytoplankton biomass and production
primarily originate from well-mixed shallow ecosystems with high benthic biomass.
These conditions along with requirements of no nutrient or light limitation of
phytoplankton production are the basic assumptions of the Officer et al. (1982) model
that use an idealized Lotka-Volterra approach of population interactions:
τg = H/RB = τp = 1/µp
where τg is the grazing and τp is production time scale, H is the system water depth in
m, R is specific clearance rate of the benthic suspension feeders in m3 g-1 d-1, B is
biomass of benthic suspension feeders in g m-2 and /µp is specific phytoplankton net
growth rate in d-1. Equilibrium between benthic suspension feeders and
phytoplankton is reached when the grazing time-scale equals the production timescale, which in the case of South San Francisco Bay can be reached with a biomass of
suspension feeders of a few hundred g m-2 resulting in a phytoplankton biomass of a
few µg chlorophyll a l-1. This is a simple but realistic situation for many well-mixed
estuaries (Herman et al. 1999). Focusing on phytoplankton bloom formation in the
presence of benthic suspension feeders, Koseff et al. (1993) showed that vertical
mixing of the water column is a critical process, which was not accounted for by
Officer et al. (1982). In non-stratified waters with ample nutrient and light
availability, the possibility of a bloom will depend on the ratios of the production
time-scale to the grazing time-scale and to the vertical (turbulent) mixing time-scale
(Koseff et al. 1993). Benthic suspension feeders can thus prevent phytoplankton
blooms if mixing is vigorous and grazing potential is sufficiently high compared to
primary production. In a more detailed, but qualitatively similar model, Lucas et al.
(1998) showed that even a mild form of permanent stratification affects the system
and renders benthic grazing unimportant in predicting bloom formation. Benthic
control is thus confined to shallow areas, though it may be significant down to 15 m
depending on actual conditions of light and stratification (Lucas et al. 1998). Bloom
formation is, however, not only dependent on the local production-loss balance but is
also affected by large-scale horizontal transport mechanisms (Lucas et al. 1999a,
Lucas et al. 1999b).
J. K. Petersen
Apart from the physical mixing, an important prerequisite for benthic suspension
feeders to act as eutrophication control is the open nature of the estuarine systems. If
there are no effective removal mechanisms, there will be an ever-increasing level of
phytoplankton when nutrients accumulate in a system, despite high grazing pressure
(Heip et al. 1995). Removal can be mediated trough different mechanism like
removal of excretion products by tidal currents, in the form of temporarily stored
nutrients released during autumn and winter (Heip et al. 1995) or by harvesting the
suspension feeders. In the highly eutrophic Mariager Fjord empirical modelling based
on monitoring data has shown that phosphorous has become limiting for primary
production during spring and early summer, while nitrogen is limiting in late summer
and fall (Nordjyllands Amt & Århus Amt 2002). The local authorities have set local
goals to reduce nutrient loading beyond the national goals in order to minimize the
risk for major anoxia events. Since additional limiting of nutrient load is difficult for
technical, economical or political reasons, the local authorities have promoted a plan
for cultivating blue mussels, Mytilus edulis, with the aim of removing nutrients. With
an annual harvest of 6,000 t live mussels, the local goal can be reached, showing that
cultivating suspension feeders can be a realistic management tool in mitigating effects
of eutrophication.
Nutrient feed-back and ecosystem structuring
Benthic suspension feeders not only remove phytoplankton and other particulate
organic matter from the water column. Specific organic constituents in the form of
phytoplankton cells and detritus consumed as food are processed and deposited as
faeces and pseudofaeces or excreted as fundamental nutrients. As an example, Jordan
& Valiela (1982) made a nitrogen budget for the ribbed mussel, Geukensia demissa,
in a salt marsh. They showed that at the level of the marsh-estuarine ecosystem, the
mussels filtered 1.8 times the particulate nitrogen exported yearly from the marsh by
tidal flushing. Of the N filtered from the water column, approximately half was
absorbed and half was deposited as faeces or pseudofaeces. Of the absorbed nitrogen,
55% was excreted as ammonia. It was calculated that the marsh mussels released
more ammonia into the system than any other component and that since nitrogen
limits the productivity of the marsh, re-mineralization by the mussels may ultimately
increase productivity. In the Askö area of the Baltic Sea, Kautsky & Wallentinus
(1980) measured in situ excretion of blue mussels and found that the regenerated
nutrients constitute a substantial input to both the pelagic and benthic system. The
excretion from mussels could explain why many of the benthic red and brown macroalgae could extend their growth maximum after nutrients are depleted in the water
mass. Thus, on one side mussels increase retention of nutrients within an estuarine
system, by filtering particulate material and depositing it as biodeposits (Kautsky &
Evans 1987) or long-lived mussel biomass. On the other side nutrients stored in
particulate planktonic material are transformed to readily accessible inorganic
nutrients by being processed by suspension feeders, thereby fuelling and thus
promoting primary production (Asmus & Asmus 1991).
Grazing on pelagic primary producers
Figure 6. Amount of nitrogen mineralized by mussel beds in central Oosterschelde compared
to total benthic and pelagic nitrogen mineralization (median and 10-90% range) from the
Oosterschelde ecosystem model SMOES (from Prins & Smaal 1994).
Apart from laboratory studies (e.g. Bayne 1976), excretion and regeneration of
nutrients have been studied in mesocosms (e.g. Doering et al. 1987) and in situ in
chambers or plastic bags (e.g. Kautsky & Wallentinus 1980, Jordan & Valiela 1982)
and in tunnels or flumes (e.g. Asmus et al. 1992). An extensive review can be found
in Dame (1996). In general, suspension feeders excrete ammonia and phosphate,
whereas mineralization of biodeposits leads to release of nitrate, nitrite and silicate
(Prins & Smaal 1994). Direct excretion can account for 17-94% of total ammonia and
37-85% of total phosphate release from mussel beds (Prins & Smaal 1994) and in
stable mussel and oyster beds, nitrogen release approximately balances input,
whereas a part of the phosphorous apparently is retained (Prins et al. 1998). The
importance of nutrient cycling by suspension feeders to phytoplankton production
will depend on general nutrient conditions in the estuary and mixing conditions, but
when turnover time of nutrients are lower than water retention time of the estuary, it
is probable that it is significant to the system (Dame 1996). Using model runs, Prins
& Smaal (1994) made an estimate of the amount of nitrogen mineralized by mussel
beds compared to total system mineralization (fig. 6) and found that it constitutes
approximately 50%. In addition, below suspended bivalve culture units, the organic
J. K. Petersen
enrichment of the sediments can lead to anaerobic conditions and thereby promote
denitrification (Kaspar et al. 1985). Bivalve cultures could thus function as a sink of
nitrogen not only through harvest of the bivalves but also due to loss of nitrogen to
the atmosphere.
Grazing by suspension feeders and the accompanying nutrient release predominantly
as easily accessible nutrients can be expected to alter phytoplankton species
composition and succession patterns. In general it can be expected that faster growing
small algal species will be favoured (Furnas 1990, Riegman et al. 1993) from grazing,
but also enhanced light availability in the water column from decreased POM
concentrations can be expected to influence phytoplankton composition and turnover
(Dame 1996). Increased phytoplankton growth (Doering et al. 1987) and shifts in
community composition towards smaller, faster growing species (Riemann et al.
1988, Olsson et al. 1992, Graneli et al. 1993, Prins et al. 1995) have been
demonstrated in mesocosm experiments. In the field, Noren et al. (1999)
demonstrated changes in phytoplankton composition in a water mass passing a
mussel bed, but field experiments from marine areas are few. Experiences from
freshwater systems are more abundant and invasion of zebra mussels has been shown
to cause complex changes in phytoplankton species composition and community
structure (e.g. Fahnenstiel et al. 1995a, Lavrentyev et al. 1995, Smith et al. 1998,
Rutherford et al. 1999) though results are not identical between systems with regard
to e.g. post-dreissenid dominance of cyanobacteria (Smith et al. 1998).
There is also evidence of effects of benthic grazing on the zooplankton community,
although more scarce. Effects may be mediated through competition for food or by
direct grazing of benthic suspension feeders on zooplankton. Mesocosm experiments
have shown that micro-zooplankton populations can be directly affected by bivalve
grazing (Horsted et al. 1988, Sullivan et al. 1991, Prins et al. 1995) and some
indications of ingestion of copepods or their development stages by mussels have
been reported (Kimmerer et al. 1994, Davenport et al. 2000). Freshwater zebra
mussels have been shown to affect ciliates (Lavrentyev et al. 1995) and daphnids
(Horgan & Mills 1999, Idrisi et al. 2001) though results are not clear. Filtration and
ingestion of larger zooplankton forms must depend on their ability to escape inhalant
siphon currents (Titelman 2001, Jakobsen 2002) and to the degree these are interfered
by turbulence (Singarajah 1975).
In summary, we now have a strong sense that benthic suspension feeders can act as a
filter in relation to eutrophication of shallow estuaries (Cloern 2001). This sense
stems from a series of different types of approaches that despite their individual
strengths and weaknesses together are convincing in demonstrating the importance of
benthic suspension feeders in the flux of particulate matter and dissolved nutrients. In
addition to the methodological problems discussed, it is hopefully apparent from the
present review, that the knowledge of the ecological role of benthic suspension
feeders in estuaries are limited both geographically, historically and quantitatively. In
Grazing on pelagic primary producers
an extensive review of ecological importance of bivalves (Dame 1996, Dame & Prins
1998), 15 estuaries were listed, from which bivalve impact could be appreciated.
Generally speaking, the vast majority of information stem from a few selected and
not necessarily representative estuaries and definitely with an under representation of
micro-tidal areas. As our knowledge is limited geographically, it is also apparent that
most information was gathered up to around 1990 and that the last decade has
produced little and mostly review-like information, especially on the part dealing
with benthic grazing impact on phytoplankton biomass and production. There is thus
a need of a proliferation of studies expanding into other geographical areas with
different physical and chemical forcing, having other benthic suspension feeders than
mussels as key organism and using different scientific approaches, preferably asking
both the water and the animals. The existing literature gives ample proof that
changing locality or organism provides new insight.
Recommendations for future work can be summarized on 3 levels: 1) scientifically a
number of questions need to be addressed. The “α-factor”, i.e. the transformation of
potential grazing capacity to actual grazing in relation to turbulent conditions needs
further attention on all scales: individual, patch and estuary. Impact of benthic
grazing (and re-mineralization) on planktonic (including importance for fish
populations) and benthic biodiversity is at the moment very little investigated both
theoretically and practically. Food sources for benthic suspension feeders and the
relative importance of these under various mixing conditions also need further
attention, especially the importance of epibenthic micro-algae. Finally, the current
controversies on basic physiological rates like clearance and ingestion in relation to
POM concentration need to be resolved. 2) For monitoring, new approaches like
remote sensing of surface waters from above water or ships of opportunity and
modelling need to address physical mixing and benthic grazing in areas with dense
populations of benthic suspension feeders. 3) Managers of cultured or fished
populations of especially bivalves need to address environmental issues in order to be
able to maintain a sustainable exploitation. These issues are the community added
value from mussel cultures as means of mitigating effects of eutrophication, as well
as the damaging effects on the benthic environment below these units and the
possible cascading effects that changes in planktonic diversity potentially can have on
e.g. the populations and recruitment of commercial fish in estuaries. Further, when
using models as management tool, these should be carefully examined for how
physical forcing is taken into account.
Alpine, A. E., & Cloern, J. E. (1992). Trophic interactions and direct physical effects control
phytoplankton biomass and production in an estuary. Limnology and Oceanography, 37(5), 946-955.
Asmus, H., Asmus, R. M., Prins, T. C., Dankers, N., Frances, G., Maass, B., et al. (1992). Benthic-pelagic
flux rates on mussel beds: Tunnel and tidal flume methodology compared. Helgoländer
Meeresuntersuchungen, 46(3), 341-361.
Asmus, R. M., & Asmus, H. (1991). Mussel beds: limiting or promoting phytoplankton. Journal of
Experimental Marine Biology and Ecology, 148, 215-232.
Bayne, B. L. (1976). Marine mussels: their ecology and physiology. Cambridge, U.K.: Cambridge
University Press.
Beukema, J. J., & Cadée, G. C. (1986). Zoobenthos responses to eutrophication of the Dutch Wadden Sea.
Ophelia, 26, 55-64.
J. K. Petersen
Beukema, J. J., & Cadée, G. C. (1991). Growth rates of the bivalve Macoma balthica in the Wadden Sea
during a period of eutrophication: relationships with concentrations of pelagic diatoms and flagellates.
Marine Ecology Progress Series, 68, 249-256.
Beukema, J. J., & Cadée, G. C. (1997). Local differences in macrozoobenthic response to enhanced food
supply caused by mild eutrophication in a Wadden Sea area: Food is only locally a limiting factor.
Limnology and Oceanography, 42(6), 1424-1435.
Beukema, J. J., Cadée, G. C., & Dekker, R. (2002). Zoobenthic biomass limited by phytoplankton
abundance: evidence from parallel changes in two long-term data series in the Wadden Sea. Journal of
Sea Research, 48, 111-125.
Blanton, J. O., Tenore, K. R., Castillejo, F., Atkinson, L. P., Schwing, F. B., & Lavin, A. (1987). The
relationship of upwelling to mussel production in the rias on the western coast of Spain. Journal of
Marine Research, 45, 497-511.
Buss, L. W., & Jackson, J. B. C. (1981). Planktonic food availability and suspension-feeder abundance:
evidence of in situ depletion. Journal of Experimental Marine Biology and Ecology, 49, 151-161.
Butman, C. A., Fréchette, M., Geyer, W. R., & Starczak, V. R. (1994). Flume experiments on food supply
to the blue mussel Mytilus edulis L. as a function of boundary-layer flow. Limnology and
Oceanography, 39(7), 1755-1768.
Carlson, D. J., Townsend, D. W., Hilyard, A. L., & Eaton, J. F. (1984). Effect of an intertidal mudflat on
plankton of the overlying water column. Canadian Journal of Fisheries and Aquatic Sciences, 41,
Cloern, J. E. (1982). Does the benthos control phytoplankton biomass in South San Francisco Bay. Marine
Ecology Progress Series, 9, 191-202.
Cloern, J. E. (1991). Tidal stirring and phytoplankton bloom dynamics in an estuary. Journal of Marine
Research, 49, 203-221.
Cloern, J. E. (1996). Phytoplankton bloom dynamics in coastal ecosystems: a review with some general
lessons from sustained investigation of San Fransisco Bay, California. Reviews of Geophysics, 34(2),
Cloern, J. E. (2001). Our evolving conceptual model of the coastal eutrophication problem. Marine
Ecology Progress Series, 210, 223-253.
Coma, R., Ribes, M., Gili, J.-M., & Hughes, R. N. (2001). The ultimate opportunists: consumers of seston.
Marine Ecology Progress Series, 219, 305-308.
Conley, D. J., Kaas, H., Møhlenberg, F., Rasmussen, B., & Windolf, J. (2000). Characteristics of Danish
estuaries. Estuaries, 23(6), 820-837.
Dame, R. F. (1996). Ecology of marine bivalves. An ecosystem approach. Boca Raton: CRC Press.
Dame, R. F., & Prins, T. C. (1998). Bivalve carrying capacity in coastal ecosystems. Aquatic Ecology, 31,
Davenport, J., Smith, R. J. J. W., & Packer, M. (2000). Mussels Mytilus edulis: significant consumers and
destoyers of mesozooplankton. Marine Ecology Progress Series, 198, 131-137.
Doering, P. H., Kelly, J. R., Oviatt, C. A., & Sowers, T. (1987). Effect of the hard clam Mercenaria
mercenaria on benthic fluxes of inorganic nutrients and gases. Marine Biology, 94, 377-383.
Dolmer, P. (2000a). Algal concentration profiles above mussel beds. Journal of Sea Research, 43(2), 113119.
Dolmer, P. (2000b). Feeding activity of mussels Mytilus edulis related to near-bed currents and
phytoplankton biomass. Journal of Sea Research, 44(3-4), 221-231.
Duarte, C. M., Benavent, E., & Sánchez, M. d. C. (1999). The microcosm of particles within seagrass
Posidonia oceanica canopies. Marine Ecology Progress Series, 181, 289-295.
Duren, L. A. v., Hendriks, I., Petersen, J. K., Larsen, J., Sikkelerus, P. B. C. v., Herman, P. M. J., et al.
(2002). Combined flume and field experiments on boundary layer hydrodynamics over mussel beds.
Paper presented at the ASLO Conference, June 2002, Victoria BC.
Emerson, G. W. (1990). Influence of sediment disturbance and water flow on the growth of the soft-shell
clam, Mya arenaria L. Canadian Journal of Fisheries and Aquatic Sciences, 47, 1655-1663.
Escaravage, V., & Prins, T. C. (2002). Silicate availability, vertical mixing and grazing control of
phytoplankton blooms in mesocosms. Hydrobiologia, 484, 33-48.
Fahnenstiel, G. L., Bridgeman, T. B., Lang, G. A., McCormick, M. J., & Nalepa, T. F. (1995).
Phytoplankton productivity in Saginaw Bay, Lake Huron: effects of zebra mussel (Dreissena
polymorpha) colonization. Journal of Great Lakes Research, 21(4), 465-475.
Fahnenstiel, G. L., Lang, G. A., Nalepa, T. F., & Jahnengen, T. H. (1995). Effects of zebra mussel
(Dreissena polymorpha) colonization on water quality parameters in Saginaw Bay, Lake Huron.
Journal of Great Lakes Research, 21(4), 435-448.
Grazing on pelagic primary producers
Fréchette, M., Aitken, A. E., & Pagé, L. (1992). Interdependence of food and space limitation of a benthic
suspension feeder: consequences for self-thinning relationships. Marine Ecology Progress Series, 83,
Frechette, M., & Bourget, E. (1985). Energy flow between the pelagic and benthic zones: Factors
controlling particulate organic matter available to an intertidal mussel (Mytilus edulis) bed. Canadian
Journal of Fisheries and Aquatic Sciences, 42(6), 1158-1165.
Fréchette, M., & Bourget, E. (1985). Food-limited growth in Mytilus edulis L. in relation to the benthic
boundary layer. Canadian Journal of Fisheries and Aquatic Sciences, 42, 1166-1170.
Frechette, M., Butman, C. A., & Geyer, W. R. (1989). The importance of boundary-layer flows in
supplying phytoplankton to the benthic suspension feeder, Mytilus edulis L. Limnology and
Oceanography, 34(1), 19-36.
Fréchette, M., & Lefaivre, D. (1990). Discriminating between food and space limitation in benthic
suspension feeders using self-thinning relationships. Marine Ecology Progress Series, 65, 15-23.
Furnas, M. J. (1990). In situ growth of marine phytoplankton: approaches to measurement, community and
species growth rate. Journal of Plankton Research, 12(6), 1117-1151.
Gacia, E., Duarte, C. M., & Middelburg, J. J. (2002). Carbon and nutrient deposition in a Mediterranean
seagrass (Posidonia oceanica) meadow. Limnology and Oceanography, 47(1), 23-32.
Graneli, E., Olsson, P., Carlsson, P., Graneli, W., & Nylander, C. (1993). Weak "top-down" control of
dinoflagellate growth in the coastal Skagerrak. Journal of Plankton Research, 15(2), 213-237.
Haamer, J., & Rodhe, J. (2000). Mussel Mytilus edulis (L.) filtering of the Baltic Sea outflow through the
Oeresund - an example of a natural, large-scale ecosystem restoration. Journal of Shellfish Research,
19(1), 413-421.
Heip, C. H. R., Goosen, N. K., Herman, P. M. J., Kromkamp, J., Middelburg, J. J., & Soetaert, K. (1995).
Production and consumption of biological particles in temperate tidal estuaries. Oceanography and
Marine Biology: an Annual Review, 33, 1-149.
Héral, M. (1993). Why carrying capacity models are useful tools for management of bivalve molluscs
culture. In R. F. Dame (Ed.), Bivalve filter feeders in estuarine and coastal ecosystem processes (pp.
421-454). Berlin: Springer.
Herman, P. M. J., Middelburg, J. J., van de Koppel, J., & Heip, C. H. R. (1999). Ecology of Estuarine
macrobenthos. Advances in Ecological Research, 29, 195-240.
Herman, P. M. J., & Scholten, H. (1990). Can suspension-feeders stabilise estuarine ecosystems. . In M. A.
Barnes & R: N: Gibson (Eds), Trophic relations in the marine environment (pp 104-116). Aberdeen
University Press.
Hily, C. (1991). Is the activity of benthic suspension feeders a factor controlling water quality in the Bay of
Brest. Marine Ecology Progress Series, 69, 179-188.
Horgan, M. J., & Mills, E. L. (1999). Zebra mussel filter feeding and food-limited production of Daphnia:
recent changes in lower trophic level dynamics of Oneida Lake, New York, U.S.A. Hydrobiologia,
411, 79-88.
Horsted, S. J., Nielsen, T. G., Riemann, B., Pock-Steen, J., & Bjørnsen, P. K. (1988). Regulation of
zooplankton by suspension-feeding bivalves and fish in estuarine enclosures. Marine Ecology
Progress Series, 48, 217-224.
Idrisi, N., Mills, E. L., Rudstam, L. G., & Stewart, D. J. (2001). Impact of zebra mussels (Dreissena
polymorpha) on the pelagic lower trophic levels of Oneida Lake, New York. Canadian Journal of
Fisheries and Aquatic Sciences, 58, 1430-1441.
Jakobsen, H. H. (2002). Escape of protists in predator-generated feeding currents. Aquatic Microbial
Ecology, 23(6), 271-281.
Jordan, T. E., & Valiela, I. (1982). A nitrogen budget for the ribbed mussel, Geukensia demissa, and its
significance in nitrogen flow in a New England salt marsh. Limnology and Oceanography, 27(1), 7590.
Josefson, A. B., & Rasmussen, B. (2000). Nutrient retention by benthic macrofaunal biomass of Danish
estuaries: Importance of nutrient load and residence time. Estuarine, Coastal and Shelf Science, 50,
Kaas, H., Møhlenberg, F., Josefson, A., Rasmussen, B., Krause-Jensen, D., Jensen, H. S., et al. (1996).
Marine areas. Danish inlets - state of the environment, trends and causal relations. The monitoring
programme under the Action Plan for the Aquatic Environment 1995 (in Danish) (Faglig rapport fra
DMU No. 179): Minestry of Environment and Energy, National Environmental Research Institute.
Kaspar, H. F., Gillespie, P. A., Boyer, I. C., & MacKenzie, A. L. (1985). Effects of mussel aquaculture on
the nitrogen cycle and benthic communities in Kenepuru Sound, Marlborough Sounds, New Zealand.
Marine Biology, 85, 127-136.
J. K. Petersen
Kautsky, N., & Evans, S. (1987). Role of biodeposition by Mytilus edulis in the circulation of matter and
nutrients in a Baltic coastal ecosystem. Marine Ecology Progress Series, 38, 201-212.
Kautsky, N., & Wallentinus, I. (1980). Nutrient release from a Baltic Mytilus-red algal community and its
role in benthic and pelagic productivity. Ophelia, Suppl. 1, 17-30.
Kimmerer, W. J., Gartside, E., & Orsi, J. J. (1994). Predation by an introduced clam as the likely cause of
substantial declines in zooplankton of San Fransisco Bay. Marine Ecology Progress Series, 113, 8193.
Koseff, J. R., Holen, J. K., Monismith, S. G., & Cloern, J. E. (1993). Coupled effects of vertical mixing and
benthic grazing on phytoplankton populations in shallow, turbid estuaries. Journal of Marine
Research, 51, 843-868.
Larsen, P. S., & Riisgård, H. U. (1997). Biomixing generated by benthic filter feeders: a diffusion model
for near-bottom phytoplankton depletion. Journal of Sea Research, 37, 81-90.
Lavrentyev, P. J., Gardner, W. S., Cavaletto, J. F., & Beaver, J. R. (1995). Effects of the zebra mussel
(Dreissena polymorpha Pallas) on protozoa and phytoplankton from Saginaw Bay, Lake Huron.
Journal of Great Lakes Research, 21(4), 545-557.
Loo, L.-O., & Rosenberg, R. (1989). Bivalve suspension-feeding dynamics and benthic-pelagic coupling in
an eutrophicated marine bay. Journal of Exprimental Marine Biology and Ecology, 130, 253-276.
Loo, L.-O., & Rosenberg, R. (1996). Production and energy budget in marine suspension feeding
populations: Mytilus edulis, Cerastoderma edule, Mys arenaria and Amphiura filiformis. Journal of
Sea Research, 35(1-3), 199-207.
Lozano, S. J., Scharold, J. V., & Nalepa, T. F. (2001). Recent declines in benthic macroinvertebrate
densities in Lake Ontario. Canadian Journal of Fisheries and Aquatic Sciences, 58, 518-529.
Lucas, L. V., Cloern, J. E., Koseff, J. R., Monismith, S. G., & Thompson, J. K. (1998). Does the Sverdrup
critical depth model explain bloom dynamics in estuaries? Journal of Marine Research, 56, 375-415.
Lucas, L. V., Koseff, J. R., Cloern, J. E., Monismith, S. G., & 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., & Thompson, J. K. (1999b). Processes
governing phytoplankton blooms in estuaries. II: The role of horizontal transport. Marine Ecology
Progress Series, 187, 17-30.
Meeuwig, J. J. (1999). Predicting coastal eutrophication from land-use: an empirical approach to small
non-stratified estuaries. Marine Ecology Progress Series, 176, 231-241.
Muschenheim, D. K. (1987). The dynamics of near-bed seston flux and suspension-feeding benthos.
Journal of Marine Research, 45(2), 473-496.
Muschenheim, D. K., & Newell, C. R. (1992). Utilization of seston flux over a mussel bed. Marine
Ecology Progress Series, 85, 131-136.
Møhlenberg, F. (1995). Regulating mechanisms of phytoplankton growth and biomass in a shallow estuary.
Ophelia, 42(1), 239-256.
Møhlenberg, F., & Petersen, J. K. (1998). Realised grazing impact by inter- and sub-tidal mussel
populations in the Oosterschelde. Paper presented at the ELOISE 2nd Annual Scientific Conference,
Huelva, Spain.
Möller, P., Pihl, L., & Rosenberg, R. (1985). Benthic faunal energy flow and biological interaction in some
shallow marine soft bottom habitats. Marine Ecology Progress Series, 27, 109-121.
Nalepa, T. F., & Fahnenstiel, G. L. (1995). Dreissena polymorpha in the Saginaw Bay, Lake Huron
ecosystem: overview and perspective. Journal of Great Lakes Research, 21(4), 411-416.
Newell, C. R. (1990). The effects of mussel (Mytilus edulis, Linnaeus, 1758) position in seeded bottom
patches on growth at subtidal lease sites in Maine. Journal of Shellfish Research, 9(1), 113-118.
Nichols, F. H. (1985). Increased benthic grazing: an alternative explanation for low phytoplankton biomass
in Northern San Fransisco Bay during the 1976-1977 drought. Estuarine, Coastal and Shelf Science,
21, 379-388.
Nordjyllands Amt & Århus Amt (2002). Debatoplæg om Mariager Fjord. Ålborg.
Noren, F., Haamer, J., & Lindahl, O. (1999). Changes in the plankton community passing a Mytilus edulis
mussel bed. Marine Ecology Progress Series, 191, 187-194.
O'Riordan, C. A., Monismith, S. G., & Koseff, J. R. (1995). The effects of bivalve excurrent jet dynamics
on mass transfer in a benthic boundary layer. Limnology and Oceanography, 40(2), 330-344.
Officer, C. B., Smayda, T. J., & Mann, R. (1982). Benthic filter feeding: A natural eutrophication control.
Marine Ecology Progress Series, 9, 203-210.
Olsson, P., Graneli, E., Carlsson, P., & Abreu, P. (1992). Structuring of a postspring phytoplankton
community by manipulation of trophic interactions. Journal of Experimental Marine Biology and
Ecology, 158(2), 249-266.
Grazing on pelagic primary producers
Petersen, J. K., Møhlenberg, F., Jonsson, P. R., Mathieu, P.-P., & Burchard, H. (2000). Grazing impact by
benthic suspension-feeders in a microtidal estuary. Paper presented at the ASLO Conference 2000,
Petersen, J. K., & Riisgård, H. U. (1992). Filtration capacity of the ascidian Ciona intestinalis and its
grazing impact in a shallow fjord. Marine Ecology Progress Series, 88, 9-17.
Petersen, J. K., Stenalt, E., & Hansen, B. W. (2002). Invertebrate re-colonisation in Mariager Fjord
(Denmark) after a severe hypoxia. II. Blue mussels (Mytilus edulis L.). Ophelia, 56(3), 215-226.
Prins, T. C., Escaravage, V., Smaal, A. C., & Peeters, J. C. H. (1995). Nutrient cycling and phytoplankton
dynamics in relation to mussel grazing in a mesocosm experiment. Ophelia, 41, 289-315.
Prins, T. C., Escaravage, V., Wetsteyn, L. P. M. J., Peeters, J. C. H., & Smaal, A. C. (1999). Effects of
different N- and P-loading on primary and secondary production in an experimental marine ecosystem.
Aquatic Ecology, 33, 65-81.
Prins, T. C., & Smaal, A. C. (1994). The role of the blue mussel Mytilus edulis in the cycling of nutrients
in the Oosterschelde estuary (The Netherlands). Hydrobiologia, 282-283(0), 413-429.
Prins, T. C., Smaal, A. C., & Dame, R. F. (1998). A review of the feedbacks between bivalve grazing and
ecosystem processes. Aquatic Ecology, 31, 349-359.
Riegman, R., Kuipers, B. R., Noordeloos, A. A. M., & Witte, H. J. (1993). Size-differential control of
phytoplankton and the structure of plankton communities. Netherlands Journal of Sea Research, 31(3),
Riemann, B., Nielsen, T. G., Horsted, S. J., Bjørnsen, P. K., & Pock-Steen, J. (1988). Regulation of
phytoplankton biomass in estuarine enclosures. Marine Ecology Progress Series, 48, 205-215.
Riisgård, H. U., Christensen, P. B., Olesen, N. J., Petersen, J. K., Møller, M. M., & Andersen, P. (1995).
Biological structure in a shallow cove (Kertinge Nor, Denmark) - Control by benthic nutrient fluxes
and suspension-feeding ascidians and jellyfish. Ophelia, 41, 329-344.
Riisgård, H. U., Jensen, A. S., & Jürgensen, C. (1998). Hydrography, near-bottom currents and grazing
impact of the benthic filter-feeding ascidian Ciona intestinalis in a Danish fjord. Ophelia, 49(1), 1-16.
Riisgård, H. U., Jürgensen, C., & Clausen, T. (1996). Filter-feeding ascidians (Ciona intestinalis) in a
shallow cove: implications of hydrodynamics for grazing impact. Journal of Sea Research, 35(4), 293300.
Riisgård, H. U., Poulsen, L., & Larsen, P. S. (1996). Phytoplankton reduction in near-bottom water caused
by filter-feeding Nereis diversicolor - implications for worm growth and population grazing impact.
Marine Ecology Progress Series, 141, 47-54.
Rutherford, E. S., Rose, K. A., Mills, E. L., Forney, J. L., Mayer, C. M., & Rudstam, L. G. (1999).
Individual-based model simulations of a zebra mussel (Dreissena polymorpha) induced energy shunt
on walleye (Stizostedion vitreum) and yellow perch (Perca flavescens) populations in Oneida Lake,
New York. Canadian Journal of Fisheries and Aquatic Sciences, 56, 2148-2160.
Singarajah, K. V. (1975). Escape reactions of zooplankters: effect of light and turbulence. Journal of the
Marine Biological Association of the United Kingdom, 55, 627-639.
Smaal, A. C., & Prins, T. C. (1993). The uptake of organic matter and the release of inorganic nutrients by
bivalve suspension feeder beds. In R. F. Dame (Ed.), Bivalve filter feeders in estuarine and coastal
ecosystem processes (pp. 273-298). Berlin: Springer.
Smaal, A. C., & van Stralen, M. R. (1990). Average annual growth and condition of mussels as a function
of food source. Hydrobiologia, 195, 179-188.
Smaal, A. C., van Stralen, M. R., & Schuiling, E. (2001). The interaction between shellfish culture and
ecosystem processes. Canadian Journal of Fisheries and Aquatic Sciences, 58(5), 991-1002.
Smaal, A. C., & Zurburg, W. (1997). The uptake and release of suspended and dissolved material by
oysters and mussels in Marennes-Oleron Bay. Aquatic Living Resources, 10(1), 23-30.
Smith, T. E., Stevenson, R. J., Caraco, N. F., & Cole, J. J. (1998). Changes in phytoplankton community
structure during the zebra mussel (Dreissena polymorpha) invasion of the Hudson River (New York).
Journal of Plankton Rsearch, 20(8), 1567-1579.
Sullivan, B. K., Doering, P. H., Oviatt, C. A., Keller, A. A., & Frithsen, J. B. (1991). Interactions with the
benthos alter pelagic food web structure in coastal waters. Canadian Journal of Fisheries and Aquatic
Sciences, 48, 2276-2284.
Terrados, J., & Duarte, C. M. (2000). Experimental evidence of reduced particle resuspension within a
seagrass (Posidonia oceanica L.) meadow. Journal of Experimental Marine Biology and Ecology, 243,
Titelman, J. (2001). Swimming and escape behaviour of copepod nauplii: implications for predator-prey
interactions among copepods. Marine Ecology Progress Series, 213, 203-213.
J. K. Petersen
Vedel, A. (1998). Phytoplankton depletion in the benthic boundary layer caused by suspension-feeding
Nereis diversicolor (Polychaeta): grazing impact and effect of temperature. Marine Ecology Progress
Series, 163, 125-132.
Wildish, D. J., & Kristmanson, D. D. (1979). Tidal energy and sublittoral macrobenthic animals in
estuaries. Journal of Fisheries Research Board of Canada, 36, 1197-1206.
Wildish, D. J., & Kristmanson, D. D. (1984). Importance to mussels of a benthic boundary layer. Canadian
Journal of Fisheries and Aquatic Science, 41, 1618-1625.
Wildish, D. J., & Kristmanson, D. D. (1997). Benthic suspension feeders and flow. New York: Cambridge
University Press.
Yu, N., & Culver, D. A. (2000). Can zebra mussels change stratification patterns in a small reservoir?
Hydrobiologia, 431, 175-184.
J.K. Petersen: Department of Marine Ecology, National Environmental Research
Institute, P.O. Box 358, DK-4000 Roskilde, Denmark.
The importance of herbivores as controls of nutrient storage as producer biomass
depends to a great extent on the percentage of primary production they consume.
This is so because the quantity of primary production that remains stored as producer
biomass decreases as consumption by herbivores increases. Hence, herbivores may
keep low levels of producer biomass through intense consumption (Carpenter 1986,
Valentine & Heck 1991, Cebrian & Duarte 1994). Similarly, the contribution of
herbivores to producer-bound nutrient recycling through faeces excretion and
exudation also increases with the percentage of primary production consumed.
Herbivore excreta are usually nutrient-rich in relation to producer detritus and thus
are quickly decomposed by detritivorous organisms. Thus, large percentages of
primary production consumed can significantly increase the rate of nutrient recycling
in the system (Elser & Urabe 1999, Sterner & Elser 2002).
The existing literature indicates that the intensity of herbivory on benthic producers
can be very variable. For instance, the percentage of primary production removed by
herbivores ranges from negligible values to ca. 100% in communities of
microphytobenthos (i.e. benthic microalgae that live in the sediment; Montagna 1984,
Baird & Ulanowicz 1993), macroalgae (Mann 1972, Hagen 1995, Ferreira et al.
1998) and seagrasses (Cebrian & Duarte 1998, Valentine & Heck 1999, Valentine et
al. 2000). It therefore seems that herbivores can play a variable role as controls of the
dynamics of producer biomass and nutrient recycling in coastal communities.
Unfortunately, there have been few comparisons of the variability in herbivory within
and across types of benthic producers (Valiela 1995, Duarte & Cebrian 1996, Alongi
1998, Cebrian 2002). In addition, most of the existing comparisons are not extensive.
This has impaired our ability to elucidate if there are any patterns in how herbivory
varies within and across types of benthic producers and to identify the controls and
consequences of that variability. It is clear that herbivores may play an important role
in the storage and trophic transference of carbon and nutrients in coastal systems, but
our knowledge of how much and why that role differs across communities of benthic
producers is poor.
There is another reason that justifies the importance of understanding the extent and
controls of herbivory across communities of benthic producers: the change in
dominant benthic producers that often results from increased anthropogenic
eutrophication. Coastal eutrophication is one of the most wide-spread, pervasive
human-induced environmental disturbances (Nixon 1995, Valiela 1995, Cloern
2001). It results from changes in land use that accompanies the occupation of coastal
areas by human populations. Urbanization and deforestation into agricultural and
farm-lands of the coastal watershed leads to higher rates of nutrient input into
receiving coastal waters. In turn, if seagrasses are the dominant producers in the
receiving waters, increased nutrient loading rates often lead to seagrass decline and
S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling:
The IInfluence of Primary Producers, 153-185.
© 2004 Kluwer Academic Publishers. Printed in the Netherlands.
the proliferation of loose and/or epiphytic macroalgae. The proximate cause for that
shift seems to be that macroalgae can generally take up water-column nutrients much
faster than seagrass can (Sand-Jensen & Borum 1991, Duarte 1995, Pedersen &
Borum 1997); thereby quickly building large stocks of biomass that can impose
severe shading and adverse biochemical conditions on seagrasses and lead to their
decline. There are numerous examples of the replacement of seagrasses by
macroalgae as dominant benthic producers following increased eutrophication (e.g.
Sfriso 1987, Reise et al. 1989, Rosenberg et al. 1990, Ciszewski et al. 1992,
Bombelli & Lenzi 1996, Niell et al. 1996, Nienhuis 1996, Hauxwell et al. 2003), and
that process is analysed at length in Chapter 3. Increased eutrophication also
promotes the accumulation of phytoplankton in the water-column, which can exert
substantial shading on benthic producers and further contribute to seagrass decline
(Valiela et al. 1992, 2000b, Taylor et al. 1995a, b, 1999). In addition, some recent
evidence suggests that large phytoplankton accumulations under high eutrophication
may also limit the growth of benthic macroalgae through intense light attenuation,
and even cause macroalgal decline (Funen Island Council 1991, Taylor et al. 1995a,
b, 1999, Duarte 1995, Schramm 1996). Thus, while it is well known that the
eutrophication of pristine environments often leads to the replacement of seagrasses
by macroalgae as dominant benthic producers, it seems that further eutrophication
reaching intense levels may lead to macroalgal decline and the replacement by
sediment flats dominated by benthic microalgae. Regardless of what outcomes may
occur under intense eutrophication, it is clear that increasing loading rates affect the
composition of assemblages of benthic producers and may shift the dominant types.
Hence, elucidating the extent and controls of the variability in herbivory among
different types of benthic producers is important to understand the changes in
herbivory that follow the replacement of dominant benthic producers under
increasing eutrophication and the consequences for coastal systems.
This chapter presents a summary of the extent and some of the controls and
consequences of herbivory on benthic primary producers under pristine conditions
(i.e. not affected by anthropogenic eutrophication). It also analyses how increasing
eutrophication, through the promotion of benthic algal growth and associated seagrass
decline, may affect the intensity of herbivory on benthic communities. The chapter is
thus divided in two parts. In the first part, using an extensive compilation of the
literature, I examine the variability in herbivory within and across pristine
communities of microphytobenthos, macroalgae and seagrasses. I also discuss some
of the controls and consequences of herbivory on benthic producers under pristine
conditions. In the second part, I first comment on several considerations as to how
the replacement of seagrasses by algae as dominant benthic producers following
increased eutrophication may affect herbivory. I then summarize what several
experiments comparing the response of macrofaunal densities to macroalgal blooms
following enhanced eutrophication have taught us. I finish by discussing the changes
in herbivory on benthic producers observed in two well-known ecosystems that have
undergone intense eutrophication over the past few decades, Waquoit Bay
(Massachusetts, USA) and the Venice Lagoon (Italy). The chapter focuses on
temperate systems, but reference to tropical systems is also made when pertinent and
possible. In choosing the tone of the chapter, I have intended to (1) review existing
information relevant to the goals of the paper, (2) present hypotheses regarding the
Grazing on benthic primary producers
control of herbivory in pristine communities and how eutrophication may affect
herbivory, and (3) suggest some recommendations for future research.
The question of whether herbivores are important controls of the biomass of benthic
producers (i.e. top-down control) has received considerable attention over the past
decades (Neckles et al. 1994, Valentine & Heck 1999, Heck et al. 2000, Worm et al.
2000, Lotze & Worm 2002). As a consequence, a wealth of reports quantifying
herbivory on communities of benthic microalgae, macroalgae, and seagrasses are now
available (for extensive reviews see Cebrian et al. 1998, Cebrian 1999, 2002). Those
reports tell us a clear message: the extent of herbivory on marine benthic producers is
highly variable, ranging from negligible levels to the removal of almost all available
producer biomass. For instance, meio- (<500µm) and macroherbivores (>500µm)
may remove a large percentage of the production of microphytobenthic communities
(Nicotri 1977, Baird & Ulanowicz 1993), but in other instances herbivory only
represents a minor loss for the community (Admiraal et al. 1983, Montagna 1984).
The extent of herbivory may even be more variable in communities of marine
macroalgae. Tropical communities often support intense consumption by herbivores
and, as a consequence, they are mainly composed of turf and/or unpalatable (i.e. with
chemical defences against herbivores) species. In fact, tropical macroalgal
communities are the template for most seminal examples of herbivore control of the
primary production and composition of marine communities (Samarco et al. 1977,
Hay 1983, Hay 1984, Hatcher & Larkum 1983, Morrison 1988). In contrast, it seems
that herbivores, by removing a smaller percentage of primary production, generally
play a lesser role on the dynamics of temperate macroalgal communities (Newell et
al. 1982, Alongi 1998). However, herbivory may also be intense in some temperate
macroalgal communities, such as kelp beds, where large grazing events leading to the
denudation of extensive kelp areas following the outburst of sea urchin populations
have been often documented (Lawrence 1975, Breen & Mann 1976, Dayton 1985,
Hagen 1995). Existing reports show a similarly broad range in herbivory on
seagrasses. Although many authors have found that herbivory accounts for a small
loss of seagrass production (Thayer et. al 1984, Mukai & Nojima 1985, Nienhuis &
Groenendijk 1986, Cebrian et al. 1996, Cebrian & Duarte 1998), herbivores may also
remove a large percentage of the production of tropical and temperate seagrass
communities (Greenway 1976, Keller 1983, Vermaat & Verhagen 1996, Valentine &
Heck 1999, Kirsch et al. 2002).
The wide range in herbivory found for communities of benthic microalgae,
macroalgae and seagrasses imply an important corollary; the role of herbivores as
controls of producer biomass dynamics and nutrient recycling, which largely depends
on the percentage of production removed, may differ greatly for each community
type. A logical question arises then: are there any patterns in that variability within
and across community types? In other words, does the extent of herbivory, and thus
the importance of herbivores as top-down controls of producer biomass and nutrient
recycling, vary in a certain way within and across community types? And if it does,
what are the factors responsible for those patterns? Answering these questions
requires extensive comparisons of representative ranges and frequencies of herbivory
values reported in the literature. Unfortunately, this kind of comparisons has been
scarce. Here, I capitalize on previous literature compilations (Cebrian et al. 1998,
Cebrian 1999, Cebrian 2002) and add recently published data to search for patterns in
the extent and control of herbivory within and between communities of
microphytobenthos, macroalgae and seagrasses.
The values compiled are
representative of the community considered (i.e. they integrate the most abundant
producers and herbivores in the community), correspond to natural conditions (i.e.
communities not deliberately impacted by human disturbances such as
eutrophication), and cover at least one year of observations or the growing season for
annual producers. The data set is available at under
"data sets".
The compilation done confirms previous results that herbivory on benthic producers
may be very variable. Indeed, herbivores may remove small or large percentages of
primary production in communities of microphytobenthos (Fig. 1A), macroalgae
(Fig. 1B) and seagrasses (Fig. 1C).
The results, however, reveal an obvious
difference between seagrasses and the two types of algal producers; the percentage of
production removed is strongly skewed to the right for seagrasses, whereas this is not
the case for microphytobenthos and macroalgae. Indeed most (>80%) of the seagrass
communities compiled lose a relatively modest percentage (<10%) of production to
herbivores (Fig. 1C). Thus, in spite of the large variability found within each
community type, communities of microphytobenthos and macroalgae tend to lose
higher percentages of primary production to herbivores than do seagrass communities
(non-parametric multiple comparison Q test; microphytobenthos vs. macroalgae,
P>0.10; microphytobenthos vs. seagrasses, P<0.01; macroalgae vs. seagrasses,
P<0.01). This difference indicates that in general herbivores, by removing a larger
percentage of primary production, play a greater role in the dynamics of producer
biomass and nutrient recycling in communities of microphytobenthos and macroalgae
than in seagrass communities. Accordingly, reports of significant top-down effects
by herbivores on benthic micro- and macro-algae are abundant (e.g. Littler et al.
1983, Hackney et al. 1989, Steneck & Dethier 1994, Bennett et al. 1999, Blanchard et
al. 2001), but much less so for seagrasses (Tribble 1981, Valentine & Heck 1991,
Heck & Valentine 1995).
As it is expected from the large variability found in the percentage of production
consumed, absolute herbivory also varies greatly within communities of
microphytobenthos (Fig. 1D), macroalgae (Fig. 1E) and seagrasses (Fig. 1F). For
each community type, absolute consumption by herbivores ranges over three orders
of magnitude (i.e. from <1 to ca. 1000 g C m-2 yr-1). However, and unlike the
percentage of production consumed, absolute consumption is only marginally smaller
in seagrass communities than in communities of microphytobenthos and macroalgae
(Kruskal-Wallis, P=0.06). In other words, in spite of the fact that seagrass
communities generally have smaller percentages of production removed by
herbivores than do communities of microphytobenthos and macroalgae, the absolute
flux of producer biomass channelled to herbivores varies little among the three
community types. The explanation for these results lies in the differences in primary
production found between the three communities. Seagrass communities tend to
reach larger levels of primary production than many benthic micro- and macroalgal
communities (Cebrian 2001).
Higher levels of primary production in seagrass
communities partially compensate for lower percentages of production removed by
Grazing on benthic primary producers
herbivores and, as a consequence, absolute consumption is only marginally smaller in
seagrass communities than in communities of benthic micro- and macroalgae. These
patterns suggest an important corollary; among communities of benthic producers,
production of herbivore biomass (as indicated by the absolute flux of producer
biomass transferred to the herbivore compartment) should not be as variable as the
extent of herbivore control of producer biomass and nutrient recycling (as indicated
by the percentage of primary production removed by herbivores).
Many factors can contribute to the large variability found in the extent of herbivory
on benthic producers. An extensive treatment of those factors is beyond the scope of
this chapter because the aim is to summarize information that can be instrumental in
predicting and understanding the changes in herbivory associated with the
replacement of dominant benthic producers due to eutrophication. Here I comment
on some factors that seem most relevant to that aim. One prominent factor that seems
responsible for large differences in herbivory is the ample temporal and spatial
variability that herbivore abundance often exhibits. Differences in recruitment
success (Valiela 1995, Mann 2000), predator abundance (Vadas & Steneck 1995,
Sala et al. 1998), degree of shelter in the habitat (Himmelman 1984, van Tamelen
1996) and frequency and duration of anoxic conditions (Sagasti et al. 2001, Gray
2002) may lead to large temporal and spatial differences in the abundance of
herbivores in benthic communities. In turn, fluctuating herbivore densities may entail
substantial variability in the extent of producer biomass consumed. Indeed, higher
herbivore abundances have been experimentally associated with larger levels of
consumption in communities of microphytobenthos (Admiraal et al. 1983, Blanchard
et al. 2001), macroalgae (Hauxwell et al. 1998, Ruesink 2000) and seagrasses
(Valentine & Heck 1991, Macia 2000). One of the best-known examples of control
of herbivore abundance by predation and resulting effects on producers is the
interaction between kelps, sea urchins (Strongylocentrotus sp.) and sea otters
(Enhydra lutris) reported in the North American Pacific Coast (Estes & Palmisano
1974, Foster & Schiel 1988). When sea urchins are abundant, they remove large
amounts of kelp biomass and may even denude extensive beds. However, under
intense predation by sea otters, sea urchin abundance remains low and kelps flourish.
Similar trophic cascade effects where the extent of herbivory is controlled by the
intensity of predation on the herbivore have also been reported in communities of
microphytobenthos (Posey et al. 1995, 1999) and seagrasses (Macintyre et al. 1987,
Jackson 1997, Valentine & Heck 1999).
Differences in the type of dominant herbivores in the community may also be
conducive to substantial variability in the intensity of herbivory among marine
benthic communities. For instance, in the seagrass communities where they are still
present, large vertebrate herbivores such as dugongs, manatees, green turtles and
waterfowl remove large quantities of seagrass and epiphyte biomass needed to
support the high metabolical demands imposed by their large body mass (Ogden et al.
1983, Lanyon et al. 1989, de Iongh et al. 1995, Masini et al. 2001).
Figure 1. The histograms of the percentage of primary production consumed by herbivores
(plots A-C) and absolute consumption by herbivores (plots D-F) in communities of
microphytobenthos, macroalgae and seagrasses. Dashed lines represent medians.
Similarly, green turtles and herbivorous fish, such as parrotfishes and surgeonfishes,
often consume large quantities of algal biomass in tropical beds and coral reef
communities (Hixon & Brostroff 1996, Musick & Limpus 1997). Consumption rates
of producer biomass per individual are far greater for large vertebrate herbivores than
for smaller invertebrate herbivores, and thus communities with different ratios of
vertebrate vs. invertebrate herbivore abundance should also differ in the intensity of
consumption of producer biomass. The feeding specificity of the dominant
Grazing on benthic primary producers
herbivores in the community may also contribute to the large variability in herbivory
observed among marine benthic communities. Some herbivores are general feeders
and consume a wide range of producer types, whereas some other herbivores are
more specific and strongly select a limited number of producers and/or target certain
parts of the producers. Sea urchins are often an example of extremely generalist
herbivores in seagrass meadows. They may consume seagrass blades and sheaths,
epiphytic micro- and macro-algae on the seagrass blades and rhizomes, and the
macroalgae that grow on bare patches interspersed in the meadow (Bergin 1987,
Klinger et al. 1994, Valentine & Heck 2001). On the contrary, dugongs select a
number of seagrass species (De Iongh et al. 1995, Preen 1995) and green turtles target
the basal part of turtlegrass (Thalassia testudinum) leaf blades (i.e. younger, more
tender parts, Thayer et al. 1984, Zieman et al. 1984). Similar examples exist for other
benthic communities. In sediment flats, harpacticoid copepods, nematods, ostracods
and other important groups of herbivores often differ greatly in their choice and use
of benthic microalgae as a food resource (Carman & Thistle 1985, Carman & Fry
2002). In temperate macroalgal communities, sea urchins often consume a variety of
host and epiphyte species (Bergin 1987, Verlaque 1987) whereas other invertebrate
grazers, such as isopods and amphipods, tend to be more selective (Salemaa 1987,
Pavia et al. 1999, Karez et al. 2000). Therefore, it can be expected that differences in
the relative abundance of generalist vs. selective herbivores and in the availability of
preferred producer diets may also induce substantial variability in the intensity of
herbivory among marine benthic communities.
Another factor that may also be responsible for differences in herbivory among
benthic communities is the nutrient content of the dominant producers. There is now
extensive experimental evidence that the growth rates of many herbivores are limited
by the nutrient content of their diets (Sterner & Hessen 1994, Elser et al. 2000a,
Sterner & Elser 2002, Stelzer & Lamberti 2002, Urabe et al. 2002). Many lab and
field manipulations have shown that aquatic and terrestrial herbivores that feed on
producers with enhanced nitrogen and phosphorus contents grow faster. The
rationale for these observations is that the growth rate of organisms is highly
dependent on the concentrations of RNA and enzymes in the cells (Elser et al. 1996,
2000b, Stener & Elser 2002). Hence, because RNA is rich in phosphorus and
enzymes are rich in nitrogen, higher nitrogen and phosphorus contents in the diet
allow for a larger production of RNA and enzymes and, thus, faster growth rates. In
turn, faster herbivore growth rates imply higher rates of biomass build-up and, thus,
they should be conducive to larger rates of consumption of producer carbon. In fact,
recent comparisons encompassing a wide range of community types, from marine to
freshwater to terrestrial, have shown that, at that large across-system scale,
communities composed of producers with higher nutrient contents support larger
rates of consumption by herbivores (Cebrian et al. 1998, Cebrian 1999).
Nevertheless, when the range of comparison is reduced to marine benthic
communities, the evidence for the association between higher producer nutrient
contents and faster consumption rates by herbivores seems equivocal. Most tests of
this association in marine benthic communities refer to seagrass communities because
the low nutrient contents of seagrass leaves relative to other benthic producers
(Duarte 1992, Nielsen et al. 1996) and the high fiber content in the leaves (Bjorndal
1980, Dawes & Lawrence 1980) have made many authors hypothesize that these
characteristics limit the intensity of herbivory on seagrasses. That hypothesis has
been confirmed for some vertebrate herbivores, such as the parrotfish, dugongs and
green sea turtles that select species or blade parts with higher nitrogen contents
(Bjorndal 1980, Zieman et al. 1984, Williams 1988, McGlathery 1995, Preen 1995).
However, recent papers comparing different seagrass species and herbivore
populations have found no association between leaf nitrogen content and herbivory
intensity (Cebrian & Duarte 1998, Mariani & Alcoverro 1999). Even more
discrepant is the fact that certain herbivores, such as the pink sea urchin (Lytechinus
variegatus), seem able to compensate for low nutritional quality in seagrass leaves
and consume larger quantities of the leaves with lower nitrogen contents (Valentine
& Heck 2001). Observations of the nutrient contents and consumption rates of
benthic micro-and macroalgae also indicate controversial results. For instance, in
coral reefs and other tropical algal communities most herbivores feed intensively on
small turf algae that normally have higher nutrient contents than most other algae
(Hay 1984, Russ & Alcala 1989, McClanahan et al. 1996). In contrast some
macroalgae that also have high nutrient contents suffer little or no herbivory due to
the presence of deterrent compounds (Hay & Fenical 1996, Hay 1997) or because of
restricted accessibility by herbivores (Carpenter 1986, Alongi 1998, Mann 2000).
To shed some light on whether a significant relationship between producer nutrient
contents and herbivory intensity exists for marine benthic communities, I have
compiled literature values of producer nitrogen and phosphorus contents and
consumption by herbivores in communities of microphytobenthos, macroalgae and
seagrasses. These values meet the same criteria as the values of absolute
consumption and percentage of primary production consumed compiled for Figure 1
(i.e. representative of the community considered, referring to natural conditions, and
encompassing one year of observations or the growing season for annual producers),
and are also available at under "data sets". The results
show that, when the three types of community are compared, larger percentages of
primary production consumed by herbivores tend to be associated with higher
nitrogen (Fig. 2A, Pearson correlation coefficient=0.40, P<0.01) and phosphorus
contents (Fig. 2B, Pearson correlation coefficient=0.49, P<0.01), with seagrasses
showing lower percentages consumed and nutrient contents in relation to
microphytobenthos and macroalgae. However, the association does not hold within
any of the community types; when communities of microphytobenthos, macroalgae
and seagrasses are examined separately, the percentage of production consumed is
independent of the nitrogen and phosphorus contents in the producer (P>0.10 for all
within-community Pearson correlation coefficients). Neither are there any significant
correlations between absolute herbivory and producer nitrogen and phosphorus
contents either across or within community types (Figs. 2C and D, P>0.10 for all
Pearson correlation coefficients). Taken together, these results convey a clear
message; the role of producer nutrient content as a control of herbivory in benthic
communities seems, at the best, modest. Significant, albeit not strong, associations
are found between the percentage consumed and producer nutrient contents, but not
with absolute herbivory. In addition, those associations only occur when different
types of benthic community are compared and a wide range of percentage (from <0.1
to ca. 100%) and nitrogen (from 1 to 10 %DW) and phosphorus (from 0.1 to 1 %DW)
values are encompassed. At scales smaller than comparing little consumed, nutrientpoor seagrass species with highly consumed, nutrient-rich macroalgae and
Grazing on benthic primary producers
microphytobenthos, it seems that factors other than producer nutrient contents are
responsible for the variability in herbivory.
Figure 2. The relationships between (A) the percentage of primary production consumed by
herbivores and nitrogen producer content, (B) the percentage of primary production consumed
by herbivores and phosphorus producer content, (C) absolute consumption by herbivores and
nitrogen producer content and (D) absolute consumption by herbivores and phosphorus
producer content in communities of microphytobenthos (open squares), macroalgae (open
circles) and seagrasses (filled circles). Data have been log-transformed to comply with the
assumptions of correlation analyses (see text for statistical results).
Despite all the many factors that can affect herbivory in marine benthic communities,
the absolute consumption of producer biomass remains closely associated with the
magnitude of primary production both across and within community types (Fig. 3A).
The compilation of published values shows high correlation coefficients across the
three types of community (Pearson correlation coefficient=0.81, P<0.01) and within
communities of microphytobenthos (Pearson correlation coefficient=0.96, P<0.01),
macroalgae (Pearson correlation coefficient=0.94, P<0.01) and seagrasses (Pearson
correlation coefficient=0.59, P<0.01). Thus, primary production appears as a good
predictor of the absolute flux of producer biomass transferred to herbivores in marine
benthic communities. The reason is entirely mathematical. Both across and within
types of community, the percentage of production consumed varies little in
comparison with the variability in primary production (Fig. 3B). For instance, across
community types the percentage consumed varies from 10 to ca. 100% for production
values comprised between 0.1 and 100 g C m-2 yr-1, and from 1 to ca. 100% for
production values between 100 and ca. 1000 g C m-2 yr-1. Thus, whereas primary
production ranges about 4 orders of magnitude across community types, the
percentage consumed only ranges about 2 orders of magnitude. Much larger ranges
in primary production than in the percentage consumed are also found within
communities of microphytobenthos, macroalgae and seagrasses. As a consequence,
absolute consumption, which corresponds to the product of primary production and
the percentage consumed, remains closely associated with primary production across
and within the three community types compared. Thus, regardless of the type of
marine benthic community compared, it seems that more productive communities
transfer a larger flux of producer biomass to herbivores and, thus, they should support
higher levels of herbivore production. Another corollary is that, since absolute
consumption is strongly associated with primary production but independent of
producer nutrient content, primary production and producer nutrient content should
also be independent in marine benthic communities. This is supported by the data
compiled (P>0.05 for all Pearson correlation coefficients between primary production
and nitrogen and phosphorus contents across and within community types) and by
other past reports (Cebrian et al. 1998, Cebrian 1999).
As mentioned above, there are now numerous observations that increasing
eutrophication of pristine shallow environments dominated by seagrasses often leads
to their replacement by loose and epiphytic macroalgae (e.g. Valiela et al. 1992,
1997a, 2000b, Short et al. 1995, Niell et al. 1996, Nienhuis 1996, Raffaelli et al.
1998, Hauxwell et al. 2001, 2003). These macroalgae are frequently species of
opportunistic green filamentous genera such as Enteromorpha, Chaetomorpha, Ulva
and Cladophora (Schramm & Booth 1981, Lapointe & O'Connell 1989, Jeffrey et al.
1992, McComb & Humphries 1992, Jeffrey 1993, Dion & Bosec 1996, Romero et al.
1996), although some other brown and red genera, such as Porphyra, Gracilaria,
Polysiphonia, Pilayella, and Ectocarpus (Fletcher 1996a, Peckol & Rivers 1995) may
also respond quickly to enhanced nutrient availability.
In some instances, however, factors such as unfavourable temperatures (Jeffrey et al.
1992, Nienhuis 1996), low light irradiance (Taylor et al. 1995a, b, 1999), exposure to
wind and wave action (De Vries et al. 1996, Raffaelli et al. 1998) and elevated
grazing (Neckles et al. 1994, Heck et al. 2000, Worm et al. 2000) may impair, and
even abort, the formation of macroalgal blooms under enhanced nutrient delivery.
The regulation of the appearance and duration of macroalgal blooms is not well
understood. It seems that increased nutrient availability is a necessary condition, but
not sufficient, for the build-up of macroalgal biomass (Raffaelli et al. 1998).
Grazing on benthic primary producers
Figure 3. The relationships between (A) absolute consumption by herbivores and primary
production and (B) the percentage of primary production consumed by herbivores and primary
production in communities of microphytobenthos (open squares), macroalgae (open circles)
and seagrasses (filled circles). Data have been log-transformed to comply with the
assumptions of correlation analyses (see text for statistical results).
Indeed, several authors have shown that blooms only occur when favourable
conditions of potentially limiting factors accompany enhanced nutrient availability
(Aubert 1990, DeVries et al. 1996, Fletcher 1996b, Poole & Raven 1997). The
nature and controls of macroalgal blooms are examined in Chapter 3 and also in
adequate recent reviews (Valiela et al. 1997a, Rafaelli et al. 1998, Cloern 2001).
Here, I focus on the effects that those blooms and subsequent seagrass decline may
have on the intensity of benthic herbivory. To do that, I first discuss some potential
implications of the replacement of seagrasses by macroalgae on local herbivore
populations and expected consequences on the intensity of herbivory. I then compare
those expectations with the results of several case studies and search for trends in the
response of herbivores to algal build-up and seagrass decline following
eutrophication. Results from various mesocosm experiments also suggest that intense
shading imposed by large phytoplankton accumulations under high eutrophication
may limit macroalgal blooms, and even lead to their replacement by sediments
dominated by benthic microalgae (Taylor et al. 1995a, b, Taylor et al. 1999). These
results obtained in laboratory mesocosms are yet to be demonstrated in field
conditions but, here, I also discuss how the replacement of macroalgal canopies by
bare sea beds inhabited by microphytobenthos could affect benthic herbivory.
3.1 Potential Effects of Shifting Assemblages of Benthic Producers on Herbivores
3.1.1 Changes in Producer Nutrient Content
Seagrass leaves tend to have low nutrient contents in relation to macroalgal tissues.
This was first showed in two extensive comparisons done in the last two decades
(Atkinson & Smith 1983, Duarte 1992) and I now confirm it with newly compiled
data (available at under data sets). As mentioned above,
the data compiled refer to the community level (i.e. include the most important
producers in the community), correspond to natural communities (i.e. not deliberately
altered by human impacts such as eutrophication) and integrate several seasons of
observations. The compilation shows that, in spite of substantial variability within
each producer type, seagrass leaves tend to have lower nitrogen (Figs. 4B and C) and
phosphorus (Figs. 4E and F) contents than macroalgae (non-parametric Q multiple
comparison test, P<0.05 for both tests). Hence, it may be expected that a replacement
of seagrass by macroalgae as dominant producers under increasing eutrophication
enhances the nitrogen and phosphorus contents of the producer community. In
addition, under high nitrogen availability bloom-forming macroalgal genera may
increase their nitrogen contents to levels that stand high in relation to most other
macroalgae (e.g. > 4%DW for Polysiphonia, Ulva and Cladophora, Peckol et al.
1994, Campbell 2001; >3%DW for Enteromorpha, Barr & Rees 2003), which further
supports the hypothesis of an overall increase in producer nitrogen content as a result
of macroalgal pile-ups and seagrass decline. Higher nutrient availability can also
increase the nutrient content of seagrass leaves (Fourqurean et al. 1997, van Katwijk
et al. 1999), so declining but richer seagrass leaves under increasing eutrophication
could also contribute to the higher overall nutrient contents of the producer
If the replacement of seagrasses by bloom-forming macroalgae does indeed lead to
higher nutrient contents in the producer assemblage, and based on the results of
Figures 2A and B, it could then be expected that herbivores exert a greater pressure
on the newly formed macroalgal canopy by removing a larger percentage of primary
production. No significant changes, however, would necessarily be expected in
absolute consumption since a concomitant decrease in total primary production could
offset the increase in the percentage consumed (Figs. 2C and D; Borum & SandJensen 1996, Cebrian 1999). At any rate, the association between larger percentages
of production removed and higher producer nutrient contents found across producer
types (Figs 2 A and B) is not strong and it is possible that other concomitant
deleterious effects of macroalgal build-ups on herbivory override the expected
positive effect of enhanced producer nutrient contents. Some of those possible
negative effects are discussed below.
My compilation also shows that microphytobenthos tends to have higher nitrogen
(Figs. 4A and B) and phosphorus (Figs. 4D and E) cell contents than do macroalgal
tissues (non-parametric Q multiple comparison test, P<0.05 for both tests). Values of
nitrogen and phosphorus contents in microphytobenthic cells are scarce, but the result
is robust in spite of having few values available for the comparison. Hence, and
based again on the association between larger percentages of production removed and
Grazing on benthic primary producers
higher nutrient contents found across different types of benthic producers (Figs. 2A
and B), it may seem that a hypothetical replacement of macroalgal canopies by
sediment microalgae due to intense shading by phytoplankton that would occur under
intense eutrophication could also be conducive to higher intensities of herbivory.
This logic, however, is fallacious, since the association between larger percentages
consumed and higher producer nutrient contents does not hold when only macroalgae
and microphytobenthos are compared (P>0.05 for both correlation coefficients). A
replacement of macroalgae by microphytobenthos under intense eutrophication
undoubtedly seems to lead to enhanced nutrient contents in the producer community,
but we do not have a strong empirical basis to predict how that enhancement can
affect herbivory.
3.1.2 Changes in Structure
A wealth of reports have demonstrated that the structurally complex web of vertical
shoots, leaves and attached epiphytes in seagrass meadows offers numerous refuge
possibilities to many organisms (e.g. Orth 1977, Heck & Orth 1980, Heck & Crowder
1991, Sogard, 1992, Edgard & Shaw 1995). Indeed, seagrass meadows are
considered as one of the most important habitats in coastal ecosystems (Orth et al.
1984, Hemminga & Duarte 2000, Williams & Heck 2001). Some organisms remain
in the meadows throughout their entire life span (i.e. permanent residents), whereas
some others use the meadow only during certain stages of their life cycles (i.e.
temporal residents). Notorious among those temporal residents are the juveniles of
many fish species with commercial and recreational interest. Upon recruitment from
offshore waters, the juveniles find shelter and abundant food in the meadows and,
after reaching a certain size, swim to deeper waters (Coles et al. 1993, Ros & Moser
1995, Rooker et al. 1998). Seagrass meadows thus play an important role as
"nurseries" for numerous fish species (Heck et al. 2003).
The dense macroalgal canopies that are formed under increasing eutrophication are
also structurally complex and can thus offer shelter to many organisms. Indeed,
several studies have shown that increasing algal abundance reduces predation rates on
mobile epifauna. For instance, Isaksson et al. (1994) and Pihl et al. (1995) showed
that increasing cover by Enteromorpha and Cladophora depressed the predation
rates of the crustaceans Crangon crangon and Carcinus maenus by cod (Gadus
morhua). Bluecrab (Callinectes sapidus) may also find refuge from predation in mats
formed by the macroalgae Ulva lactuca (Wilson et al. 1990). More recently, Norkko
(1998) examined how algal blooms may influence the trophic relationships between
infaunal preys and epibenthic predators and provided experimental evidence that
blooms of Ectocarpus and Pilayella offer substantial refuge to the infaunal amphipod
Corophium volutator and polychaete Nereis diversicolor from the epibenthic predator
Crangon crangon. Dense macroalgal blooms may also be valuable habitats for a
number of fish species, such as the four-spined stickleback (Apeltes quadracus,
Raposa & Oviatt 200) and tautog (Tautoga onitis, Dorf & Powell 1997).
Figure 4. The histograms of nitrogen (plots A-C) and phosphorus (plots D-F) contents in
microphytobenthos, macroalgae and seagrasses. Dashed lines represent medians.
Overall, thus, it is not straightforward how increasing algal abundance and declining
seagrass abundance following eutrophication may affect the degree of refuge offered
by the producer assemblage. Based on the arguments presented above, it may seem
that the structural complexity gained with the building of dense macroalgal blooms
can compensate for the loss due to declining seagrass, with the overall level of refuge
offered by the producer community thus being little altered. Unfortunately, empirical
tests of this hypothesis are lacking. Common methods to measure structural
complexity in seagrass and macroalgal canopies have not been developed. Fractal
geometry may prove as a good solution (Sugihara & May 2000, Chaplin 2001), but
much research is needed to decide how the technique can be commonly applied to
seagrass and macroalgal canopies. Elucidating how the replacement of seagrasses by
macroalgae under increasing eutrophication affects the degree of refuge provided by
the producer assemblage to organisms is an important question to understand how
eutrophication, one of the most pervasive world-wide human impacts, may affect the
dynamics of coastal ecosystems. Future research would do well in addressing that
question. Under further eutrophication, the potential loss of macroalgal canopies due
to intense shading by phytoplankton and subsequent replacement by bare bottom
Grazing on benthic primary producers
populated by microphytobenthos would certainly decrease drastically the extent of
refuge offered to organisms. It may be expected that the faunal population would
then shift from being dominated by epibenthic species to being dominated by infaunal
species that live and hide in the sediment. This hypothesis is in accordance with
recent reports of the response of faunal populations to the experimental removal of
macroalgal canopies (Franz & Friedman 2002).
3.1.3 Changes in Biochemical Conditions
Perhaps one of the best documented processes that often results from the
accumulation of large macroalgal blooms is the occurrence of anoxic conditions
within the algal canopy. In dense algal canopies, photosynthesis is usually restricted
to the top layers of the canopy due to the intense self-shading. This was first shown
by Peckol and Rivers (1996), who found that, in thick canopies of Cladophora
vagabunda and Gracilaria tikvahiae in Waquoit Bay (Massachusetts, USA), 90% and
70% of incident light was intercepted in the first 2cm of the canopy (Fig. 5A). Part of
the oxygen produced in the top centimetres of the canopy diffuses downward and it is
rapidly taken up by shaded, non-photosynthesizing algae. In addition, macroalgal
canopies, in virtue of their intertwined texture, retain many of the detrital particles
that settle from the water column (Hull 1987, Worcester 1995), which serve as
substrata for bacterial growth and further consume the oxygen that diffuses from the
top layers. As a consequence, oxygen concentrations are reduced to hypoxic/anoxic
levels within the top centimetres of the canopy even on sunny days and high
photosynthetic activity of the top layer (Fig. 5B). The problem is worsened at night,
when anoxia occurs throughout most of the canopy. Anoxic conditions can even be
found at dawn above the canopy up in the stratified water column after cloudy, hot
summer days (D'Avanzo & Kremer 1994).
Many lab experiments and field surveys have shown that macroalgal blooms may
induce hypoxic and/or anoxic conditions within the canopy and overlying water
column. The large macroalgal blooms observed in Venice Lagoon (Italy) are a
classical example; the dynamics of macroalgal biomass in the Lagoon displays large
variability among seasons and locations, but when a large biomass of Ulva
accumulates, water-column anoxia may occur on hot summer days under cloudy
conditions or high water-column turbidity (Sfriso et al. 1992). The large canopies of
Ulva lactuca that accumulate in much of the shoreline of Jamaica Bay (New York,
USA) in the summer also produce anoxic conditions within the canopy, and
sometimes in the water column, during the night and early morning hours (Franz &
Friedman 2002). In turn, hypoxic/anoxic conditions are extremely harmful for most
organisms that inhabit the algal canopy. Some organisms have a higher tolerance to
oxygen deprivation (Sagasti et al. 2001), but in general, most fauna readily migrates
to the top canopy layer when oxygen becomes limiting or, if exposed to oxygen
deprivation for a sustained period, they perish (Rafaelli et al. 1998, Tagliapietra et al.
1998, Osterling & Pihl 2001, Franz & Friedman 2002, Gray et al. 2002). Not all
macroalgal blooms are conducive to anoxic conditions. In some cases, the blooms
occur as sparse drifting algal masses that are relatively well flushed and exposed to
continuous air exchange with the atmosphere (Rafaelli et al. 1998). Under these
conditions, hypoxic/anoxic conditions are rarely encountered within the algal mass
(Thybo-Christensen et al. 1993, Norkko & Bonsdorff 1996, Norkko et al. 2000).
However, if the blooms occur as thick macroalgal canopies that remain stagnant on
the bottom and subject to hydrographical isolation through stratification,
anoxic/hypoxic conditions and subsequent detrimental effects for the resident fauna
can be expected.
Oxygen depletion is not the only harmful condition that organisms face within the
canopies of algal blooms because the highly reduced environment that occurs with
anoxia favours the formation of sulphides and ammonia (Gray et al. 2002) In fact,
Nedergaard et al. (2002) reported high hydrogen sulphide concentrations within
canopies of Ulva lactuca. There have been no direct measurements of ammonia
concentrations within canopies of bloom-forming algae, but McGlathery et al. (1997)
and Krause-Jensen et al. (1999) measured high ammonium concentrations within
canopies of Chaetomorpha linum, and Hauxwell et al. (2001) measured high
concentrations within mixed canopies of Cladophora vagabunda and Gracilaria
tikvahiae (Fig. 5B). In turn, elevated ammonium concentrations should also increase
the concentration of ammonia since the two products are in equilibrium in the marine
environment (Gray et al. 2002). Sulphides and ammonia are extremely toxic to many
marine organisms and can inflict substantial mortality even at low concentrations
(Gray et al. 2002). In addition, in a highly reduced environment the effects of anoxia
may interact with the effects of sulphide and ammonia and cause increased damage.
For instance, hypoxic tolerance in three species of marine gastropods (Littorina
ziczac, Neritina vorginea and Olivella vereuxii) decreased in the presence of
hydrogen sulphide (Hiroki 1978). Therefore, it seems that steady, thick macroalgal
canopies are, except in the upper layers, a hostile environment for many organisms.
Grazing on benthic primary producers
Figure 5. (A) Light attenuation through mats of Gracilaria tikvahiae (open squares) and
Cladophora vagabunda (filled squares) in Waquoit Bay (Massachusetts, USA). Light
attenuation is expressed as the percentage of irradiance measured at the mat surface that
remains at a given mat depth. From Peckol & Rivers (1996). (B) Profiles of oxygen and
ammonium concentrations throughout the water column and natural ("no algae" treatment)
and experimentally thickened ("algae" treatment) algal mats in the subestuaries Sage Lot Pond
and Hamblin Pond of the Waquoit Bay estuarine system . In Sage Lot Pond, natural and
experimental algal mats were 2 and 25cm tall, respectively, and in Hamblind Pond they were 9
and 18 cm tall. The algal mat consisted of a mixture of G. tikvahiae and C. vagabunda. In
each subestuary, dusk and dawn concentrations were taken at a number of depths (i.e. labelled
as "distance from the sediment" in the figure) through the water column and canopy of two
natural and two experimentally thickened mats. Average values and standard deviations are
plotted (see legend in figure). Measurements were done after a typical sunny summer day.
Dashed lines represent the height of thickened algal mats. Adapted from Hauxwell et al.
3.2 Overall Effects of Eutrophication-Induced Changes in Benthic Producer
Assemblages on Herbivory: Lessons from Case Studies
Based on the arguments elaborated above, we can expect diverse and counter effects
of eutrophication-driven macroalgal build-ups and seagrass decline on benthic
herbivory. One direct positive effect seems to be an overall increase in the nutrient
contents of the producer community. Anoxic conditions and elevated sulphide and
ammonia concentrations within the algal canopy, however, may impose severe
deleterious consequences on associated fauna and herbivory. It is not straightforward
how the replacement of seagrass meadows by macroalgal mats may alter the
structural complexity of the community and degree of refuge offered to organisms.
Both types of producer community are structurally very complex and important
habitats for many organisms. It thus seems that, when replacing seagrass meadows
by macroalgal mats, the effects on associated fauna due to changes on structure
would be modest in comparison with the effects of enhanced producer nutrient
contents and adverse biochemical conditions.
A logical question arises; what is the net effect of the replacement of seagrass
meadows by macroalgal blooms on benthic herbivory? In other words, how do two
of the most important consequences of that replacement, i.e. an increase in the
nutrient content of the producer community but adverse biochemical conditions,
interact and what are the overall implications on the extent of benthic herbivory?
Unfortunately, very little work has been done to answer that question, even though
elucidating how increased eutrophication, through changes in the dominant
assemblages of benthic producers, may affect herbivory is important to understand
associated changes in secondary production and nutrient recycling and storage in
coastal ecosystems. Many authors have examined how macroalgal blooms on bare
substratum affect faunal diversity and abundance (e.g. Rafaelli et al. 1998, Bolam et
al. 2000, Norkko et al. 2000, Osterling & Pihl 2001, Cardoso et al. 2002, Franz &
Friedman 2002), and a few authors have assessed the extent of herbivory on
macroalgal blooms (Hauxwell et al. 1998, Sfriso & Marcomini 1997), but no
comparison of how herbivory changes from seagrass-dominated to macroalgaldominated communities following increased eutrophication, and of the controls
responsible for those changes, has been done to date. Here, I attempt to shed some
light on that question by (1) examining a number of reports of the response of
macrofauna to macroalgal blooms on bare sediment and the role of anoxia on that
response and (2) analysing how herbivory on macroalgae has changed in two well
known ecosystems, Waquoit Bay (Massachusetts, USA) and Venice Lagoon (Italy),
where increased eutrophication has caused the persistence of dense macroalgal mats
and intense seagrass loss.
3.2.1 Macrofaunal response to macroalgal blooms on sediment flats.
A good number of authors have examined how macroalgal build-ups on intertidal and
subtidal sediments affect infaunal and epifaunal communities (see review by Rafaelli
et al. 1998). Those authors have focussed on such aspects as changes in abundance
(Osterling & Pihl 2001, Franz & Friedman 2002), diversity (Bolam et al. 2000,
Norkko et al. 2000), growth (Sogard 1992, Cardoso et al. 2002), recruitment
(Bonsdorf et al. 1996, Cardoso et al. 2002), and trophic relationships (Sogard & Able
Grazing on benthic primary producers
1991, Lopes et al. 2000). No report, however, has addressed how macroalgal
accumulations on bare substratum alter the intensity of benthic herbivory. Here, I
discuss existing information on the development of algal blooms on bare substratum
and effects on local fauna that can help predict what the consequences on herbivory
could be.
In what has become a classical review, Rafaelli et al. (1998) documented that
macroalgal blooms that develop on sediment flats can have contrasting effects on the
local fauna depending on the type of dominant fauna at the bloom onset, the size and
nature (drifting or steady) of the bloom, season and, to a lesser extent, the type of
bloom-forming algae. However, despite the high variability in the response of
macrofauna to macroalgal blooms, their analyses pointed to some apparent trends.
Macroalgal blooms, particularly those that remain anchored to the bottom, often
reduced the abundance of infauna (i.e. organisms living and/or hiding in sediment
burrows) and surface feeders, except for organisms that can thrive under highly
reduced conditions such as the polychaete Capitella capitata. Macroalgal blooms, on
the contrary, frequently increased the abundance of epifaunal species (i.e. organisms
living above the sediment), which readily colonized and found abundant shelter and
food in the new algal substratum. Overall, thus, one of the messages that stemmed
from the review by Rafaelli et al. (1998) is that macroalgal blooms tend to be
detrimental for most infaunal species but beneficial for many epifaunal species.
Nevertheless, research done since the review by Rafaelli et al. (1998) was published
has provided evidence that macroalgal blooms may also be detrimental for epifaunal
species through the generation of persistent anoxia. For instance, Franz & Friedman
(2002) did a series of algal removal/addition experiments to demonstrate that blooms
of Ulva lactuca drastically reduced the abundance of epibenthic copepods and
attributed that reduction to the anoxic conditions registered within the algal mats at
night. Similarly, Tagliapietra et al. (1998) showed opposite cycles between benthic
macrofaunal abundance, including many epibenthic species, and the abundance of
Ulva in the Palude della Rosa lagoon (Italy), with precipitous decreases in faunal
abundance following the formation of algal blooms. They also attributed those
changes to the severe anoxic conditions brought about by the blooms. In contrast,
Norkko et al. (2000) found high levels of epifaunal abundance in drifting algal mats
of Ectocarpus siliculosus and Pilayella littoralis, which could even exceed the high
levels typically found in seagrass meadows. Drifting algal mats are often highly
mobile and are thus relatively well flushed in relation to large, steady algal canopies.
Indeed, Norkko et al. (2000) reported intermediate oxygen levels within the drifting
mats and only found hypoxic conditions at the algal-sediment interface. Thus, the
lack of prolonged anoxia within the drifting mats, along with the enhanced levels of
refuge, food and surface for recruitment provided by the mats, probably explain the
high epifaunal abundance reported by the authors. Additional evidence of the
importance of anoxia in determining the overall effect of macroalgal blooms on
epifaunal populations is provided by Osterling & Pihl (2001). They subjected
subtidal faunal communities to moderate (i.e. mesh-bags filled with algae simulating
floating patches with loose attachment to the sediment) and intense (i.e. cages filled
up with algae simulating steadier, thicker canopies) algal disturbance and found
higher epifaunal abundances at moderate disturbance. They suggested that
differences in hypoxic/anoxic conditions were probably an important factor
explaining the results.
Hence, when one compares reports of positive and negative effects of macroalgal
blooms on epifaunal abundance, the occurrence of anoxia and probably other harmful
chemical reagents such as ammonia and sulphides within the algal canopy seems to
be of paramount importance in dictating the net consequences of the bloom; if anoxia
persists, an overall negative effect on the abundance of the epifaunal community
seems to prevail in spite of all the other possible concomitant positive effects of the
bloom (i.e. enhanced refuge, food and structure for recruitment). A number of
epifaunal species that thrive in algal canopies are herbivorous (Hicks & Coull 1983,
Salemaa 1987, Geertz-Hansen et al. 1993). Many other species are primarily
detritivorous but can also consume living tissues of the algae (Rafaelli et al. 1998,
Norkko et al. 2000). Thus, it may be hypothesised that the extent of herbivory on
macroalgal blooms is also primarily determined by the intensity and persistence of
oxygen reduction; if persistent anoxic conditions occur, herbivory, via depressed
herbivore abundance, would also be drastically reduced.
In contrast, in
comparatively well-oxygenated canopies, such as those of drifting mats, algal
consumption by herbivores could be substantial. The two case studies discussed
below represent a template to further explore this hypothesis.
3.2.2. Changes in herbivory on macroalgal mats in Waquoit Bay (Massachusetts,
USA) and the Venice Lagoon (Italy).
The estuarine system of Waquoit Bay comprises a number of subestuaries that have
undergone contrasting eutrophication over the last decades (Valiela et al. 1997b,
Valiela et al. 2000a). Some subestuaries are protected land and are thus surrounded
by Spartina marshes and forestlands. Those subestuaries receive small nutrient
loading rates from their watersheds. In contrast, some other subestuaries have been
greatly deforested mainly through intense urbanization and the construction of golf
courses. As a consequence, those subestuaries receive large nutrient loads from their
watersheds. Finally, some other subestuaries have only undergone moderate
urbanization and, consequently, they receive intermediate loads. Interestingly, the
physical and hydrographical differences among the subestuaries of the Waquoit Bay
system are small in comparison to the large variability in nitrogen loading rates,
which range from 5 to 400 Kg nitrogen ha-1 yr-1. Therefore, the system represents a
convenient scenario to test how increasing eutrophication may affect the ecology of
coastal waters.
One of the consequences of intense nitrogen loading has been the replacement of
native eelgrass (Zostera marina) meadows by massive accumulations of the algae
Cladophora vagabunda and Gracilaria tikvahiae (Valiela et al. 1992, Valiela et al.
2000b, Hauxwell et al. 2003). In the subestuaries subject to the highest loading,
eelgrass has disappeared completely and the algae form dense, thick canopies that
may reach a maximum height of ca. 80cm (Hauxwell et al. 2001).
On the contrary, in the pristine subestuaries, eelgrass meadows are still luxurious and
the algal canopy is sparse and only 2-3 cm tall on average. Hauxwell et al. (1998)
investigated how larger algal canopies produced with increasing eutrophication
affected the intensity of herbivory on the algae. To do so, they measured herbivory in
three subestuaries subject to contrasting nitrogen loading rates and, thus, with
Grazing on benthic primary producers
different algal canopies. Sage Lot Pond was the pristine environment, with a loading
rate of 5 Kg nitrogen ha-1 yr-1 and a maximum canopy height of 10 cm; Quashanet
River had intermediate values of loading rate (300 Kg nitrogen ha-1 yr-1) and
maximum canopy height (20 cm), and Childs River had the highest values (410 Kg
nitrogen ha-1 yr-1 and 75 cm, respectively).
The results obtained by Hauxwell et al. (1998) represent an exemplary case of how
the positive and negative effects entailed by macroalgal build-ups may affect the
intensity of herbivory on the algae. The seven-fold increase in maximum canopy
height observed from Sage Lot Pond to Childs River undoubtedly entailed a
substantial increase in refuge for many epifaunal species. Furthermore, as a
consequence of higher nutrient availability, the two algal species increased their
internal nutrient contents with higher nitrogen loading (Peckol et al. 1994), which
should also enhance their nutritional quality for herbivores. Accordingly, Hauxwell
et al. (1998) found that the individual consumption rates for most species of
herbivorous amphipods examined increased from Sage Lot Pond to Quashanet River
to Childs River (Fig. 4 in Hauxwell et al. (1998)). Nevertheless, they also found that
the total abundance of herbivorous amphipods and isopods within the algal canopy
decreased abruptly as nitrogen loading and height of the canopy increased, with the
summer abundance at the pristine Sage Lot Pond being four-fold higher than that at
the eutrophic Childs River. Those differences overrode the opposite trend found for
individual consumption rates and, as a consequence, the total absolute consumption
of algal biomass (in g DW per year per bottom m2) and the percentage of algal
production consumed decreased from Sage Lot Pond to Quashanet River to Childs
River. Later work by the same authors (Hauxwell et al. 2001, 2003) has shown that
the anoxic conditions created at night within the thick algal canopies at Quashanet
River and Child River, in contrast with the absence of anoxia in the sparse, thin
canopies at Sage Lot Pond, are probably responsible to a great extent for the
differences in herbivore abundance and intensity of herbivory found across the three
subestuaries (Fig. 6).
The Venice Lagoon (Italy) provides another example of the importance of anoxic
conditions in regulating the extent of herbivory on eutrophication-driven macroalgal
mats. The Venice Lagoon experienced substantial anthropogenic eutrophication
throughout the 70's and 80's via untreated sewage effluents and inputs from
agricultural fields in the watershed (Orio & Donazzolo 1987, Sfriso et al. 1992). This
caused the replacement of beds of seagrasses (Zostera marina, Z. noltii and
Cymodocea nodosa) and bulky, slow-growing red and brown macroalgae (Cystoseira
fimbriata, C. barbata and Dictyopteris membranacea) by mats of green filamentous
algal species (Ulva rigida, Enteromorpha sp. and Cladophora sp; Pignatti 1962,
Sfriso 1987, Sfriso et al. 1989) in central parts of the Lagoon where the water is
slowly renewed by offshore influx (Battiston et al. 1983). In turn, those changes in
benthic producer assemblages have entailed profound consequences on the local
fauna, which has also undergone substantial changes in diversity and abundance
(Giordani Soika & Perin 1970, Tagliapietra et al. 1998).
Sfriso et al. (1992) showed that one of the most important effects induced by
macroalgal mats on the faunal populations in the central part of the Lagoon was the
frequent occurrence of hypoxic/anoxic conditions. They sampled several stations in
the central part of the Lagoon and found that, in the stations with the highest nutrient
inputs and least exposed to flushing by offshore water, large mats of Ulva and
Enteromorpha accumulated in the summer months and anoxic conditions occurred
under reduced light intensity (i.e. cloudy days) and/or water-column stratification.
They also found large mortality events of fish and benthic animals associated with
those anoxia events. Sfriso et al. (1992) did not measure the extent of herbivory on
the macroalgal mats during well-oxygenated and anoxic conditions, but it is obvious
that herbivory would have been drastically reduced, if not completely eliminated,
under anoxic conditions. Interestingly, though, a few years later Sfriso and
Marcomini (1997) measured herbivory on the Ulva stands of one of the previously
most polluted stations and where they had found the highest occurrence of anoxic
events (Lido station). They also compared those herbivory values with those
measured in a southern station dominated by eelgrass (Z. marina) that had undergone
little eutrophication. Sfriso and Marcomini (1997) found intense herbivory on the
Ulva mats (65% of the algae production consumed), but negligible values on the
seagrass. What is very relevant to this discussion is that, in contrast to the frequent
occurrence of anoxia recorded a few years earlier, they did not find any anoxic
conditions within the algal mat at the Lido station. They attributed the lack of anoxic
conditions to reduced levels of Ulva abundance in relation to the levels measured a
few years earlier. In their conclusion, they mentioned that "in the absence of anoxic
conditions, the benthic herbivores associated with Ulva drastically reduced the
biomass produced by this species whereas, in areas dominated by rhizophytes,
herbivores feed primarily on the macroalgae growing within the Zostera shoots"
The analyses presented here confirm past observations that herbivory on benthic
producers can be very variable. However, the results reveal a strong tendency for
seagrass communities to have a smaller percentage of primary production removed by
herbivores in comparison with communities of microphytobenthos and macroalgae.
These differences suggest that herbivores should generally play a lesser role in the
dynamics of producer biomass and in the recycling of producer-bound nutrients in the
former communities. The results also suggest that the lower nutrient contents found
for seagrass leaves in relation to the contents in macroalgal tissues and
microphytobenthic cells may be partially responsible for the smaller percentages
consumed in seagrasses. Yet, the association between higher producer nutrient
contents and larger percentages of production consumed found across types of
benthic producer is not strong and other factors, such as herbivore abundance, size or
feeding specificity, may be more important at explaining the variability in percentage
consumed across benthic producers.
Grazing on benthic primary producers
Figure 6. Hypothesized causal relationships between increased eutrophication and the
changes observed in algal herbivory in subestuaries of the Waquoit Bay system. (A) Higher
nutrient loading rates are conducive to larger macroalgal mats (macroalgal biomass, mostly
accounted for by C. vagabunda, is expressed in g dry weight per square meter; Sage Lot Pond
is the "low N estuary", Quashanet River is the "mid N estuary" and Childs River is the 'high N
estuary". Adapated from Hauxwell et al. 1998). (B) Larger algal canopies induce anoxic
conditions within the canopy (see Fig. 5B for details; adapted from Hauxwell et al. (2001)).
(C) Increased anoxia within the algal canopies leads to depressed grazer abundance (legends
for estuaries as in Figure 6A; adapted from Hauxwell et al. (1998)). (D) depressed grazer
abundance results into lower leves of herbivory on the algae, both expressed as absolute
consumption (in g algal dry weight per bottom square meter per day, Y-axis on the plot) and as
percentage of algal production consumed (the ratio between absolute consumption and algal
production, which is also exprssed in g algal dry weight per bottom square meter per day and
corresponds to the X-axis in the plot) (legends for estuaries as in Figure 6A; the 1:1 line has
been depicted for comparison; adapted from Hauxwell et al. (1998))
More research is needed to elucidate why seagrasses tend to lose a lower percentage
of production to herbivores when compared with microphytobenthos and macroalgae.
However, when herbivory is regarded as an absolute flux, seagrasses tend to show
only marginally lower values than do the two types of algal producers because the
higher values of primary production typically found for seagrasses partially
compensate for the smaller percentages of production removed. Indeed, primary
production stands out as a strong indicator of the variability in absolute herbivory
when benthic producers of the same type or different types are compared; when
benthic micro- and macroalgae and seagrasses are examined in concert or separately,
larger absolute consumption is strongly associated with higher primary production.
The reason is entirely mathematical; both with and across producer types, the range in
primary production largely exceeds that in the percentage consumed and, as a
consequence, absolute consumption remains closely associated with primary
These results provide a basis to formulate hypotheses as to how the replacement of
seagrass communities by mats of filamentous algae that usually follows
anthropogenic eutrophication may affect the intensity of benthic herbivory. This is an
important question to understand how this pervasive, world-wide environmental
impact alters secondary production and carbon and nutrient cycling in marine
ecosystems but, unfortunately, the question is yet to be investigated. My results
suggest that the overall increase in producer nutrient content entailed by the
replacement of seagrasses by filamentous algae as dominant producers may be
conducive to higher percentages of primary production consumed by herbivores. In
turn, that would promote the role of herbivores as controls of carbon and nutrient
recycling and storage as producer biomass. Changes in absolute consumption and
implications on secondary production would instead be determined by the balance
between the changes in primary production and percentage consumed that would
occur with the shift from seagrasses to macroalgae. Nevertheless, abundant evidence
indicates that the adverse biochemical conditions that are usually found in
eutrophication-driven thick macroalgal mats may be of paramount importance in
determining the response of herbivory. Numerous reports of the effects of macroalgal
blooms in sediment flats on associated fauna and the history of research in the highlyeutrophic ecosystems of Waquoit Bay (Massachusetts, USA) and Venice Lagoon
(Italy) suggest that, if the macroalgal mats that outcompete seagrass beds under
enhanced eutrophication induce persistent anoxia and elevated ammonia and sulphide
concentrations within the mat, the extent of herbivory, via decreased faunal
abundance, would be drastically limited to values even lower than those on the
former seagrass community. Much less evidence exists to hypothesise how a possible
replacement of macroalgal mats by microphytobenthos-dominated bare sediment due
to severe phytoplankton shading under intense eutrophication may affect the levels of
benthic herbivory. Profound changes in producer nutrient content, spatial structure,
biochemical conditions and other factors affecting herbivory will follow the
replacement of macroalgal mats by bare bottom inhabited by microphytobenthos.
Those changes may have contrasting effects on local herbivore populations, and the
possible overall impact on the intensity of benthic herbivory is not straightforward.
As the replacement of macroalgal mats by sediment flats dominated by
microphytobenthos seems plausible under substantial phytoplankton shading,
Grazing on benthic primary producers
research should also be undertaken to investigate how that replacement may affect
herbivory and other routes of carbon and nutrient cycling.
Admiraal, W., Bouwman, L. A., Hoekstra, L. & Romeyn, K. (1983). Qualitative and Quantitative
Interactions between Microphytobenthos and Herbivorous Meiofauna on a Brackish Intertidal Mudflat.
Internationale Revue der Gesamten Hydrobiologie, 68, 175-191.
Alongi, D. M. (1998). Coastal ecosystems processes. First Edition. CRC Press, New York, USA.
Atkison, M. J. & Smith, S. V. (1983). C:N:P ratios of benthic marine plants. Limnology and
Oceanography, 28, 568-574.
Aubert, M. (1990). Mediators of microbiological origin and eutrophication phenomena. Marine Pollution
Bulletin, 21, 24-29.
Baird, D. & Ulanowicz, R. E. (1993) Comparative study on the trophic structure, cycling and ecosystem
properties of four tidal estuaries. Marine Ecology Progress Series 99, 221-237.
Barr, N. G. & Rees, T. A. V. (2003). Nitrogen status and metabolism in the green seaweed Enteromorpha
intestinalis: an examination of three natural populations. Marine Ecology Progress Series, 249, 133144.
Battison, L., Giommoni, L., Pilan, L. & Vicenzi, S. (1983). Salinity exchange induced by tide in the Venice
Lagoon (in Italian). In Conference Proceedings on Five Centuries of Water Management in the Venice
Territory (pp.1-7). Venice.
Bennett, A., Bianchi, T. S., Means, J. C. & Carman, K. R. (1999). The effects of polycyclic aromatic
hydrocarbon contamination and grazing on the abundance and composition of microphytobenthos in
salt marsh sediments. Journal of Experimental Marine Biology and Ecology, 242, 1-20.
Bergin, F. (1987). Contenus digestifs de Paracentrotus lividus et d'Arbacia lixula dans la region d'El Dabaa
(Egypte). In C. F. Boudouresque (Ed.), Colloque International sur Paracentrotus lividus et les oursins
comestibles (pp. 107-116). GIS Posidonie publications, Marseille, France.
Blanchard, G. F., Guarini, J-M., Orvain, F. & Sauriau, P-G. (2001). Dynamic behaviour of benthic
microalgal biomass in intertidal mudflats. Journal of Experimental Marine Biology and Ecology, 264,
Bjorndal, K. A. (1980). Nutrition and grazing behaviour of the green turtle Chelonia mydas. Marine
Biology, 56, 147-154.
Bolam, S. G., Fernandes, T. F., Read, P. & Raffaelli, D. (2000). Effects of macroalgal mats on intertidal
sandflats: en experimental study. Journal of Experiimental Marine Biology and Ecology, 249, 123137.
Bombelli, V. & Lenzi, M. (1996). Italy-the Orbetello lagoon and the Tuscan coast. In W. Schramm & P. H.
Nienhuis (Eds.), Marine benthic vegetation: recent changes and the effects of eutrophication (pp. 331338). Berling: Springer
Bonsdorff, E., Blomqvist, E. M., Mattila, J. & Norkko, A. (1996). Long-term changes and coastal
eutrophication. Examples from the Aland Islands and the Archipelago Sea, northern Baltic Sea.
Oceanologica Acta, 20, 319-329.
Borum, J. & Sand-Jensen, K. (1996). Is total primary production in shallow coastal marine waters
stimulated by nitrogen loading? Oikos, 76, 406-410.
Breen, P. A. & Mann, K. H. (1976). Destructive grazing of kelp by sea uchins in eastern Canada. Journal
of the Fisheries Research Board of Canada, 33, 1278-1283.
Campbell, S. (2001). Ammonium requirements of fast-growing ephemeral macroalgae in a nutrientenriched marine embayment (Port Phillip Bay, Australia). Marine Ecology Progress Series, 209, 99107.
Cardoso, P. G., Lillebo, A. I., Pardal, M. A., Ferreira, S. M. & Marques, J. C. (2002). The effect of
different primary producers on Hydrobia ulvae population dynamics: a case study in a temperate
intertidal estuary. Journal of Experimental Marine Biology and Ecology, 277, 173-195.
Carman, K. R. & Thistle, D. (1985). Microbial food partitioning by three species of benthic copepods.
Marine Biology, 88, 143-148.
Carman, K. R. & Fry, B. (2002). Small-sample methods for ∂13C and ∂15N analysis of the diets of marsh
meiofaunal species using natural-abundance and tracer-addition isotope techniques. Marine Ecology
Progress Series, 240, 85-92.
Carpenter, R. C. (1986). Partitioning herbivory and its effects on coral reef algal communities. Ecological
Monographs, 56, 345-363.
Cebrian, J. (1999). Patterns in the fate of production in plant communities. The American Naturalist, 154,
Cebrian, J. (2002). Variability and control of carbon consumption, export and accumulation in marine
communities. Limnology and Oceanography, 47, 11-22.
Cebrian, J. & Duarte, C. M. (1994). The dependence of herbivory on growth rate in natural plant
communities. Functional Ecology, 8, 518-525
Cebrian, J. & Duarte, C. M. (1998). Patterns in leaf herbivory on seagrasses. Aquatic Boatny, 60, 67-82.
Cebrian, J., Duarte, C. M., Marba, N., Enriquez, S. , Gallegos, M., Olesen, B. (1996). Herbivory on
Posidonia oceanica (L.) Delile: magnitude and variability in the Spanish Mediterranean. Marine
Ecology Progress Series, 130, 147-155.
Cebrian, 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,
Chaplin, G. I. (2001). Effects of habitat complexity, created by native and exotic macrophytes, on
secondary production and the transfer of energy to higher order consumers in the lower MobileTensaw delta, Alabama. (MS. Thesis, University of South Alabama, Alabama, USA, 2001).
Ciszewski, P., Kruk-Dowgiallo, L. & Zmudzinski, L. (1992). Deterioration of the Puck Bay and
biotechnical approaches to its reclamation. Proceedings of the 12th Baltic Marine Biology Symposium
1991 (pp. 44-46). Helsingor.
Cloern, J. E. (2001). Our evolving conceptual model of the coastal eutrophication problem. Marine
Ecology Progress Series, 210, 223-253.
Coles, R. G., Lee Long, W. J. & Watson, R A. (1993). Distribution of seagrasses, and their fish and
Penaeid prawn communities, in Cairns Harbour, a tropical estuary, Northern Queensland, Australia.
Australian Journal of Marine and Freshwater Research, 44, 193-210.
D'Avanzo, X. & Kremer, J. N. (1994). Diel Oxygen Dynamics and Anoxic Events in an Eutrophic Estuary
of Waquoit Bay, Massachusetts. Estuaries, 17, 131-139.
Dawes, C. J. & Lawerence, J. M. (1980). Seasonal changes in the proximate constituents of the seagrasses
Thalassia testudinum, Halodule wrightii and Syringodium filiforme. Aquatic Botany 8, 371-380.
Dayton, P. K. (1985). Ecology of kelp communities. Annual Reviews of Ecology and Systematics, 16, 215245.
De Vries, I., Philippart, C. J. M., De Groot, E. G. & van der Toll, M. W. M. (1996). Coastal eutrophication
and marine benthic vegetation: a model analysis. In W. Schramm & P. H. Nienhuis (Eds.), Marine
benthic vegetation: recent changes and the effects of eutrophication (pp. 79-114). Berlin: Springer.
de Iongh, H. H., Wenno, B. J. & Meelis, E. (1995). Seagrass distribution and seasonal biomass changes in
relation to dugong grazing in the Moluccas, East Indonesia. Aquatic Botany, 50, 1-19.
Dion, P. & Le Bosec, S. (1996). The French Atlantic Coasts. In W. Schramm & P. H. Nienhuis (Eds.),
Marine benthic vegetation: recent changes and the effects of eutrophication (pp. 251-264). Berlin:
Dorf, B. A. & Powell, J. C. (1997). Distribution, abundance, and habitat characteristics of juvenile tautog
(Tautoga onitis, family Labridae) in Narragansett Bay, Rhode Island, 1988-1992. Estuaries, 20, 589600.
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,
Duarte, C. M. & Cebrian, J. (1996) The fate of marine autotrophic production. Limnology and
Oceanography, 41, 1759-1766.
Edgard, G. J. & Shaw, C. (1995). The production and trophic ecology of shallow water fish assemblages in
southern Australia III. General relationships between sediments, seagrasses, invertebrates and fishes.
Journal of Experimental Marine Biology and Ecology, 194, 107-131.
Elser, J. J. & Urabe, J. (1999). The stoichiometry of consumer-driven nutrient recycling: theory,
observations, and consequences. Ecology, 80, 735-751.
Elser, J. J., Dobberfuhl, D. R., MacKay, N. A. & Schampel, J. H. (1996). Organism Size, Life History, and
N:P Stoichiometry. BioScience 46, 674-684.
Elser, J. J., Fagan, W. F., Denno, R. F., Dobberfuhl, D. R., Folarin, A., Huberty, A., Interlandi, S., Kilham,
S. K., McCauley E., Schulz, K. L., Siemann, E. H. & Sterner, R. W. (2000a). Nutritional constraints in
terrestrial and freshwater food webs. Nature 408, 578-580.
Elser, J. J., Sterner, R. W., Gorokhova, W., Fagan, W. F., Markow, T. A., Cotner, J. B., Harrison, J. F.,
Hobbie, S. H., Odell, G. M. & Weider, L. J. (2000b). Biological stoichiometry from genes to
ecosystems. Ecology Letters 3, 540-550.
Grazing on benthic primary producers
Estes, J. A. & J. F. Palmisano (1974). Sea otters: their role in structuring nearshore communities. Science,
185, 1058-1060.
Ferreira, C. E. L., Goncalves, J. E. A., Coutinho, R. & Peret, A. C. (1998). Herbivory by the Dusky
Damselfish Stegastes fuscus (Cuvier, 1830) in a tropical rocky shore: effects on the benthic
community. Journal of Experimental Marine Biology and Ecology, 229, 241-264.
Fletcher, R. L. (1996a). The occurrence of "Green tides"- a review. In W. Schramm & P. H. Nienhuis
(Eds.), Marine benthic vegetation: recent changes and the effects of eutrophication (pp. 7-43). Berlin:
Fletcher, R. L. (1996b). The British Isles. In W. Schramm & P. H. Nienhuis (Eds.), Marine benthic
vegetation: recent changes and the effects of eutrophication (pp. 223-250). Berlin: Springer.
Foster, M. S. & Schiel, D. R. (1988). Kelp communities and sea otters: keystone species or just another
brick in the wall? In G. R. VanBlaricom & J. A. Estes (Eds.), The community ecology of sea otters
(pp. 92-115). Springer-Verlag, Berlin.
Fourqurean, J. W., Moore, T. O., Fry & Hollibaugh, J. T. (1997). Spatial and temporal variation in C:N:P
ratios, ∂15N, and ∂13C of eelgrass Zostera marina as indicators of ecosystem processes, Tomales Bay,
California, USA. Marine Ecology Progress Series, 157, 147-157.
Franz, D. R. & Friedman, I. (2002). Effects of macroalgal mat (Ulva lactuca) on estuarine sand flat
copepods: an experimental study. Journal of Experimental Marine Biology and Ecology 271, 209-226.
Funnen Island Council (1991). Eutrophication of coastal waters. Coastal water quality management in the
county of Funen, Denmark 1976-1990. Denmark: Funnen Island Council.
Geertz-Hansen, O., Sand-Jensen, K., Hansen, D. F. & Christiansen, A. (1993). Growth and grazing control
of abundance of the marine macroalgae, Ulva lactuca, L. in a eutrophic Danish estuary. Aquatic
Botany 46: 101-109.
Giordiani Soika, A. & Perin, G. (1970). Changes of chemical characteristics and population variation in the
lagoon sediments in the last twenty years. I. Changes of animal populations (in Italian). In XI
Conference Proceedings A.N.L.S.B. (pp. 135-139). Venezia.
Gray, J. S., Wu, R. S. & Or, Y. Y. (2002). Effects of hypoxia and organic enrichment on the coastal
marine environment. Marine Ecology Progress Series, 238, 249-279.
Greenway, M. (1976). The grazing of Thalassia testudinum in Kingston Harbour, Jamaica. Aquatic
Botany, 2, 117-126.
Hackney, J. M., Carpenter, R. C. & Adey, W. H. (1989). Characteristic adaptations to grazing among algal
turfs on a Caribbean reef. Phycologia, 28, 109-119.
Hagen, N. T. (1995). Recurrent destructive grazing on successionally immature kelp forests by green sea
urchins in Vestfjorden, Northern Norway. Marine Ecology Progress Series, 123, 95-106.
Hatcher, B. G. & Larkum, A. W. D. (1983). An experimental analysis of factors controlling the standing
crop of the epilithic algal community on a coral reef. Journal of Experimental Marine Biology and
Ecology, 69, 61-84.
Hauxwell, J., McClelland, J., Behr, P. J. & Valiela, I. (1998). Relative importance of grazing and nutrient
controls of macroalgal biomass in three temperate shallow estuaries. Estuaries, 21, 347-360.
Hauxwell, J., Cebrian, J., Furlong, C. & Valiela, I. (2001). Macroalgal canopies contribute to eelgrass
(Zostera marina) decline in temperate estuarine ecosystems. Ecology, 82, 1007-1022.
Hauxwell, J., Cebrian, J. and Valiela, I. (2003). Eelgrass Zostera marina loss in temperate estuaries:
relationship to land-derived nitrogen loads and effect of light limitation imposed by algae. Marine
Ecology Progress Series, 247: 59-73
Hay, M. E. (1983). Spatial and temporal patterns in herbivory on a Caribbean fringing reef: the effects on
plant distribution. Oecologia (Berlin), 58, 229-308.
Hay, M. E. (1984). Patterns of fish and urching grazing on Caribbean coral reefs:are previous results
typical? Ecology 65: 446-454.
Hay, M. E. (1997). Calcified seaweeds on coral reefs: complex defenses, trophic relationships and value as
habitats. Proceedings 8th International Coral Reef Symposium 1, (pp. 713-718).
Hay, M. E. & Fenical, W. (1996). Chemical ecology and marine biodiversity: insights and products from
the sea. Oceanography, 9, 10-20.
Heck, K. L. Jr. & Orth, R. J. (1980). Seagrass habitats: the roles of habitat complexity, competition and
predation in structuring associated fish and motile macroinvertebrate assemblages. In V. S. Kennedy
(Ed.), Estuarine Perspectives (pp. 449-464). Academic Press, New York.
Heck, K. L. Jr. & Crowder, L. B. (1991). Habitat structure and predator-prey interactions in vegetated
aquatic ecosystems. In S. S. Bell, E. D. McCoy & E. R. Mushinsky (Eds.), Habitat Structure of Objects
in Space (pp. 281-299). Chapman and Hall, London.
Heck, K. L. Jr. & Valentine, J. F. (1995). Sea urchin herbivory: evidence for long-lasting effects in
subtropical seagrass meadows. Journal of Experimental Marine Biology and Ecology, 189, 205-217.
Heck, K. L. Jr., Hays, G. & Orth R. J. (2003). Critical evaluation of the nursery role hypothesis for
seagrass meadows. Marine Ecology Progress Series 253: 123-136.
Heck, K. L. Jr, Pennock, J. R., Valentine, J. F., Coen, L. D. & Sklenar, S. A. (2000). Effects of nutrient
enrichment and small predator density on seagrass ecosystems: an experimental assessment.
Limnology and Oceanography, 45, 1041-1057.
Hemminga, M. A. & Duarte, C. M. (2000). Seagrass Ecology. Cambridge University Press, Cambridge.
Hicks, G. R. F & Coull, B. C. (1983). The ecology of marine meiobenthic harpacticoid copepods.
Oceanography and Marine Biology: an Annual Review 21, 67-175.
Himmelman, J. H. (1984) Urchin feeding and macroalgal distribution in Newfoundland, Eastern Canada.
Naturaliste Canadien (Revue d'Ecologie et de Systematique), 111, 337-348.
Hiroki, K. (1978). Resistance of marine gastropods to oxygen deficiency and hydrogen sulphide. Boletim
de Fisiologia Animal (Sao Paulo), 2, 33-42.
Hixon, M. A. & Brostoff, W. N. (1996). Succession and herbivory: effects of differential fish grazing on
Hawaiian coral-reef algae. Ecological Monographs, 66, 67-90.
Hull, S. C. (1987). Macroalgal mats and species abundances: a field experiment. Estuarine, Coastal and
Shelf Science, 25, 519-532.
Isaksson, I., Pihl, L. & van Montfrans, J. (1994). Eutrophication-related changes in macrovegetation and
foraging of young cod (Gadus morhua L.): a mesocosm experiment. Journal of Experimental Marine
Biology and Ecology, 177, 203-217.
Jackson, J. (1997). Reefs since Columbus. Proc. 8th International Coral Reef Symposium (pp.97-106).
Jeffrey, D. W. (1993). Sources of nitrogen for nuisance macroalgal growths in Dublin Bay, Republic of
Ireland. The Phycologist, 34, 30 only.
Jeffrey, D. W., Madden, B., Rafferty, B., Dwyer, R., Wilson, J. & Arnott, N. (1992). Dublin Bay water
quality management plan. Technical Report 7. Algal growths and foreshore quality. Dublin:
Environmental Research Unit.
Karez, R., Engelbert, S. & Sommer, U. (2000). "Co-consumption" and "protective coating": two new
proposed effects of epiphytes on their macroalgal hosts in mesograzer-epiphyte-host interactions.
Marine Ecology Progress Series, 205, 85-93.
Keller, B. D. (1983). Coexistence of sea urchins in seagrass meadows: an experimental analyses of
competition and predation. Ecology, 64, 1581-1598.
Kirsch, K. D., Valentine, J. F. & Heck, K. L. Jr. (2002). Parrotfish grazing on turtlegrass Thalassia
testudinum: evidence for the importance of seagrass consumption in food web dynamics of the Florida
Keys National Marine Sanctuary. Marine Ecology Progress Series 227: 71-85.
Klinger, T. S., Lawrence, J. M. & Lawerence, A. L. (1994). Digestive characteristics of the sea-urchin
Lytechinus variegatus (Lamarck) (Echinodermata:Echinoidea) fed prepared foods. Journal of the
World Aquaculture Society, 25, 489-496.
Krause-Jensen, D., Christensen, P. B. & Rysgaard, S. (1999). Oxygen and nutrient dynamics within mats of
the filamentous macroalga Chaetomorpha linum. Estuaries, 22, 31-38.
Lanyon, J. M., Limpus, C. J. & Marsh, H. (1989). Dugongs and turtles: grazers in the seagrass system. In
A. W. D. Larkum, A. J. McComb & S. A. Shepard (Eds.), Biology of seagrasses: a treatise of the
biology of seagrasses with special reference to the Australian region (pp. 610-634). Elsevier,
Lapointe, B. E. & O'Connell, J. (1989). Nutrient-enhanced growth of Cladophora prolifera in Harrington
Sound, Bermuda: eutrophication of a confined, phosphorus limited marine system. Estuarine, Coastal
and Shelf Science, 28, 347-360.
Lawrence, J. M. (1975). On the relationships between marine plants and sea urchins. Oceanography and
Marine Biology: an Annual Review 13, 213-286.
Littler, M. M., Taylor, P. R. & Littler, D. S. (1983). Algal resistance to herbivory on a Caribbean barrier
reef. Coral Reefs, 2, 111-118.
Lopes, R. J., Pardal, M. A. & Marques, J. C. (2000). Impact of macroalgal blooms and wader predation on
intertidal macroinvertebrates: experimental evidence from Mondego estuary (Portugal). Journal of
Experimental Marine Biology and Ecology, 249, 165-179.
Lotze, H. K. & Worm, B. (2002). Complex interactions of climatic and ecological controls on macroalgal
recruitment. Limnology and Oceanography, 47, 1734-1741.
Macia, S. (2000). The effects of sea urchin grazing and drift algal blooms on a subtropical seagrass bed
community. Journal of Experimental Marine Biology and Ecology, 246, 53-67.
Mann, K. H. (1972). Ecological energetics of the seaweed zone in a marine bay on the Atlantic coast of
Canada. II. Productivity of the seaweed. Marine Biology, 14, 199-209.
Grazing on benthic primary producers
Mann, K. H. (2000). Ecology of Coastal Waters. Second Edition. Blackwell Science, Malden,
Massachussets, USA.
Mariani, S. & Alcoverro, T. (1999). A multiple-choice feeding-preference experiment utilising seagrass
with a natural population of herbivorous fishes. Marine Ecology Progress Series, 189, 295-299.
Masini, R. J., Anderson, P . K. & McComb, A. J. (2001). A Halodule-dominated community in a
subtropical embayment: physical environment, productivity, biomass and impact of dugong grazing.
Aquatic Botany, 71, 179-197.
McClanahan, T. R., Kamukuru, A. T., Muthiga, N. A., Yebeo, M. G. & Obura, D. (1996). Effect of sea
urchin reductions on algae, coral and fish populations. Conservation Biology, 1, 136-154.
McComb, A. J. & Humphries, R. (1992). Loss of nutrient from catchments and their ecological impacts in
the Peel-Harvey estuarine system, Western Australia. Estuaries, 15, 529-537.
McGlathery, K. J. (1995). Nutrient and grazing influences on a subtropical seagrass community. Marine
Ecology Progress Series, 122, 239-252.
McGlathery, K. J., Krause-Jensen, D., Rysgaard, S. & Christensen, P. B. (1997). Pattern of ammonium
uptake within dense mats of the filamentous macroalgae Chaetomorpha linum. Aquatic Botany, 59,
Mcintyre, I. G., Graus, R. R., Reinthal, P. N., Littler, M. M & Littler D. S. (1987). The Barrier Reef
sediment apron: Tobacco Reef, Belize. Coral Reefs, 6, 1-12.
Montagna, P. A. (1984). In situ measurement of meiobenthic grazing rates on sediment bacteria and
edaphic diatoms. Marine Ecology Progress Series 18, 119-130.
Morrison, D. (1988). Comparing fish and urching grazing in shallow and deeper coral reef algal
communities. Ecology, 69, 1367-1382.
Mukai, H. & Nojima, S. (1985). A preliminary study on grazing and defecation rates of a seagrass grazer,
Tripneustes gratilla (Echinodermata:Echinoidea), in a Papua New Guinean seagrass bed. Special
Publications of the Mukaishishima Marine Biological Station (pp. 184-191)
Musick, J. A. & Limpus, C. J. (1997) Habitat utilization and migration in juvenile sea turtles. In: P. L.
Lutz & J. A. Musick (Eds), The biology of sea turtles (pp. 137-163). CRC Press, Boca Raton, Florida.
Neckles, H. A., Koepfler, E. T., Haas, L. W., Wetzel, R. L. & Orth, R. J. (1994). Dynamics of epiphytic
photoautotrophs and heterotrophs in Zostera marina (eelgrass) microcosms: responses to nutrient
enrichment and grazing. Estuaries, 17, 597-605.
Nedergaard, R. I., Risgaard-Petersen, N. & Finster, K. (2002). The importance of sulfate reduction
associated with Ulva lactuca thalli during decomposition: a mesocosm experiment. Journal of
Experimental Marine Biology and Ecology, 275, 15-29.
Newell, R, Field, J. & Griffiths, C. (1982). Energy Balance and significance of microorganisms in a kelp
bed community. Marine Ecology Progress Series, 8, 103-113.
Nicotri, M. E. (1977) Grazing effects of four marine intertidal herbivores on the microflora. Ecology, 58,
Niell, F. X., Fernandez, C., Figueroa, F. L., Figueiras, F. G., Fuentes, J. M., Perez-Llorens, J. L., GarciaSanchez, M. J., Hernandez, I., Fernandez, J. A., Espejo, M., Buela, J., Garcia-Jimenez, M. C., Clavero,
V. & Jimenez, C. (1996). Spanish Atlantic Coasts. In W. Schramm & P. H. Nienhuis (Eds.), Marine
benthic vegetation: recent changes and the effects of eutrophication (pp. 265-281). Berlin: Springer.
Nielsen, S. L., Enriquez, S., Duarte, C. M. & Sand-Jensen, K. (1996). Scaling maximum growth rates
among photosynthetic organisms. Functional Ecology 10, 167-175.
Nienhuis, P. H. (1996). The north sea coasts of Denmark, Germany and the Netherlands. In W. Schramm &
P. H. Nienhuis (Eds.), Marine benthic vegetation: recent changes and the effects of eutrophication (pp.
187-222). Berlin: Springer.
Nienhuis, P. H. & Groenendijk, A. (1986). Consumption of eelgrass (Zostera marina) by birds and
invertebrates: an annual budget. Marine Ecology Progress Series, 29, 29-35.
Nixon, S. W. (1995) . Coastal marine eutrophication: a definition, social causes and future concerns.
Ophelia, 41, 199-219.
Norkko, A. (1998). The impact of loose-lying algal mats and predation by the brown shrimp Crangon
crangon (L.) on infaunal prey dispersal and survival. Journal of Experimental Marine Biology and
Ecology, 221, 99-116.
Norkko, A. & Bonsdorff, E. (1996). Rapid zoobenthic community responses to accumulations of drifting
algae. Marine Ecology Progress Series, 131, 143-157.
Norkko, A., Bonsdorff, E. & Norkko, A. (2000). Drifting algal mats as an alternative habitat for benthic
invertebrates: species-specific responses to a transient resource. Journal of Experimental Marine
Biology and Ecology, 248, 79-104.
Ogden, J.C., Robinson, L. Whitlock, K., Daganhardt, H. & Cebula, R. (1983). Diel foraging patterns in
juvenile green turtles (Chelonia mydas, L.) in St. Croix United States Virgin Islands. Journal of
Experimental Marine Biology and Ecology, 66, 199-205.
Orio, A. A. & Donazzolo, R. (1987). Toxic and eutrophicating substances in the lagoon and Gulf of Venice
(in Italian). Istituto Veneto di Scienze, Lettere ed Arti, 11, 149-215.
Orth, R. J. (1977). The importance of sediment stability in seagrass communities. In B. C. Coull (Ed.),
Ecology of Marine Benthos (pp. 281-300). University of South Carolina Press, Columbia.
Orth, R. J., Heck, K. L. & van Montfrans, J. (1984). Faunal components in seagrass beds: a review of the
influence of plant structure and prey characteristics on predator-prey relationships. Estuaries, 7, 339350.
Osterling, M. & Phil, L. (2001). Effects of filamentous green algal mats on benthic macrofaunal functional
feeding groups. Journal of Experimental Marine Biology and Ecology, 263, 159-183.
Pavia, H., Carr, H. & Aberg, P. (1999). Habitat and feeding preferences of crustacean mesoherbivores
inhabiting the brown seaweed Ascophyllum nodosum (L.) Le Jol. And its epiphytic algae. Journal of
Experimental Marine Biology and Ecology, 236, 15-32.
Peckol, P. & Rivers, J. S. (1995). Physiological responses of the opportunistic macroalgae, Cladophora
vagabunda and Gracilaria tikvahiae, to environmental disturbances associated with eutrophication.
Journal of Experimental Marine Biology and Ecology, 190, 1-16.
Peckol, P. & Rivers, J. S. (1996). Contribution by macroalgal mats to primary production of a shallow
embayment under high and low nitrogen-loading rates. Estuarine, Coastal and Shelf Science, 43, 311325.
Peckol, P., DeMeo-Anderson, B., Rivers, J., Valiela, I., Maldonado, M. & Yates, J. (1994). Growth,
nutrient uptake capacities and tissue constituents of the macroalgae Cladophora vagabunda and
Gracilaria tikvahiae related to site-specific nitrogen loading rates. Marine Biology, 121, 175-185.
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.
Pignatti, S. (1962). Associations of marine algae on the Venice littorals (in Italian). Memorie Istituto
Veneto di Scienze, Lettere ad Arti, 32, 1-134.
Pihl, L., Isaksoon, I., Wennhage, H. & Moksnes, P. O. (1995). Recent increases of filamentous algae in
shallow Swedish Bays: effects on the community structure of epibenthic fauna and fish. Netherlands
Journal of Aquatic Ecology, 29, 349-358.
Poole, L. J. & Raven, J. A. (1997). The biology of Enteromorpha. Progress in Phycological Research, 12,
Posey, M., Powell, C., Cahoon, L. & Lindquist, D. (1995). Top-down vs. bottom up control of benthic
community composition on an intertidal flat. Journal of Experimental Marine Biology and Ecology,
185, 19-31.
Posey, M. H., Alphin, T. D., Cahoon, L., Lindquist, D. & Becker, M. E. (1999). Interactive effects of
nutrient additions and predation on infaunal communities. Estuaries, 22, 785-792.
Preen, A. (1995). Impacts of dugong foraging on seagrass habitats: observational and experimental
evidence for cultivation grazing. Marine Ecology Progress Series, 124, 201-213.
Raffaelli, D. G., Raven, J. H. & Poole, L. J. (1998). Ecological impact of green macroalgal blooms.
Oceanography and Marine Biology: an Annual Review, 36, 97-125.
Raposa, K. B. & Oviatt, C. A. (2000). The influence of contiguous shoreline type, distance from shore, and
vegetation biomass on nekton community structure in eelgrass beds. Estuaries, 23, 46-55.
Reise, K., Heere, E. & Sturm, K. (1989). Historical changes in the benthos of the Wadden Sea around the
island of Sylt in the North Sea. Helgolander Meeresunterschungen, 43, 417-433.
Romero, J., Niell, F. X., Martinez-Arroyo, A., Perez, M. & Camp, J. (1996). The Spanish Mediterranean
coasts. In W. Schramm & P. H. Nienhuis (Eds.), Marine benthic vegetation: recent changes and the
effects of eutrophication (pp. 295-306). Berlin: Springer.
Rooker, J. R., Holt, S. A., Soto, M. A. & Golt, G. J. (1998). Postsettlement patterns of habitat use by
Sciaenid fishes in subtropical seagrass meadows. Estuaries, 21, 318-327.
Rosenberg, R., Elmgren, R., Fleischer, S., Jonsson, P., Persoon, G. & Dahlin, H. (1990). Marine
eutrophication case studies in Sweden. Ambio, 19, 102-108.
Ross, S. W. & Moser, M. L. (1995). Life history of juvenile gag, Mycteroperca microlepis, in North
Carolina estuaries. Bulletin of Marine Science, 56, 222-237.
Ruesink, J. L. (2000). Intertidal mesograzers in field mesocosm: linking laboratory feeding rates to
community dynamics. Journal of Experimental Marine Biology and Ecology, 248, 163-176.
Russ, G. R. & Alcala, A. C. (1989). Effects of intense fishing pressure on an assemblage of coral reef
fishes. Marine Ecology Progress Series, 56, 13-27.
Grazing on benthic primary producers
Sagasti, A., Schaffner, L. C. & Duffy, J. E. (2001). Effects of periodic hypoxia on mortality, feeding and
predation in an estuarine epifaunal community. Journal of Experimental Marine Biology and Ecology,
258, 257-283.
Sala, E., Ribes, M., Hereu, B., Zabala, M., Alva, V., Coma, R. & Garrabou, J. (1998). Temporal variability
in abundance of the sea urchins Paracentrotus lividus and Arbacia lixula in the northwestern
Mediterranean: Comparison between a marine reserve and an unprotected area. Marine Ecology
Progress Series, 168, 135-145.
Salemaa, H. (1987). Herbivory and microhabitat preferences of Idotea spp. (Isopoda) in the northern Baltic
Sea. Ophelia, 27, 1-15.
Sammarco, P. W., Levinton, J. S. & Ogden, J. C. (1977). Grazing and control of coral reef community
structure by Diadema antillarum Philippi (Echinodermata:Echinoidea): a preliminary study. Journal of
Marine Research, 32, 47-53.
Sand-Jensen, K. & Borum, J. (1991). Interactions among phytoplankton, periphyton, and macrophytes in
temperate freshwaters and estuaries. Aquatic Botany, 41, 137-175.
Schramm, W. (1996). The Baltic Sea and its transition zones. In W. Schramm & P. H. Nienhuis (Eds.),
Marine benthic vegetation: recent changes and the effects of eutrophication (pp. 131-164). Berlin:
Schramm, W. & Booth, W. (1981). Mass bloom of the alga Cladophora prolifera in Bermuda: productivity
and phosphorus accumulation. Botanica Marina, 24, 419-426.
Sfriso, A. (1987). Flora and vertical distribution of macroalgae in the Lagoon of Venice: a comparison with
previous studies. Giornale Botanico Italiano, 121, 69-85.
Sfriso, A., Pavoni, B. & Marcomini, A. (1989). Macroalgae and phytoplankton standing crops in the
central Venice Lagoon: Primary production and nutrient balance. Science of the Total Environment,
80: 139-159.
Sfriso, A., Pavoni, B., Marcomini, A. & Orio, A. A. (1992). Macroalgae, Nutrient Cycles and Pollutants in
the Lagoon of Venice. Estuaries, 15, 517-528
Sfriso, A. & Marcomini, A. (1997). Macrophyte Production in a Shallow Coastal Lagoon. Part I: Coupling
with Chemico-Physical Parameters and Nutrient Concentrations in Waters. Marine Environmental
Research, 44, 351-375.
Short, F. T., Burdick, D. M. & Kaldy, J. E. III. (1995). Mesocosm experiments to quantify the effects of
eutrophication on eelgrass, Zostera marina. Limnology and Oceanography, 40, 740-749.
Sogard, S. M. (1992). Variability in growth rates of juvenile fishes in different estuarine habitats. Marine
Ecology Progress Series, 85, 35-53.
Sogard, S. M. & K. W. Able (1991). A comparison of eelgrass, sea lettuce macroalgae and marsh creeks as
habitats for epibenthic fishes and decapods. Estuarine, Coastal and Shelf Science, 33, 501-519.
Stelzer, R. S. & Lamberti, G. A. (2002) Ecological stoichiometry in running waters: periphyton chemical
composition and snail growth. Ecology 83, 1039-1051.
Steneck, R. S. & Dethier, M. N. (1994). A functional group approach to the structure of algal-dominated
communities. Oikos, 69, 476-498.
Sterner, R. W. & Hessen, D. O. (1994). Algal nutrient limitation and the nutrition of aquatic herbivores.
Annual Review of Ecology and Systematics 25, 1-29.
Sterner, R. W. & Elser, J. J. (2002). Ecological Stoichiometry: The Biology of Elements from Molecules to
the Biosphere. First edition. Princeton University Press, Princeton, USA.
Sugihara, G. & May, R. M. (1990) Applications of fractals in ecology. Trends in Ecology and Evolution.
3, 79-86.
Tagliapietra, D., Pavan, M. & Wagner, C. (1998). Macrobenthic community changes related to
eutrophication in Palude della Rosa (Venetian Lagoon, Italy). Estuarine, Coastal and Shelf Science,
47, 217-226.
Taylor, D., Nixon, S., Granger, S. & Buckley, B. (1995a). Nutrient limitation and the eutrophication of
coastal lagoons. Marine Ecology Progress Series, 127, 235-244.
Taylor, D., Nixon, S. , Granger, S., Buckley, B., McMahon, J. & Lin, H. (1995b). Responses of coastal
lagoon plant communities to different forms of nutrient enrichment- a mesocosm experiment. Aquatic
Botany, 52, 19-34.
Taylor, D., Nixon, S., Granger, S & Buckley, B. (1999). Responses of Coastal Lagoon Plant Communities
to Levels of Nutrient Enrichment: a Mesocosm Study. Estuaries, 22, 1041-1056.
Thayer, G. W., Bjorndal, K. A., Ogden, J., Williams, S. & Zieman, J. (1984). Role of larger herbivores in
seagrass communities. Estuaries, 7, 351-376.
Thybo-Christensen, M., Rasmussen, M. B. & Blackburn, T. H. (1993). Nutrient fluxes and growth of
Cladophora sericea in a shallow Danish bay. Marine Ecology Progress Series, 100, 273-281.
Tribble, G. W. (1981). Reef-based herbivores and the distribution of two seagrasses (Syringodium
filiforme and Thalassia testudinum) in the San Blas Islands (Western Caribbean). Marine Biology, 65,
Urabe, J., Kyle, M., Makino, W., Yoshida, T., Andersen,T. & Elser, J. J. (2002). Reduced light increases
herbivore production due to stoichiometric effects of light/nutrient balance. Ecology, 83, 619-627.
Vadas, R. L. & Steneck, R. S. (1995). Overfishing and inferences in kelp-sea urchin interactions. In H. R.
Skjoldal, C. Hopkins, K. E. Erikstad & H. P. Leinaas (Eds.), Ecology of Fjords and Coastal Waters,
(pp. 509-524). Amsterdam: Elsevier.
Valentine, J. F. & Heck, K. L. Jr. (1991). The role of sea urchin herbivory in regulating subtropical
seagrass meadows: evidence from field manipulations in the northern Gulf of Mexico. Journal of
Experimental Marine Biology and Ecology, 154, 215-230.
Valentine, J. F. & Heck, K. L. Jr. (1999). Seagrass herbivory: evidence for the continued grazing of marine
grasses. Marine Ecology Progress Series 176: 291-302.
Valentine, J. F. & Heck, K. L Jr. (2001) The role of leaf nitrogen content in determining turtlegrass
(Thalassia testudinum) grazing by a generalized herbivore in the northeastern Gulf of Mexico. Journal
of Experimental Marine Biology and Ecology, 258, 65-86.
Valentine, J. F., Heck, K. L. Jr., Kirsch, K. D. & Webb, D. (2000) Role of sea urchin Lytechinus variegatus
grazing in regulating subtropical turtlegrass Thalassia testudinum meadows in the Florida keys (USA).
Marine Ecology Progress Series 200, 213-228.
Valiela, I. (1995). Marine Ecological Processes. Second edition. Springer-Verlag, New York, USA.
Valiela, I., Foremann, K., LaMontagne, M., Hersh, D., Costa, J., Peckol, P., DeMeo-Anderson, B.,
D'Avanzo, C., Babione, M., Sham, C., Brawley, J. & Lajtha, K. (1992). Couplings of watersheds and
coastal waters: sources and consequencesof nutrient enrichment in Waquoit Bay. Estuaries, 15, 443457.
Valiela, I., McClelland, J., Hauwxell, J., Behr, P. J., Hersh D., & Foreman, K. (1997a). Macroalgal blooms
in shallow estuaries: controls and ecophysiological and ecosystems consequences. Limnology and
Oceanography, 45, 1105-1118.
Valiela, I., Collins, G., Kremer, J., Lajtha, K., Geist, M., Seely, B., Brawley, J. & Sham C. H. (1997b).
Nitrogen loading from coastal watersheds to receiving estuaries: new method and application.
Ecological Applications, 7, 358-380.
Valiela, I., Geist, M., McClelland, J. & Tomasky, G. (2000a). Nitrogen loading from watersheds to
estuaries: verification of the Waquoit Bay Nitrogen Loading Model. Biogeochemistry, 49, 277-293.
Valiela, I., Tomasky, G., Hauxwell, J., Cole, M. L., Cebrian, J. & Kroeger, K. D. (2000b). Operationalizing
sustainability: management and risk assessment of land-derived nitrogen loads to estuaries. Ecological
Applications, 10, 1006-1023.
Van Katwijk, M. M., Schmitz, G. H. W., Gasseling, A. P. & van Avesaath, P. H. (1999). Effects of salinity
and nutrient load and their interaction on Zostera marina. Marine Ecology Progress Series, 190, 155165.
Van Tamelen, P. G. (1996). Algal zonation in tidepools: experimental evaluation of the role of physical
disturbance, herbivory and competition. Journal of Experimental Marine Biology and Ecology, 201,
Verlaque, M. (1987). Relations entre Paracentrotus lividus (Lamarck) et le phytobenthos de Mediterranee
Occidentale. In C. F. Boudouresque (Ed.), Colloque International sur Paracentrotus lividus et les
oursins comestibles (pp. 5-36). GIS Posidonie publications, Marseille, France.
Vermaat, J. & Verhagen, F. C. A. (1996). Seasonal variation in the intertidal seagrass Zostera noltii
Hornem: coupling demographic and physiological patterns. Aquatic Botany 52, 259-281.
Williams, S. L. (1988). Thalassia testudinum productivity and grazing by green turtles in a highly
disturbed seagrass bed. Marine Biology, 98, 447-455.
Williams, S. L. & Heck, K. L. Jr. (2001). Seagrass Community Ecology. In: M. D. Bertness, S. D. Gaines
& M. E. Hay (Eds.), Marine Community Ecology (pp. 317-338). Sinauer Associates, Sunderland.
Wilson, K. A., Able, K. W. & Heck, K. L. Jr. (1990). Predation rates on juvenile crabs in estuarine nursery
habitats: evidence for the importance of macroalgae (Ulva lactuca). Marine Ecology Progress Series,
58, 243-251.
Worcester, S. E. (1995). Effects of eelgrass beds on advection and turbulent mixing in a low current and
low shoot density environments. Marine Ecology Progress Series, 126, 223-232.
Worm, B., Lotze, H. K. & Sommer, U. (2000). Coastal food web structure, carbon storage and nitrogen
retention regulated by consumer pressure and nutrient loading. Limnology and Oceanography, 45,
Zieman, J. C., Iverson, R. L. & Ogden, J. C. (1984). Herbivory effects on Thalassia testudinum leaf growth
and nitrogen content. Marine Ecology Progress Series, 15, 151-158.
Grazing on benthic primary producers
J. Cebrián: Dauphin Island Sea Laboratory, Department of Marine Science,
University of Southern Alabama, 101 Bienville Boulevard, Dauphin Island, AL
36528, USA.
Coastal marine plant communities are, in contrast to the open sea where the only
primary producers are phytoplankton, made up by a number of different plant groups
that each contribute to total autotrophic biomass and production (Borum & SandJensen, 1996); Hauxwell and Valiela, Chapter 3). Coastal plant communities rank
among the most productive biomes on earth and large amounts of nutrients
(especially N and P) pass therefore through this compartment (Pedersen et al., chapter
1). Nutrients assimilated during plant growth are immobilized temporarily in living
biomass and detritus until they are mineralized through grazing or decomposition.
Incomplete or very slow decomposition of detrital components that are resistant to
decomposition may, however, lead to long term or permanent (i.e. over decades,
centuries) storage of organic matter and associated nutrients. Whether detritus-bound
nutrients are mineralized and recycled quickly or whether they become stored in
slowly degradable detritus over longer time scales is controlled by the rates and
patterns of the decomposition of plant detritus. The percentage of net primary
production that is lost through grazing versus decomposition differs among major
groups of marine primary producers (i.e. phytoplankton, macroalgae and seagrasses)
(Cebrián, 1999; Duarte, 1995; Duarte & Cebrián, 1996). Hence, slow-growing
macroalgae and seagrasses seem to lose a much smaller proportion of net primary
production to grazing than fast-growing macro- and microalgae and thus the
immediate fate of organic matter produced by the slower growing macrophytes is
therefore to end up as detritus. A large proportion of the primary production may
therefore enter the detrital pathway when coastal ecosystems are dominated by
macrophytes. The rate by which plant detritus decomposes depends on the chemical
composition, nutrient content and carbon quality of the litter (i.e., degradability or
“quality” of detritus) and therefore decomposition rates vary substantially among
plant types ranging from uni-cells to trees (Enríquez, Duarte, & Sand-Jensen, 1993).
Comparative decomposition studies of marine plants have also shown that detritus
from phytoplankton and fast-growing macroalgae generally decay faster than that
from slow-growing, perennial macroalgae and seagrasses (e.g., Buchsbaum, Valiela,
Swain, Dzierzeski, & Allen, 1991; Schmidt, 1980; Twilley, Ejdung, Romare, &
Kemp, 1986). These variations suggest that nutrients are recycled much faster in
systems dominated by the former groups (Duarte, 1995). This conclusion is based on
S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling:
The IInfluence of Primary Producers, 187-216.
© 2004 Kluwer Academic Publishers. Printed in the Netherlands.
G.T. Banta, M.F. Pedersen and S.L. Nielsen
the assumption that N and P are lost from the litter at the same rate as dry weight or
carbon. It is however evident, from the rich literature on decomposition of terrestrial
plant litter, that loss rates of N and P may differ substantially from those of dry
weight or carbon (e.g., Staaf & Berg, 1982). For example, nutrients may be
mineralized faster than organic matter are decomposed (i.e. net mineralization) or
they may be lost more slowly than organic matter are decomposed due to nutrient
assimilation by the associated microflora (i.e. net immobilization during
decomposition). A proper evaluation of mineralization of detritus-bound nutrients
through decomposition should therefore preferably be based upon observed changes
in the N and P pools of the detritus and not on mass changes in biomass. Such data
are unfortunately rare for marine plants.
The purpose of this chapter is to compare patterns of decomposition and
mineralization among different groups of marine primary producers in order to
evaluate whether dominance of certain plant groups (i.e. fast-growing ephemeral
versus slow-growing persistent types) affect the fate – that is, mineralization versus
long term storage - of detritus-bound nutrients in coastal marine ecosystems. Based
on a large data-set extracted from the literature, we will first test whether general
patterns of decomposition differ between major groups of marine plants. We intend
then to compare how major nutrients (N and P) are released during decomposition of
litter from different marine plants using data obtained by the authors. In order to
compare published data on decomposition in a systematic manner, we have chosen to
standardize these data using Westrich and Berner’s (1984) multi-G decomposition
model. Where possible, we apply this model specifically to carbon (C) and nutrient
(N and P) pools of the detritus to compare how materials, especially major nutrients,
are recycled via decomposition.
Decomposition of plant detritus can be viewed as occurring in stages (e.g., Aber &
Melillo, 2001), starting with a rapid leaching phase where labile matter (low
molecular-weight sugars and small amino-acids) are abiotically lost from the newly
dead cells, followed by a slower loss of material caused by microbial activity (i.e.,
decomposition). Substrate quality and thus, susceptibility to microbial activity, varies
among the different compounds contained in plant litter and these compounds are
therefore often decomposed at different rates. Small amino-acids and low molecularweight sugars are thus decomposed faster than high molecular-weight sugars and
cellulose, which are decomposed faster than lignin. Some of these compounds may
further contain fractions that are resistant to microbial degradation and plant litter
therefore often contains a refractory fraction that is only decomposed over very long
time scales. Decomposition and mineralization rates thus vary widely among detritus
originating from different plants groups because the composition (i.e., content of
nutrients, structural tissues etc.) of such litter may differ substantially.
Decomposition of litter is most often described as an exponential decrease in biomass
(G) with time (Olson, 1963):
Gt = G0 e − k t
(eq. 1)
Decomposition of marine primary producers
where t is time and k is the litter specific rate constant. This equation represents a
simplified description of decomposition since it is assumed that all substrates
contained in the litter are processed simultaneously and completely. This is of course
not true (see above) and a better alternative to this model is therefore Westrich and
Berner’s multi-G model (as modified in Westrich & Berner, 1984). Briefly, this
model assumes that detritus is made up by several organic fractions (G1, G2, etc.) that
decompose at different rates (ki) albeit at the same time. Thus, the overall pattern of
decomposition is described as the sum of the patterns of each component plus
eventually a refractory pool of matter (R) that does not appear to decompose within
the time studied:
Gt = G0 e − k0 t + G1 e − k1 t + G2 e − k2 t + ... + Gi e − ki t + R
(eq. 2)
Often, only one actively decomposing pool of organic matter can be deduced from
the decomposition data available in addition to an eventual refractory pool and in
such cases the model simplifies to:
Gt = G0 e − k t + R
(eq. 3)
where the amount of organic matter remaining at time t, Gt, is a function of the
exponential decay, with rate constant k, of a decomposable pool, G0, and a refractory
pool, R, of organic matter.
In this chapter we will focus on the k and R parameters as these indicate, respectively,
how rapidly organic matter is decomposed (or associated nutrients are mineralized)
and, to what extent organic matter (or nutrients) may be preserved in refractory
detritus. It is not, however, common practice to report both these parameters (i.e., k
and R) since most studies apply the simple model of exponential decay (equation 1).
The application of a model without a refractory pool (R) may, however, have
important consequences for the estimation of decomposition rates as can be illustrated
by the example in Figure 1. In this example, weight losses of litter (i.e.,
decomposition) proceeded for a period of approximately 150 days after which further
changes occurred very slowly. Thus, in this example there was a refractory, or
residual, pool of organic matter corresponding to ca. 13% of the original sample
weight. Two different models – one with a residual component included (equation 3)
and one without (equation 1) – were fitted to the same data and decomposition rates
(k) were estimated. The estimated value of k was 3.7 d-1 when a model including a
residual fraction was applied whereas it was 2.9 d-1 when a model without a residual
fraction was applied. Hence, the estimated decomposition rate (k) becomes severely
underestimated (by more than 20% in this case) when a refractory pool is present, but
not considered, since an exponential decomposition model without a refractory term
must approach zero by some time.
G.T. Banta, M.F. Pedersen and S.L. Nielsen
Single exponential decay + refractory pool
Single exponential decay
% Remaining
Gt = G0e-kt + R
G0 = 86%
G = G0e-kt
k = 3.7% d
k = 2.9% d-1
R = 14%
Figure 1. Application of simple G decomposition models (Westrich & Berner, 1984) with (left)
and without (right) a parameter describing the refractory pool (R).
The larger the size of the refractory pool, the larger the error related to estimates of k.
This can lead to poor model fits and large errors for estimates for decomposition rate
constants. We have therefore only included data in this review for which we were
able to determine both the decomposition rate constant (k) and the size of a refractory
pool (R) if present.
While the focus of this chapter is nutrient remineralization through decomposition,
there is remarkably little data collected on nutrient release from marine plant detritus
during decomposition and even fewer studies where rates of mineralization are
compared across marine plant types. Most studies on decomposition of marine
primary producers report decomposition rates, i.e. temporal changes in biomass
(expressed in units of dry weight or carbon) and not mineralization rates (see, e.g.,
Enríquez et al., 1993). We take therefore two different approaches in this chapter.
First, we briefly review the literature available on decomposition for marine primary
producers and compare and interpret these data in terms of its potential consequences
for nutrient release or storage, assuming that the element composition remains
constant during decomposition (i.e., that the rate and degree of N and P
mineralization is the same as that of C). Next, we present the limited data set of
directly measured changes in N and P pools during decomposition for selected plant
groups to test whether N and P are released faster or slower than organic matter is
decomposed. We will also present some of our own data to examine the
stoichiometry of C, N and P during the decomposition of a range of coastal
macrophytes. Finally, we will try to illustrate the consequences of these differences
in decomposition patterns for nutrient retention on coastal ecosystems based on two
case studies based on two Danish estuaries which differ in macrophyte communities.
The information reviewed in this chapter should be interpreted in light of what are the
consequences of changes in coastal plant communities for nutrient cycling, a main
theme of this book, via the detrital decomposition pathway.
Decomposition of marine primary producers
Decomposition parameters k and R for different groups of coastal marine primary
producers that could be extracted from the available literature are summarized in
Figure 2. Only data sets where it was possible to assess the size of R (whether =0 or
not) were included in this review. Decomposition rate constants (k) range from
0.0006 to 0.71 d-1 across the entire data set and vary substantially within each plant
group. Some general patterns among the three plant groups are evident, however.
The median decomposition rate for of seagrass litter (median value = 0.020 d-1) is
significantly lower than that of detritus originating from macro- and microalgae
(median values = 0.042 d-1 and 0.046 d-1, respectively) (Kruskal Wallis, p = 0.003).
The decomposition rates for microalgae and macroalgae were quite similar, however,
both in terms of the range observed and media values, and were not distinguishably
It is evident from Figure 2 that the detritus from all the major plant groups often
contain an apparently refractory pool of detritus (R). This is not always the case,
however, as R was equal to 0 (i.e., no refractory pool could be detected) for nearly
half (46%) of the litter decomposition curves we fitted. The size of R ranged widely
from 0% to 86% of the total detrital biomass across the entire data set. Surprisingly,
macroalgae had significantly smaller refractory pools (median = 0%) than either
seagrasses (median = 16%) or microalgae (median = 33%) (Kruskal Wallis, p =
0.006). The latter two groups did not differ significantly in the size of the refractory
pool due to the surprisingly high refractory pools estimated for microalgae.
Part of the explanation for the relatively large refractory pools for microalgae may be
due to the fact that studies of microalgae decomposition were generally shorter in
duration than those for larger macrophytes and decomposition may not be fully
completed, making it difficult to estimate the final stages of decomposition. This
may lead to an overestimate of R. Another factor, however, that may lead to an
underestimate of R for macrophyte detritus (both macroalgae and seagrasses) is that
the litter bag technique typically used for studying macrophyte detritus is inherently
different than the techniques used for studying the decomposition of microalgae (e.g.,
culture studies, flask incubations, radioactive labeling experiments). With the litter
bag technique, once particles are reduced to below the mesh size of a litter bag they
are likely to be lost and counted as decomposed, even if they may not truly be fully
decomposed. In contrast, the techniques used with microalgae are more likely to
fully account for all particulate microalgae detritus until it is completely decomposed.
Still, it is apparent that all detritus types, including microalgae, often contain
refractory pools which may be significant when considering long-term nutrient
Benthic microalgae
Twilley et al., 1986
Lee & Fisher, 1992
Andersen, 1996
Andersen & Kristensen, 1992
Westrich & Berner, 1984
Garber, 1984
Thalassiosira angstii
Scrippsiella trochoidea
Skeletonema costatum
Chaetoceros tricornutum
Skeletonema costatum
Natural community (dominated
by the diatom Leptocylindrus
Chlorella spp.
Thalassiosira psuedonana
Skeletonema costatum
Benthic penate diatoms
N data?
R >0?
estuarine phytoplankton
Scenedesmus sp.
Primary producer
Jewell & McCarty, 1971
Otsuki & Hanya, 1972a; Otsuki &
Hanya, 1972b
R. C. Newell, Lucas, & Linley, 1981
P data?
Table 1. Studies included in our review of decomposition of estuarine primary producers (Figs. 2&3). Indicated is what type of primary producer was studied,
whether it was possible to explicitly determine that the refractory fraction (R) was >0 (in some cases* it was necessary to assume R=0 to fit the data). Also
indicated is whether changes in nutrient content (N & P) during decomposition were reported.
G.T. Banta, M.F. Pedersen and S.L. Nielsen
Rice & Tenore, 1981
Birch, Gabrielson, & Hamel,
Gabrielson, Birch, & Hamel, 1983
Twilley et al., 1986
Buchsbaum et al., 1991
Kristensen, Andersen, & Blackburn, 1992
Christensen et al., 1994
Kristensen, 1994
Bourgues, Auby, deWit, & Labourg, 1996
Chaeteromorpha linum
Ulva lactuca
Fucus vesiculosus
Monostroma obscurum
Ulva lactuca
Ulva lactuca
Gracilaria tikvahiae
Fucus vesiculosus
Chondrus crispus
Josselyn & Mathieson, 1980
Fucus vesiculosus
Macrocystis intergrifolia
Nerocystis luetkeana
Ascophyllum nodosum
Fucus vesiculosus
Gracelaria foliifera
Hypnea musciformis
Sargassum filapendula
Spatoglossum schroederi
Cladophora albida
Hunter, 1976
Albright, Chocair, Masuda, & Valdés, 1980
Primary producer
Table 1 (continued)
N data?
R >0?
P data?
Decomposition of marine primary producers
Pedersen, unpublished
Godshalk & Wetzel, 1978
Josselyn & Mathieson, 1980
Rice & Tenore, 1981
Pellikaan, 1982
Rublee & Roman, 1982
Gallagher, Kibby, & Skirvin, 1984
Kenworthy & Thayer, 1984
Kristensen, 1994
Brouwer, 1996
Bourgues et al., 1996
Paalme, Kukk, Kotta, & Orav, 2002
Primary producer
Table 1 (continued)
Zostera marina
Zostera marina
Thalassia testudinum
Syringodium filiform
Zostera marina
Thalassia testudinum
Zostera marina
Zostera marina
Zostera marina
Ruppia maritima
Desmarestia anceps
Monostroma obscurum
Cladorphora glomerata
Pilayella littoralis
Ulva lactuca
Polysiphonia fucoids
Fucus vesiculosus
Ceramium rubrum
Sargassum muticum
Halidrys sliquosa
N data?
R >0?
P data?
G.T. Banta, M.F. Pedersen and S.L. Nielsen
Halophila stipulacea
Ruppia maritima
Zostera marina
Zostera noltii
Cymodocea nodosa
Cymodocea nodosa
Zostera marina
Zostera noltii
Posidonia oceanica
Enhalus acoroides
Zostera marina
(including roots)
Peduzzi & Herndl, 1991
Christensen et al., 1994
Bourgues et al., 1996
Mateo & Romero, 1997
Holmer & Olsen, 2002
Thalassia testudum
S. Y. Newell, Fell, Statzelltallman,
Miller, & Cefalu, 1984
Wahbeh & Mahasneh, 1985
Twilley et al., 1986
Buchsbaum et al., 1991
Hemminga & Nieuwenhuize, 1991
Primary producer
Table 1 (continued)
R >0?
N data?
P data?
Decomposition of marine primary producers
G.T. Banta, M.F. Pedersen and S.L. Nielsen
R (%)
k (d-1)
c ro
Figure 2. Box plot of decomposition parameters (k and R) for different groups of marine
primary producers based on changes in biomass (dry weight) or carbon (C). Note log scale for
k-values (left). Boxes encompass 25 and 75 quartiles while median values are shown as
horizontal lines within each box. Error bars represent the 5 and 95 percentiles. Data outside
the 5 and 95% percentiles are plotted as single observations. Note log scale for k (left). The
number of observations is 38, 58 and 20 for seagrasses, macroalgae and microalgae,
respectively. The studies included are listed in Table 1.
The studies reviewed here show that detritus from micro- and macroalgae are
decomposed faster than detritus originating from seagrasses, which corresponds very
well to the findings of Enríquez et al. (1993) and Cebrián (1999) although their datasets on marine plants were smaller than the one presented here. In addition, it is
evident from our review that macroalgae have a tendency to decompose more
completely than other detritus types. Thus, the slower and more incomplete
decomposition of litter from slow-growing macrophytes (seagrasses in particular)
may lead to a larger accumulation of dead organic matter in systems dominated by
seagrasses when compared to systems dominated by macro- and microalgae provided
that the net production per unit area is the same (Cebrián, 1999; Duarte & Cebrián,
1996). Seagrasses especially may therefore act as a sink for carbon due to their slow
decomposition and the preservation of refractory organic matter.
As previously mentioned, nutrient mineralization patterns should be evaluated from
changes in nutrient pools of the detritus during decomposition rather than from
changes in biomass of that detritus because nutrients may be lost at different rates
than biomass is decomposed. There are, however, fewer available data sets on
changes in pool size of major nutrients during decomposition so estimates for the
decomposition parameters (k and R) expressed in units of N and P are therefore less
common. The data we could find where it was possible to estimate k and R for N
pools are presented in Figure 3. The k-values based on temporal changes in N-pools
span a similar range as those based on changes in biomass (Figure 2) although it was
Decomposition of marine primary producers
not possible to detect any significant differences among plant groups (Kruskal Wallis,
p = 0.30). This is despite the fact that it appeared that N in microalgae decomposes
more rapidly (median = 0.068 d-1) compared to seagrasses and macroalgae (median =
0.040 d-1 and 0.038 d-1, respectively). The highest values of k based on N tend to be
higher than those based on changes in biomass (kC), but it is difficult to make a
proper comparison between the two data sets as there are fewer studies where we
could assess N dynamics.
R (%)
k (d-1)
Figure 3. Box plot of decomposition parameters (k and R) based on changes in nitrogen (N)
pools in different groups of marine primary producers. Note log scale for k-values (left).
Boxes encompass 25 and 75 quartiles while median values are shown as horizontal lines
within each box. Error bars represent the 5 and 95 percentiles. Data outside the 5 and 95%
percentiles are plotted as single observations. Data where extremely high k-values (> 1 d-1)
indicated leaching, not decomposition losses, were excluded. The number of observations is
29, 29 and 6 for seagrasses, macroalgae and microalgae, respectively. The studies included
are listed in Table 1.
Examining the size of the refractory N-pool (R) in those studies, we observed that
macroalgae (median = 0%) again had significant smaller refractory N-pools than
microalgae (median = 31%) (Kruskal Wallis, p = 0.04). Seagrasses had intermediate
sizes of refractory N-pools (median = 21%) and were not significantly different than
the other two plant groups. This was similar to the pattern we observed for organic
matter pools.
There were even fewer studies where we could determine the decomposition
parameters for P-pools (Fig. 4). It was not possible to detect significant differences
among plant groups for either k-values or R (Kruskal Wallis p = 0.51 and 0.15,
respectively) for P data, given in part the fewer data points. The patterns suggested in
Figure 4 are different than those observed for organic matter (Fig. 2) or N (Fig. 3),
however, with kP’s being greatest for microalgae (median = 0.13 d-1) and RP’s being
lowest (median = 4 %) compared to macroalgae and seagrasses. This pattern is more
typical of what one would have expected due to differences in composition (i.e.,
degree of structural components such as cellulose and lignin) of the different plant
G.T. Banta, M.F. Pedersen and S.L. Nielsen
groups. As expected, microalgae lose P more rapidly and completely compared to
macrophytes. Macroalgae and seagrasses lose P at similar rates (k medians = 0.039 d1
and 0.047 d-1, respectively) while that latter group has the largest amount of P
retained in refractory pools (R medians = 16% and 25%, respectively).
R (%)
k (d-1)
Figure 4. Box plot of decomposition parameters (k and R) based on changes in phosphorous
(P) pools in different groups of marine primary producers. Note log scale for k-values (left).
Boxes encompass 25 and 75 quartiles while median values are shown as horizontal lines
within each box. Error bars represent the 5 and 95 percentiles. Data outside the 5 and 95%
percentiles are plotted as single observations. Data where extremely high k-values (> 1 d-1)
indicated leaching, not decomposition, losses were excluded. The number of observations is
15, 24 and 4 for seagrasses, macroalgae and microalgae, respectively. The studies included
are listed in Table 1.
Comparing the decomposition dynamics for the different elements within coastal
marine plant groups, it seems that C, N and P are decomposed and mineralized at
similar rates for macroalgae as k-values for macroalgae in Figures 2-4 are similar. It
appears that P may be retained somewhat more than C or N as the median value of R
for P (16%) is larger than for either organic matter (C) or N (both medians = 0%).
There is a great deal of variability in R values for macroalgae, however, indicating a
wide range of nutrient retentions in refractory pools for macroalgae.
In contrast to macroalgae, the decomposition dynamics for C, N and P are not similar
for either seagrasses or microalgae. Mineralization rates for N and P are generally
faster (i.e., higher k-values) than for C for both plant groups, especially for P. For
seagrasses there is also a tendency for there to be a larger refractory pool of N and P
compared to C. That is not the case for microalgae, however, where there are smaller
refractory pools, especially of P. This conclusion is based on precariously few data
sets, however, and should be considered cautiously. In any case, it is clear that
mineralization dynamics for nutrients cannot be easily predicted based on patterns for
organic matter decomposition for seagrasses and microalgae, while it appears that one
might have better success with macroalgae.
Decomposition of marine primary producers
Trying to summarize from this review of the literature, we conclude that the
mineralization dynamics of nutrients do differ among coastal marine plant groups.
The decomposition of microalgae is generally more rapid than for macrophytes and
nutrient mineralization rates for microalgae are more rapid than organic matter
decomposition rates. There does appear to be the potential for nutrient retention (N
in particular) in a refractory component of microalgae detritus and this is a
phenomenon that warrants further research. Seagrasses on the other hand decompose
more slowly although they do exhibit higher rates of nutrient mineralization (which
still are lower than microalgae, though) compared to rates of C decomposition.
Seagrasses thus have a greater potential to retain nutrients than microalgae, both due
to the slower rates of mineralization and to significant refractory pools. Macroalgae
are quite variable, but can characterized as having relatively rapid rates of organic
matter decomposition and nutrient mineralization and containing small refractory
pools. Thus, they appear to have a much lower potential to retain nutrients than the
other plant groups. These conclusions are greatly generalized given the fact that they
are based on a wide range of data sets collected for many species under differing
circumstances and should be examined more specifically, based on the specific plant
groups in various environments. Still, they do present an overall hypothesis worth
considering about the importance and significance of the various plant groups in
coastal environments in terms of their role for nutrient retention.
The data reviewed in Figures 2-4 suffer from substantial variation within each plant
group which makes specific comparisons of decomposition dynamics somewhat
unreliable. Part of the variation can be attributed to species-specific variations within
each group. For example, morphological, physical and functional properties that may
serve to determine decomposition patterns vary enormously among different types of
macroalgae and great variations in k and R-values are therefore to be expected. Part
of the observed variation may further be due to the variable conditions under which
these data were collected by different authors (i.e. in situ versus culture studies,
aerobic versus anaerobic conditions, variable temperatures, etc.). These sources of
variation make it difficult to precisely compare decomposition parameters (i.e. k and
R) between groups and difficult to compare specific patterns of decomposition and
mineralization for different plant groups. Overall, the results in Figures 2-4 do,
however, suggest general differences among plant groups and do indicate that
decomposition patterns for organic matter, N and P are not the same and need to be
assessed individually.
Surprisingly, relatively few studies have systematically compared the changes that
occur in the C, N and P pools of detritus during decomposition in marine plants,
which is in contrast to the extensive work published on decomposition of terrestrial
and salt marsh detritus (e.g., Melillo, Aber, & Muratore, 1982; Staaf & Berg, 1982;
Valiela et al., 1985). The following study was therefore designed to allow a direct
comparison of decomposition and mineralization patterns among four species of
marine macrophytes, which represent different functional types of the most common
macrophytes in temperate (specifically Northern European) coastal waters. Briefly,
G.T. Banta, M.F. Pedersen and S.L. Nielsen
we investigated the anaerobic decomposition of two species of ephemeral macroalgae
(Ulva lactuca, Ceramium rubrum), one species of persistent macroalgae (Fucus
vesiculosus) and one seagrass species (Zostera marina, both below- and aboveground organs were considered) using litterbags (1 mm mesh) which were buried in
estuarine sediments in the laboratory for nearly two years. We did not include
microalgae in this study, even though phytoplankton and other microalgae often
contribute significantly to total production of organic matter in coastal ecosystems,
because the litter bag technique is not well-suited to study the compositional changes
during decomposition of such small cells.
Long term changes in carbon biomass (i.e., decomposition) are shown in Figure 5
while decomposition rates and the estimated size of refractory pools are given in
Table 2. Decomposition rates based on changes in carbon biomass varied
substantially and were fastest for the two ephemeral species and slowest for the
below-ground parts of Zostera marina while the decay rates of detritus from slowgrowing macroalgae Fucus vesiculosus and leaves of Zostera were intermediate
between those extremes (Table 2).
% C remaining
Ulva fit
Fucus fit
Ceramium fit
Zostera roots
Zostera fit
Zostera roots fit
Figure 5. Long term changes in carbon biomass during anaerobic decomposition of detritus
from three species of macroalgae (Ulva lactuca, Ceramium rubrum and Fucus vesiculosus;
left) and the seagrass Zostera marina (above- and below-ground organs; right). Data are
presented as % of the original sample size. Experimental temperature = 15° C and salinity =
15‰ (Banta, Pedersen and Nielsen, unpublished). The curves represent the fits of a 1-G
decomposition model with a refractory pool (equation 3). Parameters for these fits are
summarized in Table 2.
The pool size of refractory detritus differed also substantially between the detritus
types examined - detritus from Ulva lactuca and Ceramium rubrum was completely
decomposed after 9 months while there was 7%, of the original Fucus vesiculosus
detritus left. The seagrass species studied, Zostera marina, contained even larger
refractory pools with 15% and more than 50% of the detritus from Zostera leaves and
Zostera roots and rhizomes, respectively, remaining after almost two years (Fig. 5,
Table 2).
Ulva lactuca
Ceramium rubrum
Fucus vesiculosus
Zostera marina
Zostera marina
Primary producer
Green macroalgae
Red macroalgae
Brown macroalgae
Seagrass (leaves)
Seagrass (roots & rhizomes)
0.026 ± 0.011
0 ± 26
0.040 ± 0.010
0.038 ± 0.014
0.028 ± 0.005
0.033 ± 0.006
15 ± 3
0.013 ± 0.005
57 ± 4
0.003 ± 0.012
86 ± 80
0.053 ± 0.012
14 ± 3
0.005 ± 0.007
0 ± 79
0.022 ± 0.010
0.006 ± 0.003
26 ± 14
0.055 ± 0.007
12 ± 2
0.019 ± 0.010
3 ± 11
0.029 ± 0.015
0.039 ± 0.011
Table 2. Decomposition parameters k (d-1) and R (% of initial sample size) for selected estuarine macrophytes. Data were obtained by following long term
changes in C, N and P pools under anoxic conditions. Parameters were obtained by fitting the data to a 1-G decomposition model (Westrich and Berner 1984)
and are presented as mean ± SE. * No refractory pool was evident so R was set to 0 and a simple exponential decay was fitted for the data. (Banta, Pedersen &
Nielsen, unpublished).
Decomposition of marine primary producers
G.T. Banta, M.F. Pedersen and S.L. Nielsen
These patterns, as well as the specific decomposition parameters, are consistent with
the general patterns we observed in our review of macrophyte decomposition studies
(Figs. 2-4), although we are able to distinguish differences between the fast-growing
ephemeral and slow-growing perennial macroalgae species in our study.
Seagrass detritus, especially that from the underground parts, clearly stands out from
the other plant types with lower rates of decomposition and much greater refractory
pools. Given that the production of seagrasses roots and rhizomes often represents
20-50% of the total seagrass production (Duarte et al., 1998), the slower and less
complete decomposition of this detritus type may have significant consequences for
the accumulation of organic matter in shallow coastal ecosystems (Duarte &
Cebrián, 1996). Dominance of seagrasses and, to some extent also by slow-growing
macroalgae, may potentially have a similar effect on accumulation and burial of
plant-bound nutrients and, hence, on nutrient retention at the ecosystem level.
Ulva fit
Fucus fit
Ceramium fit
% N remaining
Ulva fit
Fucus fit
Ceramium fit
% P remaining
Zostera roots
Zostera fit
Zostera roots fit
Zostera roots
Zostera fit
Zostera roots fit
Figure 6. Long term changes in N (top) and P(bottom) pools during anaerobic
decomposition of detritus from three species of macroalgae (Ulva lactuca, Ceramium rubrum
and Fucus vesiculosus; left) and the seagrass Zostera marina (above- and below-ground
organs; right). Data are presented as % of the original sample size. Experimental
temperature = 15° C and salinity = 15‰ (Banta, Pedersen and Nielsen, unpublished). The
curves represent the fits of a 1-G decomposition model with a refractory pool (equation 3).
Parameters for these fits are summarized in Table 2.
Decomposition of marine primary producers
Whether or not mineralization will proceed with the same speed as decomposition
depends on the rate and extent by which nutrients are released from detritus during
decay. The decomposition patterns and parameters for N and P for the studied
macrophytes are presented in Figure 6 and Table 2. While the overall patterns
among macrophytes are similar to that of C, with the ephemeral macroalgae being
decomposed more rapidly and completely than the slower growing macrophytes,
there are noticeable differences in the decomposition dynamics of C, N and P for all
of the macrophytes.
Decay rates for N (kN) were generally lower than when expressed in units of carbon
(Table 2). This was especially evident for detritus from Fucus and the below-ground
parts of Zostera, which experienced a net increase in N-pool size during the initial
stages of decomposition. Substantial losses of N from Fucus detritus appeared first
after 2 months of decomposition, after which N was lost rapidly and almost
completely (RN not significantly different from 0). The N-pool within detritus from
Zostera roots/rhizomes remained, in contrast, nearly intact during 2 years of
decomposition (figure 7) resulting in a very low decomposition rate (kN = 0.003 d-1)
and a substantial refractory pool of detritus-bound N (RN = 86% of initial sample
size). Surprisingly, the loss rate of N in detritus from Zostera leaves was high, but
this detritus type contained also a significant refractory N pool (RN = 14%).
Decay rates for P (kP) were generally slower than for C but faster than for N (Table
2) but rates for the different types of detritus were ranked as for N, i.e., Zostera
leaves>Ulva>Ceramium>Fucus>>Zostera roots/rhizomes. Refractory pools of P
were generally of the same size as those for C except in the case of Zostera
roots/rhizomes where the refractory detritus pool was severely depleted in P.
The fact that both N and P mineralization generally occurred more slowly than C
decomposition in our study differs from the overall patterns observed in our review
of the decomposition literature (Figs. 2-4). There N and P mineralization either
occurs at similar rates (macroalgae) or more rapidly (microalgae and seagrasses)
than C decomposition. Part of the explanation for this apparent discrepancy lies in
the fact that the literature review allows general but not specific comparisons under
a given set of conditions. One of the many factors that can vary both between and
within species is their nutrient composition. Therefore we now will consider the
results of our decomposition study based on the specific nutrient composition of the
macrophytes studied.
Changes in nutrient composition during decomposition are most easily analyzed and
interpreted by examining changes in tissue nutrient ratios with time. Microbial
decomposers are nutrient rich and balanced bacterial growth require substrates with
an approximate C:N:P-ratio of about 106:12:1 (Goldman, Caron, & Dennett, 1987).
The nutritional quality of the substrate may thus determine whether microbial
activity becomes limited by lack of energy (i.e., carbon) or by nutrients and, hence,
determine how the elemental composition of detritus will change during
decomposition. Nutrient rich detritus may experience a net loss of nutrients relative
to carbon because the decomposers become carbon-limited and excrete excess
nutrients, while nutrient poor detritus may be enriched with nutrients because
G.T. Banta, M.F. Pedersen and S.L. Nielsen
C:P-ratio (molar)
C:N-ratio (molar)
decomposers contribute nutrients to the matrix of detritus and bacteria from the
surrounding medium (Swift, Heal, & Anderson, 1979). In extreme cases, the nutrient
pool size of detritus may remain almost intact or even increase while the total
biomass of detritus decreases (e.g., Staaf & Berg, 1982).
Different plant types differ considerably from each other in their chemical
composition (e.g., Duarte, 1992). Microalgae and fast-growing macroalgae tend to
be richer in both N and P than slow-growing macroalgae and seagrasses, because the
former groups contain more actively metabolizing tissues and the latter contains
more structural tissues. The same pattern was seen among the five detritus types
included here – fresh detritus from Zostera marina and Fucus vesiculosus were
deprived in both N and P relative to C, whereas detritus from Ulva and Ceramium
were relatively rich on these nutrients (Fig. 5). Tissue concentrations of N and P are
seasonally variable, but these plants were all sampled during mid-summer where N
and P concentrations are generally low. Tissue concentrations of N and P and N:Pratios indicated that the detritus from Ceramium was well balanced relatively to the
substrate requirements of bacteria, while Ulva lactuca and leaves of Zostera were
marginally deprived of both N and P relative to C. Detritus from Fucus vesiculosus
and the below-ground parts of Zostera were, in comparison, very poor in both
nutrients, particularly N (both had N:P-ratios ≤10).
Figure 7. Initial C:N and C:P ratios (mean ± s.d.) of Ulva lactuca, Ceramium rubrum, Fucus
vesiculosus and Zostera marina (above- and below-grounds organs, respectively) used in the
experiment (Banta, Pedersen and Nielsen, unpublished).
The nutritional quality of detritus from the different plant types leads us to expect
that losses of N and P during decomposition should follow losses of C in Ceramium,
but be slower than for C (i.e. increasing C:N and C:P-ratios) among the other species
and especially so for Fucus vesiculosus and the below-ground parts of Zostera.
The data in Figure 8 show that the C:N-ratio of detritus from Ceramium rubrum and
from leaves of Zostera marina did remain more or less constant during
decomposition. Carbon and nitrogen were thus lost at approximate similar rates for
these detritus types. Constant C:N-ratios and relatively fast decomposition rates for
both these detritus types suggest that the quality of this detritus was relatively high
as expected based on start C:N. The C:N-ratios decreased, in contrast, significantly
Decomposition of marine primary producers
with time in detritus originating from Ulva lactuca, Fucus vesiculosus and the roots
and rhizomes of Z. marina indicating a more rapid loss of C relative to N or,
alternatively, that microbial immobilization of N during the decomposition was able
to enrich these N-poor detritus types. The tissue concentration of C tended to
decrease with time for all three detritus types while N concentrations increased
significantly (data not shown). The increase in N concentration within detritus from
Fucus and from the below-ground parts of Zostera was even so large that the total
pool size of nitrogen increased to more than 100% of the initial size during early
stages of decomposition (Fig. 6, top). The detritus of Ulva, Fucus and below-ground
parts of Zostera experienced thus a net enrichment of N during decomposition,
which may be important since it may slow down the rates by which net
mineralization from these detritus types occur. These changes during decomposition
and colonization by microorganisms are consistent with starting C:N ratios observed
for those detritus types
With the exception of the below-ground portions of Zostera, there were no marked
differences in the relative size of the refractory pools of C and N among the different
macrophytes (Table 2). The refractory pool of Zostera root and rhizomes did contain
substantially higher amounts of N than fresh detritus (much lower C:N ratio, Fig. 8),
however, which may be important in relation to burial and retention of N.
The C:P-ratios of detritus originating from Ceramium and Ulva decreased during
decomposition (Fig. 8) showing that C was lost faster than P and/or that detritus was
enriched with P by microflora associated with the detritus. The patterns were
opposite for detritus from Fucus and Zostera leaves and root/rhizomes – C:P-ratios
increased (albeit not significantly for Fucus) especially during the later stages of
decomposition. Thus, P was lost much faster than C from the detritus.
The decrease in C:P-ratios of detritus from Ceramium and Ulva corresponded well
to the observation that fresh detritus of these plants were relatively depleted in P.
Colonization of microbial decomposers should in theory contribute P to the mixture
of detritus and microflora. Fucus and Zostera had even higher C:P-ratios than the
two ephemeral algae, but experienced never the less a rapid loss of P during
decomposition. Low N:P-ratios (range 10-15) indicated, however, that these plants
were more severely depleted in N relative to P. P may therefore have been available
in excess from the detritus for the microbes relative to N, which was limiting
relative to P for their requirements, and therefore P may have been preferentially
excreted (i.e. mineralized).
Changes in N:P-ratios obviously reflect largely the changes in C:N and C:P-ratios,
but they do provide insights into the relative nutrient limitations for microbes
utilizing these detritus types. The N:P-ratio within detritus from Ceramium
decreased slightly due to the weak immobilization of P while that of Ulva remained
largely unaffected over time suggesting good nutrient balance for microorganisms.
The N:P-ratios of detritus from Fucus and Zostera increased, in contrast, markedly
due to the parallel immobilization of N and rapid losses of P. The immobilization of
P and N, respectively, is a good indication of which nutrient is limiting for the
decomposer microbes and therefore which nutrient is most likely to be retained
during the decomposition of that detritus.
G.T. Banta, M.F. Pedersen and S.L. Nielsen
Zostera Roots
Zostera roots
Zostera roots
Figure 8. C, N and P relationships (mean ± s.d.) during decomposition of detritus from three
species of macroalgae and the seagrass Zostera marina (both leaves and roots/rhizomes).
Details of the study are given in Figure5 (Banta, Pedersen & Nielsen, unpublished).
In overall terms, decomposition lead to a net enrichment of detritus by the most
depleted nutrient (most often N) and rapid losses of replete nutrient species (often P
in our study). Only few studies of decomposition among marine plants also include
information on stoichiometric changes during decomposition. Results from those
few studies vary in similar fashion as our results did. For example, Buchsbaum et
al. (1991) and Schmidt (1980) found that N and C was released at approximately
similar rates during decomposition of detritus from Fucus vesiculosus, Laminaria
Decomposition of marine primary producers
sp. and Zostera marina leaves under in situ conditions. In contrast, Rublee and
Roman (1982) and Kristensen (1994) found that N was released slower than C in
detritus from macroalgae (Ulva lactuca and Fucus vesiculosus) and seagrasses
(Zostera marina, Ruppia maritima and Thalassia testudinum). Finally, Paalme et al.
(2002) found that N was released more slowly and P faster than dry weight biomass
was decomposed in detritus from the two ephemeral macroalgae Cladophora
glomerata and Pilyella littoralis which is consistent with our results for Fucus and
Zostera. The most likely explanation of these varying degrees of N and P release or
immobilization is the starting nutrient status (i.e., C:N:P) of the detritus studied
relative to the needs of decomposer microorganisms. Availability of nutrients in the
environment may also play a role, but a systematic investigation of nutrient
availability and coastal macrophyte decomposition is lacking. Finally, whether
decomposition occurs aerobically or anaerobically also affects nutrient
stoichiometry because anaerobic bacteria are generally less efficient in utilizing C
and thus retain less of the nutrients remineralized per mole C during decomposition
(Blackburn, 1979; Canfield, Kristensen, & Thamdrup, in press). Results from both
our study and others do suggest that stoichiometric changes in detritus during
decomposition can vary substantially, likely governed by the availability of internal
and external nutrients. More work should be focused on this area if we are to better
understand how the decomposition of macrophytes affects the recycling of nutrients
in coastal ecosystems.
If major nutrients are immobilized in detritus by microbial decomposers in
significant amounts due to nutrient limitation this may have practical implications
for the patterns of nutrient mineralization and recycling in coastal ecosystems.
Substantial nutrient immobilization (N or P) may result in slower mineralization
rates and refractory pools of detritus that are more nutrient rich when expressed in
units of N or P instead of C. In contrast, if nutrients are available in excess of
microbial needs, this may lead to faster nutrient mineralization and recycling rates
and the formation of refractory pools of detritus depleted in nutrients relative to C in
such environments.
The overall impression from our investigation of decomposition patterns of coastal
primary producers is that the decomposition of fast growing, ephemeral algae occurs
rapidly and completely with only slight differences in the rates of release of different
components (C, N and P). The below-ground parts of seagrasses represent the
opposite extreme where decomposition occurs very slowly and is less complete.
Nutrients (especially N) are released much more slowly from seagrass roots and
rhizomes when compared to C, indicating that limiting nutrients are immobilized
during the decomposition of this poor quality detritus. Slower growing, perennial
macroalgae and seagrass leaves represent an intermediate between these two
extremes of decomposition with faster and more complete decomposition compared
to seagrass roots/rhizomes, but slower than for the fast-growing ephemeral algae.
This suggests that slow-growing macroalgae may play an intermediate role in
nutrient retention, slowing the recycling of nutrients via decomposition compared to
the case of fast growing, ephemeral macroalgae.
G.T. Banta, M.F. Pedersen and S.L. Nielsen
The role of microalgae in terms of nutrient retention is less clear as it has not been
studied as extensively or in comparable fashion to macrophytes. Decomposition of
microalgae occurs rapidly, but the presence of a refractory component could lead to
nutrient retention in preserved organic matter deposits. More attention should be
focused on refractory components of microalgae detritus and its potential role in
nutrient retention. Also, much more work should be done on the decomposition of
benthic microalgae. This important group of coastal primary producers has been
overlooked almost entirely, probably due in part to the difficulties in studying the
decomposition of microalgae so closely associated with sediment particles.
Given the extensive background information on decomposition and nutrient
mineralization dynamics for coastal primary producers presented so far, we can now
consider the consequences of changes in coastal “plant” communities due to, e.g.,
eutrophication, for nutrient recycling and retention via detritus and decomposition.
As reviewed in Chapter 3, one of the main consequences of eutrophication is to
cause a shift in primary producer dominance from slow-growing seagrasses and
brown macroalgae to fast-growing, ephemeral macroaglae (greens and reds) and
phytoplankton in coastal ecosystems. Based on the review in this chapter, such
changes in coastal plant communities towards the latter groups would be expected to
lead to faster and more complete nutrient recycling, i.e., less nutrient retention.
Systems dominated by seagrasses, on the other hand, will both recycle nutrients
more slowly and permanently retain a larger proportion of nutrients in those
systems. Based on the expected changes in nutrient cycling, community changes of
that type could have a reinforcing effect on the effects of eutrophication by ensuring
that the excess nutrients entering eutrophied coastal ecosystems are recycled more
rapidly through those systems and remain available to support the greater nutrient
demands of those fast-growing primary producers (i.e., ephemeral macroalgae and
It is worth examining this general conclusion in a bit more detail in terms of the
decomposition dynamics examined in this chapter, however. While it is true that
there are significantly different rates of decomposition and nutrient mineralization
among the different coastal primary producers, this, in and of itself, may have little
effect on permanent nutrient retention and thus little effect on, for example,
eutrophication. If detritus is completely mineralized, then all nutrients will
ultimately be returned to the ecosystem via decomposition, albeit over different time
scales. (We ignore here other ecosystem processes which remover nutrients from
active cycling such as denitrification – see Chapter 10). Thus, the differences in
rates of decomposition and nutrient mineralization (i.e., k’s) observed in this chapter
are not expected to have as important ecological implications over the long term as
might first be expected as k only determines how quickly detritus is decomposed
and nutrient are mineralized, not whether they are retained.
What is of more importance for nutrient retention is the presence of refractory pools
of nutrients for different types of detritus. Such pools are especially prevalent in
seagrasses, particularly in their underground components which make up a
significant fraction of their biomass. The other macrophytes considered had only
small refractory pools, although there was a great deal of variation. Phytoplankton
appear to have a significant refractory pool based on the data available, but it is
unclear what the ultimate fate of nutrients retained in refractory phytoplankton
detritus is in the environment. Based on these observations, we suggest that the
Decomposition of marine primary producers
most important alteration on nutrient retention due to plant uptake and sequestering
in biomass occurs when costal communities go from being dominated by rooted
macrophytes (seagrasses) to other primary producers. This shift in community
composition is one of the first effects of eutrophication (see Chapter 3) which may,
in turn, make it difficult to reverse the effects of eutrophication later given the
changes in nutrient cycling dynamics of the different primary producer
communities. Unfortunately, few studies have examined the role of coastal “plant”
community changes on nutrient dynamics from this perspective. It is an area worth
further consideration in the future.
The foregoing comparison shows that patterns of decomposition and mineralization
differ substantially among different types of marine primary producers.
Furthermore, we and others have inferred that such differences may have significant
consequences for the turn-over, accumulation and, eventually, burial of detritus and
associated nutrients such that more detritus and nutrients accumulate and become
buried in systems dominated by slow-growing, perennial macrophytes (e.g.,
Cebrián, 1999; Duarte, 1995; Enríquez et al., 1993). Whether or not dominance of
slow-growing, perennial macrophytes does in fact lead to a higher accumulation of
detritus and retention of associated nutrients remains, however, to be systematically
The best way to evaluate this problem is to compare the detritus dynamics in
systems dominated by different types of primary producers. We conducted a detailed
survey based upon repeated sampling in two shallow estuaries in Isefjorden
(Denmark) which provided data on seasonal changes in plant community
composition, biomass, primary production, grazing losses and tissue nutrient
concentrations. The daily production of detritus from each of the major plant types
was estimated from measured production, grazing losses and changes in biomass
between successive sampling events. Export losses were not accounted for in this
study and the estimated production of detritus may therefore have been slightly
over-estimated (see Chapter 4). Dynamic modeling was subsequently used to
simulate the dynamics of plant-derived detritus and associated nitrogen over a year
in both systems. Mineralization of nitrogen was simulated by use of the equations
that describe decomposition of different types of plant litter expressed in units of
nitrogen. Decomposition of microalgae was simulated on the basis of the
decomposition patterns of Skeletonema costatum presented by Garber while
decomposition of the remaining macrophytes was based on the parameters presented
in Figures 5 & 6 and Table 2.
The two estuaries, Vellerup Vig and Tempelkrogen, are small and shallow and they
receive approximately the same amounts of nitrogen from external sources on an
annual basis (Table 3). The plant communities of the two estuaries differed,
however, considerably from each other. Total plant biomass and total primary
production were greater in Vellerup Vig than in Tempelkrogen. Furthermore, the
plant biomass in Vellerup Vig was dominated by perennial macrophytes (mainly
seagrasses) whereas the plant biomass in Tempelkrogen was completely dominated
G.T. Banta, M.F. Pedersen and S.L. Nielsen
by microalgae (phytoplankton and benthic microalgae) and ephemeral macroalgae
belonging to the genera Ulva, Enteromorpha, Ceramium, Polysiphonia and Pilayella
(Fig. 9, top).
Table 3. Approximate area, depth, annual nitrogen loading, total plant biomass and annual
primary production in Vellerup Vig and Tempelkrogen, Isefjorden, Denmark.
Vellerup Vig
Area (m )
Average depth (m)
External N-loading (g N m year )
Plant biomass (g C m )
Total net primary
(g C m-2 year-1)
2,5 (max: 6 m)
1,5 (max: 4 m)
Total annual production of autochthonous detritus reached ca. 710 g C m-2 year-1 in
Vellerup Vig and almost 480 g C m-2 year-1 in Tempelkrogen (Fig. 9, middle),
corresponding to ca. 79 and 52 g detritus-bound N m-2 year-1 (Table 4). The
production of detritus-bound nitrogen from slow-growing, perennial macrophytes
was relatively small in both systems, even in Vellerup Vig where perennial
macrophytes dominated plant biomass. This is partly because the production by
these plant types contribute relatively little to total primary production (Fig. 9,
bottom) but also because the detritus derived from these plants contained much less
nitrogen than detritus from microalgae and ephemeral macroalgae.
Seasonal variations in the standing stocks of detritus-bound nitrogen reflect the
balance between inputs (i.e. production of detritus-bound nitrogen) and
remineralization through decomposition. The amount of detritus-bound N in
Vellerup Vig (Fig. 10) varied substantially across the season, ranging from a few
grams of N m-2 in winter to more than 20 g N m-2 in April-May. The spring peak in
detritus-bound nitrogen resulted from a substantial production of detritus, which
appeared at the termination of a macroalgal bloom by the end of April. The detritus
from these plants decomposed rapidly, however, and mineralization rates exceeded
those of production during summer and fall, leading to a rapid decline in the
standing stocks of detritus-bound nitrogen. The input of detritus-bound nitrogen
from slow-growing, perennial macrophytes remained low relative to that from
microalgae and ephemeral macroalgae, but this detritus decomposed slowly and
incompletely, leading to a continuous accumulation of refractory detritus over time.
The amount of nitrogen contained in detritus from these plants represented therefore
more than 70% of the detritus-bound nitrogen by the end of the year (3.7 g N m-2,
Table 4).
The clear dominance of detritus from plant types that decompose rapidly and nearly
completely resulted in low standing stocks of detritus-bound nitrogen by the end of
Decomposition of marine primary producers
the year. The amount of detritus-bound nitrogen in Vellerup Vig was therefore only
5 g N m-2 by the end of December, off which 62% was contained in the refractory
component of the detritus.
Vellerup Vig
Total biomass: 17,7 g C m-2
Total biomass: 33,1 g C m-2
M icroalgae
Total net production: 956 g C m-2 year-1
M icroalgae
Total net production: 684 g C m-2 year-1
Total production of detritus: 707 g C m-2 year-1
Total production of detritus: 483 g C m-2 year-1
Figure 9. The contribution of different plant types (microalgae, ephemeral macroalgae and
perennial macrophytes) to total plant biomass, net primary production and production of
detritus in Vellerup Vig (left column) and Tempelkrogen (right column), Isefjorden, Denmark
(Pedersen, Banta and Nielsen, unpublished).
G.T. Banta, M.F. Pedersen and S.L. Nielsen
Figure 10. Seasonal changes in the standing stock of detritus-bound nitrogen in Vellerup Vig
(left) and Tempelkrogen (right), Isefjorden, Denmark. Temporal changes reflect the balance
between production of detritus-bound nitrogen (from different plant types) and losses of
nitrogen through mineralization. Data from Pedersen, Banta and Nielsen (unpublished)
The annual production of detritus-bound nitrogen in Tempelkrogen reached 52 g N
m-2 and microalgae and ephemeral macroalgae provided by far most of that detritus
(Table 4). Decomposition and mineralization was fast and the standing stocks of
detritus-bound nitrogen remained always smaller and less seasonal variable than in
Vellerup Vig. Maximum levels of detritus-bound nitrogen reached only 6-7 g m-2 in
mid-summer and declined to 1.5 g N m-2 by the end of the year, corresponding to
less than 3% of the amount produced annually (Fig. 10). The amount of nitrogen
bound in refractory detritus from slow-growing perennial macrophytes by the end of
December was only 0.4 g N m-2, or about 40% of the total standing stock (Table 4).
The amount of detritus-bound nitrogen that accumulates within these systems is thus
markedly higher when the plant community is dominated by slow-growing,
perennial macrophytes rather than by fast-growing, ephemeral algae (including
microalgae). Part of this difference can be attributed to differences in the amount of
detritus produced, but most is due to the slow and incomplete turn-over of detritus
originating from perennial plants. In the present case, for example, the production of
detritus-bound N in Vellerup Vig exceeded that of Tempelkrogen by 50%, while the
standing-stock of detritus-bound N in Vellerup Vig was 3.6 times higher than that in
Tempelkrogen after one year. Dominance of slow-growing, perennial macrophytes
did not only stimulate a greater accumulation of total and “labile” detritus (1.9 g N
m-2 in Vellerup Vig versus 1.1 m-2 in Tempelkrogen), but also to a much more
dramatic increase in the amount of nitrogen bound in refractory detritus (3.3 versus
0.4 g N m-2). The latter is important in relation to retention of nutrients at the
ecosystem scale since refractory detritus and associated nutrients may be especially
prone to permanent burial.
Decomposition of marine primary producers
Table 4. Summary of the dynamics of detritus-bound nitrogen in Vellerup Vig and
Tempelkrogen, Isefjorden, Denmark (Pedersen, Banta and Nielsen, unpublished).
Vellerup Vig
Production of detritus-bound N (g N m-2 year-1):
Ephemeral macroalgae
Perennial macrophytes
Standing stock of detritus-bound N after one year (g N m-2):
Ephemeral macroalgae
Perennial macrophytes
Labile detritus (g N m-2)
Refractory detritus (g N m-2)
It is finally worth noting that most of the nutrients that were assimilated by plants
and became tied up in detritus were in fact recycled rapidly and almost completely
through decomposition over rather short time-scales. Thus, 93% and 97% of the
detritus-bound nitrogen that was “produced” in Vellerup Vig and Tempelkrogen,
respectively, was mineralized within the same year as it was produced.
Consequently, less than 7% and 3%, respectively, of the nitrogen that was originally
taken up by primary producers and incorporated in “plant” detritus could ever be
subject to permanent burial, suggesting that marine plants may play a relatively
minor role for the burial and the permanent retention of nutrients in coastal marine
environments compared to nutrient loadings.
Burial of nitrogen in estuarine and coastal sediments average about 5 g N m-2 year-1
(range: 0.2 – 14.3 g N m-2 year-1)(Nixon, Granger, & Nowicki, 1996). We do not
know the rates of permanent N-burial in Vellerup Vig and Tempelkrogen but the
estimated amount of nitrogen accumulated in refractory detritus was 3.1 g N m-2
year-1 in Vellerup Vig and 0.4 g N m-2 year-1 in Tempelkrogen. Long term burial of
plant derived detritus along with associated nutrients could thus contribute
significantly to total burial of nitrogen in Vellerup Vig but not in Tempelkrogen,
assuming that total rates of burial are similar to those found elsewhere. The key to
this difference is the difference in plant community composition, namely the
presence of significant perennial macrophytes, especially seagrasses. Thus, while
the absolute effects of permanent nutrient retention via plant detritus were relatively
small in terms of nutrient loadings in both systems, the only system where it
occurred at all was Vellerup Vig where seagrasses were prevalent. This is consistent
with the conclusions we have drawn earlier in this chapter.
G.T. Banta, M.F. Pedersen and S.L. Nielsen
In this chapter we have reviewed decomposition and nutrient mineralization patterns
for coastal primary producers with the goal of shedding light on the role of different
plant groups for retaining assimilated nutrients in their detritus. While it is clear that
there is a great deal of variation, both within and between species, there are general
differences in decomposition patterns that have implications for nutrient recycling
and retention. The production of refractory pools of nutrients, which particularly
occurs with seagrass detritus, is especially relevant for the long-term retention of
nutrients. Other plant groups (micro- macroalgae) are not expected to retain
nutrients in detrital pools on a permanent basis. The extent to which particular
nutrients are retained depends on a variety of factors including the availability or
limitation of nutrients for the decomposer microorganisms in the environment as this
affects the stoichiometry of nutrient release during decomposition. As such, this
could be expected to vary greatly from system to system. More attention should be
given to nutrient stoichiometry during detrital mineralization and recycling.
Finally, as was evident form the case study examined, while nutrient uptake by
coastal primary producers is very important for coastal nutrient dynamics, most of
those nutrients are in fact recycled to the environment again via decomposition.
Still, a fraction of those nutrients can be permanently retained in refractory detrital
pools provided they are taken up by slow-growing perennial macrophytes
(especially seagrasses). Nutrients assimilated by faster-growing primary producers
will be efficiently recycled within the ecosystem via decomposition with little loss
or retention. Thus changes in coastal plant communities will have some impact on
coastal nutrient cycling dynamics, the importance of which will depend on the
degree of external loadings in the environment of interest.
Aber, J. D., & Melillo, J. M. (2001). Terrestrial Ecosystems (2nd ed.). San Diego, USA: Academic Press.
Albright, L. J., Chocair, J., Masuda, K., & Valdés, M. (1980). In situ degradation of the kekps
Macrocystis integrifolia and Nerocystis luetkeana in British Columbia coastal waters. Le Naturaliste
Canadien, 107(1), 3-10.
Andersen, F. O. (1996). Fate of organic carbon added as diatom cells to oxic and anoxic marine sediment
microcosms. Marine Ecology Progress Series, 134(1-3), 225-233.
Andersen, F. O., & Kristensen, E. (1992). The importance of benthic macrofauna in decomposition of
microalgae in a coastal marine sediment. Limnology and Oceanography, 37, 1392-1403.
Birch, P. B., Gabrielson, J. O., & Hamel, K. S. (1983). Decomposition of Cladophora .1. Field Studies in
the Peel-Harvey Estuarine System, Western Australia. Botanica Marina, 26(4), 165-171.
Blackburn, T. H. (1979). Nitrogen/carbon ratios and the rates of ammonia turnover in anoxic sediments.
Paper presented at the Microbial Degradation of Pollutants in Marine Environments, Gulf Breeze,
Borum, J., & Sand-Jensen, K. (1996). Is total primary production in shallow coastal marine waters
stimulated by nitrogen loading? Oikos, 76(2), 406-410.
Bourgues, S., Auby, I., deWit, R., & Labourg, P. J. (1996). Differential anaerobic decomposition of
seagrass (Zostera noltii) and macroalgal (Monostroma obscurum) biomass from Arcachon Bay
(France). Hydrobiologia, 329(1-3), 121-131.
Brouwer, P. E. M. (1996). Decomposition in situ of the sublittoral Antarctic macroalga Desmarestia
anceps Montagne. Polar Biology, 16(2), 129-137.
Buchsbaum, R., Valiela, I., Swain, T., Dzierzeski, M., & Allen, S. (1991). Available and refractory
nitrogen in detritus of coastal vascular plants and macroalgae. Marine Ecology Progress Series, 72(12), 131-143.
Decomposition of marine primary producers
Canfield, D. E., Kristensen, E., & Thamdrup, B. (in press). Aquatic Geomicrobiology.
Cebrián, J. (1999). Patterns in the fate of production in plant communities. American Naturalist, 154(4),
Christensen, P. B., Møhlenberg, F., Krause-Jensen, D., Jensen, H. S., Rysgaard, S., Clausen, P., et al.
(1994). Stoftransport og stofomsætning i Kertinge Nor/Kertiminde Fjord (No. 43). Copenhagen,
Denmark: Miljøstyrelsen.
Duarte, C. M. (1992). Nutrient Concentration of Aquatic Plants - Patterns across Species. Limnology and
Oceanography, 37(4), 882.
Duarte, C. M. (1995). Submerged Aquatic Vegetation in Relation to Different Nutrient Regimes. Ophelia,
41, 87.
Duarte, C. M., & Cebrián, J. (1996). The fate of marine autotrophic production. Limnology and
Oceanography, 41(8), 1758-1766.
Duarte, C. M., Merino, M., Agawin, N. S. R., Uri, J., Fortes, M. D., Gallegos, M. E., et al. (1998). Root
production and belowground seagrass biomass. Marine Ecology Progress Series, 171, 97-108.
Enríquez, S., Duarte, C. M., & Sand-Jensen, K. (1993). Patterns in decomposition rates among
photosynthetic organisms: the importance of detrital C:N:P content. Oecologia., 94, 457-471.
Gabrielson, J. O., Birch, P. B., & Hamel, K. S. (1983). Decomposition of Cladophora .2. Invitro Studies
of Nitrogen and Phosphorus Regeneration. Botanica Marina, 26(4), 173-179.
Gallagher, J. L., Kibby, H. V., & Skirvin, K. W. (1984). Detritus Processing and Mineral Cycling in
Seagrass (Zostera) Litter in an Oregon Salt-Marsh. Aquatic Botany, 20(1-2), 97-108.
Garber, J. H. (1984). Laboratory study of nitrogen and phosphorus remineralization during the
decomposition of coastal plankton and seston. Estuarine and Coastal Shelf Science, 18, 685-702.
Godshalk, G. L., & Wetzel, R. G. (1978). Decomposition of Aquatic Angiosperms .3. Zostera-Marina L
and a Conceptual-Model of Decomposition. Aquatic Botany, 5(4), 329-354.
Goldman, J. C., Caron, D. A., & Dennett, M. R. (1987). Regulation of Gross Growth Efficiency and
Ammonium Regeneration in Bacteria by Substrate C-N Ratio. Limnology and Oceanography, 32(6),
Hemminga, M. A., & Nieuwenhuize, J. (1991). Transport, Deposition and Insitu Decay of Seagrasses in a
Tropical Mudflat Area (Banc Darguin, Mauritania). Netherlands Journal of Sea Research, 27(2),
Holmer, M., & Olsen, A. B. (2002). Role of decomposition of mangrove and seagrass detritus in sediment
carbon and nitrogen cycling in a tropical mangrove forest. Marine Ecology Progress Series, 230, 87101.
Hunter, R. D. (1976). Changes in Carbon and Nitrogen-Content During Decomposition of 3 Macrophytes
in Freshwater and Marine Environments. Hydrobiologia, 51(2), 119-128.
Jewell, W. J., & McCarty, P. L. (1971). Aerobic decomposition of algae. Environmental Science and
Technology, 5, 1023-1031.
Josselyn, M. N., & Mathieson, A. C. (1980). Seasonal Influx and Decomposition of Autochthonous
Macrophyte Litter in a North Temperature Estuary. Hydrobiologia, 71(3), 197-207.
Kenworthy, W. J., & Thayer, G. W. (1984). Production and Decomposition of the Roots and Rhizomes of
Seagrasses, Zostera-Marina and Thalassia-Testudinum, in Temperate and Sub-Tropical Marine
Ecosystems. Bulletin of Marine Science, 35(3), 364-379.
Kristensen, E. (1994). Decomposition of Macroalgae, Vascular Plants and Sediment Detritus in Seawater
- Use of Stepwise Thermogravimetry. Biogeochemistry, 26(1), 1-24.
Kristensen, E., Andersen, F. O., & Blackburn, T. H. (1992). Effects of benthic macrofauna and
temperature on degradation of macroalgal detritus: The fate of organic carbon. Limnology and
Oceanography, 37, 1404-1419.
Lee, B.-G., & Fisher, N. S. (1992). Degradation and elemental release rates from phytoplankton debris
and their geochemical implications. Limnology and Oceanography, 37, 1345-1360.
Mateo, M. A., & Romero, J. (1997). Detritus dynamics in the seagrass Posidonia oceanica: Elements for
an ecosystem carbon and nutrient budget. Marine Ecology Progress Series, 151(1-3), 43-53.
Melillo, J. M., Aber, J. D., & Muratore, J. F. (1982). Nitrogen and Lignin Control of Hardwood Leaf
Litter Decomposition Dynamics. Ecology, 63(3), 621-626.
Newell, R. C., Lucas, M. I., & Linley, E. A. S. (1981). Rate of degradation and efficiency of conversion
of phytoplankton debris by marine micro-organisms. Marine Ecology Progress Series, 6, 123-136.
Newell, S. Y., Fell, J. W., Statzelltallman, A., Miller, C., & Cefalu, R. (1984). Carbon and Nitrogen
Dynamics in Decomposing Leaves of 3 Coastal Marine Vascular Plants of the Subtropics. Aquatic
Botany, 19(1-2), 183-192.
G.T. Banta, M.F. Pedersen and S.L. Nielsen
Nixon, S. W., Granger, S. L., & Nowicki, B. L. (1996). An assessment of the annual mass balance of
carbon, nitrogen, and phosphorus in Narragansett Bay (vol 31, pg 15, 1995). Biogeochemistry, 33(3),
Olson, J. S. (1963). Energy-Storage and Balance of Producers and Decomposers in Ecological-Systems.
Ecology, 44(2), 322-&.
Otsuki, A., & Hanya, T. (1972a). Production of Dissolved Organic-Matter from Dead Green Algal Cells
.1. Aerobic Microbial Decomposition. Limnology and Oceanography, 17(2), 248-&.
Otsuki, A., & Hanya, T. (1972b). Production of Dissolved Organic-Matter from Dead Green Algal Cells
.2. Anaerobic Microbial Decomposition. Limnology and Oceanography, 17(2), 258-&.
Peduzzi, P., & Herndl, G. J. (1991). Decomposition and Significance of Seagrass Leaf Litter
(Cymodocea-Nodosa) for the Microbial Food Web in Coastal Waters (Gulf of Trieste, Northern
Adriatic Sea). Marine Ecology Progress Series, 71(2), 163-174.
Pellikaan, C. G. (1982). Decomposition processes of eelgrass, Zostera marina L. Hydrobiological
Bulletin, 16(1), 83-92.
Paalme, T., Kukk, H., Kotta, J., & Orav, H. (2002). 'In vitro' and 'in situ' decomposition of nuisance
macroalgae Cladophora glomerata and Pilayella littoralis. Hydrobiologia, 475(1), 469-476.
Rice, D. L., & Tenore, K. R. (1981). Dynamics of Carbon and Nitrogen During the Decomposition of
Detritus Derived from Estuarine Macrophytes. Estuarine Coastal and Shelf Science, 13(6), 681-690.
Rublee, P. A., & Roman, M. R. (1982). Decomposition of Turtlegrass (Thalassia-Testudinum Konig) in
Flowing Sea-Water Tanks and Litterbags - Compositional Changes and Comparison with Natural
Particulate Matter. Journal of Experimental Marine Biology and Ecology, 58(1), 47-58.
Schmidt, C. (1980). Some aspects of marine algae decomposition. Ophelia., Suppl.1, 257-264.
Staaf, H., & Berg, B. (1982). Accumulation and Release of Plant Nutrients in Decomposing Scots Pine
Needle Litter - Long-Term Decomposition in a Scots Pine Forest .2. Canadian Journal of BotanyRevue Canadienne De Botanique, 60(8), 1561-1568.
Swift, M. J., Heal, O. W., & Anderson, J. M. (1979). Decomposition in terrestrial systems (Vol. 5).
Oxford: Blackwell.
Twilley, R. R., Ejdung, G., Romare, P., & Kemp, W. M. (1986). A Comparative-Study of Decomposition,
Oxygen-Consumption and Nutrient Release for Selected Aquatic Plants Occurring in an Estuarine
Environment. Oikos, 47(2), 190-198.
Valiela, I., Teal, J. M., Allen, S. D., Van Etten, R., Goehringer, D., & Volkmann, S. (1985).
Decomposition in salt marsh ecosystems: the phases and major factors affecting disappearance of
above-ground organic matter. Journal of Experimental Marine Biology and Ecology, 89, 29-54.
Wahbeh, M. I., & Mahasneh, A. M. (1985). Some Aspects of Decomposition of Leaf Litter of the
Seagrass Halophila-Stipulacea from the Gulf of Aqaba (Jordan). Aquatic Botany, 21(3), 237-244.
Westrich, J. T., & Berner, R. A. (1984). The role of sedimentary organic matter in bacterial sulfate
reduction: the G model tested. Limnology and Oceanography, 29, 236-249.
G.T. Banta, M.F. Pedersen & S.L. Nielsen: Department of Life Sciences and
Chemistry, Roskilde University. P.O. Box 260, DK-4000 Roskilde, Denmark.
Anthropogenic nutrient inputs to aquatic systems exceed those under pristine
conditions. This has resulted in numerous changes in their functioning of which
eutrophication effects, the development of nuisance algal blooms, seasonal anoxia,
the disappearance of seagrasses are illustrative, well-documented examples (Heip,
1995; Duarte, 1995). Concern about these changes in ecosystem functioning has
initiated large amount of research on nutrient cycling with the ultimate aim to
predict consequences of policy measures and management options. One of the basic
tools used in the study of nutrient cycling in aquatic ecosystems is the construction
of nutrient budgets.
Nutrient budgets are very much alike financial bookkeeping systems with revenues
and assets on one side and expenses and debts on the other side. Accountancy has
developed into a mature discipline that is able to deal with internal and external
losses and revenues and with disparate, not always accurate or quantifiable data. For
instance, long-term depreciation costs have to be balanced against the expected
enhanced profit of an investment.
Biogeochemists and ecologists constructing nutrient budgets are faced with similar
complexities. On the one hand there are event-like input terms such as atmospheric
deposition of nutrients during rain that cover a period of a few hours at most and on
the other hand there are long-term (years) loss terms such as the nutrient contained
in accumulating sediments. For ecosystem budgets, it is important to distinguish
between internal sinks (primary production and nutrient assimilation) and sources
(remineralisation) and external (river) input and output terms (denitrification and
In this chapter we will address the burial term for nutrients in coastal sediment
ecosystems. We will first provide a definition and then discuss the two main
components (sediment accumulation and nutrient concentration). Special attention
will be given to the influence of primary producers on nutrient burial via
enhancement of sediment accumulation rates or via their effect on organic matter
production or preservation.
The term burial is used within various subdisciplines with the consequences that it is
used for different processes that operate on different time scales. Here we define
burial as the longer-term removal (annual to decadal scale) of nutrients from the
S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling:
The IInfluence of Primary Producers, 217-230.
© 2004 Kluwer Academic Publishers. Printed in the Netherlands.
J.J. Middelburg et al.
pelagic system, by accumulation of sediments. Geoscientists working on carbon and
nutrient recycling often use the term burial efficiency, which expresses the fraction
of carbon or nutrient delivered to the sediment that becomes buried.
Soil scientists and some plant ecologists use the term burial to distinguish between
aerial, exposed degradation and decomposition within the soil or sediments, where
biogeochemical conditions are different (e.g. McInerney & Bolger, 2000; Rovira &
Ramón Vallejo, 2002). Burial is then assessed for instance by transferring leaf
material within litterbags into a deeper layer to study decomposition of organic
matter under anoxic conditions.
Burial should also not be confused with storage or refractory accumulation within
ecosystems sensu Duarte & Cebrian (1996) and Cebrian (2002). Detritus that is not
decomposed within the study period (typically a year) accumulates temporarily on
the sediment surface. However, before it is removed for a longer period of time, it is
subject to further degradation within the sediment with losses of 50 to > 90 % (Heip
et al., 1995).
Our definition of burial derives from the geosciences (Berner, 1980; Boudreau,
1997). Sediments comprise water and particles, both of which may contain nutrients
and which interact. Mineralisation and dissolution will provide a transition from
solid to liquid phase, whereas primary production, adsorption and precipitation work
the other way around. Two different types of transport affect the distribution of these
constituents in the sediment. Accretion of sediment and compaction will remove
both liquid and solid substances away from the sediment surface. Molecular
diffusion and bioirrigation will further affect dissolved constituents, whilst sediment
reworking by animal feeding and movement (bioturbation) or by the action of waves
and tides affects particulates. A thorough discussion of these various modes can be
found in Boudreau (1997). For our current interest, it suffices to note that
(bio)turbation generally operates much slower than molecular diffusion and is
limited to the upper layer of the sediment only. Consequently, burial of the solid
phase is much more effective than burial of the liquid phase and the latter is usually
ignored. Once particulate matter is transferred to below the sediment layer, which is
biogeochemically and ecologically active, it is isolated from the short-term
biologically controlled cycles and may either become involved in a rapid geological
cycle (decades to thousands of years) or even in the long-term geological cycle
(million of years). Rapid geological recycling can be related to migrating gullies
that erode at one place and cause deposition at another location, or to changes in the
relative rise of sea level that can cause drowning or emergence of depositional areas.
Mathematically speaking, burial of particulate matter is the product of the solid
fraction (1-porosity), concentration of the constituents (Ci mass/mass), the dry
density of particles (ρ; mass volume-1) and sediment accretion rate (ω, length yr-1 ).
Burial = (1 − φ )Ci ρω
where φ is the sediment porosity (volume of water/volume of total sediment).
It is important to realize that the accumulation rate (ω) in eq. 1 refers to long-term
sediment accretion at sufficient depth below the sediment surface and not to the
Burial of nutrients in coastal sediments
deposition of material at the sediment surface. Particle deposition rates at the
sediment surface can be much higher than net accumulation rates because many
particles that are deposited on the sediment surface are resuspended. These
deposition-resuspension cycles may occur at hourly-daily time scales in tidal
systems or at seasonal time scales in many coastal, temperate sediments (Herman et
al., 2001). This difference between temporary deposition and long-term
accumulation should be kept in mind when constructing nutrient budgets. Dynamic
nutrient budgets may include this temporal repository of sedimentary nutrient
accumulation, but long-term, annual-scale nutrient budgets should include the burial
term as defined in this chapter.
Four terms define the removal of nutrients by burial (eq.1): porosity (φ), the
dry density of particles (ρ), the deep concentration (Ci) and the sediment
accumulation rate (ω). The first two terms are rather invariant and consequently do
not contribute significantly to the variance in nutrient burial estimates. Porosity, the
fraction of water, varies from 0.4 for coarse sands to 0.8 for mud (Boudreau, 1997).
Similarly, the dry density of inorganic sedimentary particulates is usually in the
range of 2.4 to 2.6 g cm-3, except for peat with an organic matter content of >45 wt%
which may have bulk dry densities less than 2.0 g cm-3. The variability in porosity
and dry densities can be neglected relative to those in nutrient concentration and the
sediment accumulation rate. Sediment accumulation rate varies over orders of
magnitude in coastal sediments and is the prime factor governing burial of carbon
and nutrients (Middelburg et al., 1997a). It should be remarked that there is not
necessarily a relationship between sediment accumulation rates and nutrient
concentration. The mere enhancement of mineral deposition without associated
organic matter and nutrients could in principle lead to a lowering of nutrient
concentration by dilution but usually there is a coupling between particle and
(organic) nutrient deposition and enhanced deposition rates will therefore not result
in lower nutrient concentrations.
Sediment accumulation refers to the permanent transfer of material from the
seasonally influenced, biogeochemically and ecologically active layer to the deeper
Quantification of sediment accumulation rates in coastal ecosystems is usually based
on the radionuclides such as the primordial ones (210Pb), atmospheric ones (7Be, 14C)
or artificial, human-produced radionuclides such as 137Cs (Thomsen et al., 2002).
Determination of sediment accumulation rates based on artificial radionuclides boils
down to determining the sediment depth at which the event (e.g. Chernobyl May
1986) or maximum release (e.g. 1963 atmospheric testing of nuclear bombs treaty)
can be recognized and division by time since that period (Figure 1). Natural and
primordial radionuclides normally show an exponentially decaying profile with
depth in the sediment due to input from the water column and subsequent
radioactive decay. Fitting of an exponential regression allows then calculation of the
sediment accumulation (Figure 1). Care should be taken not to apply the procedure
to the surface layer which is mixed by wave or current activities
J.J. Middelburg et al.
Figure 1. Sediment accumulation rate determinations based on the radionuclides 137Cs and
Pb. Activity versus depth profiles of 137Cs usually show a subsurface maximum related to
maximum above-ground nuclear bombing testing in 1963. Dividing the depth of the 137Cs
peak with the time period between sampling and 1963 provides an estimate for sediment
accumulation (i.e. 20/(2003-1963) = 0.5 cm yr-1). Activity versus depth profiles of 210Pb
typically show an exponential decline (straight line on semi-logarithmic graph) with the depth
attenuation coefficient providing an estimate of mixing or sediment accumulation rate. The
line shown is consistent either with a sediment accumulation rate of 0.5 cm yr-1, a
bioturbation coefficient of 8 cm2 yr-1 or a sediment accumulation rate of 0.4 cm yr-1 and a
mixing rate of 1.5 cm2 yr-1. Sediment accumulation rate determinations should therefore
always be based on more than one natural radionuclide if sediment mixing cannot be
excluded a priori, as for instance in deep, anoxic basins.
Burial of nutrients in coastal sediments
and animals moving and feeding as here the rate of sediment mixing and
accumulation cannot be independently derived on the basis of one radionuclide. One
either has to apply additional constraints or use a second radionuclide having a
different depth profile, i.e. different decay rate. Similarly, with the interest in longterm accumulation, one should use radionuclides that have comparable long halflives. Given the short half lives of 234Th (24 days) and 7Be (53 days), these
radionuclides can only be used for dating in rapidly accumulating sediments subject
to limited mixing. However, a detailed study of their inventories (activity per unit
area) may provide useful information on seasonal accumulation of material (Canuel
et al., 1990). Sediment accumulation rate may vary anywhere between zero or even
negative (erosion) to more than 10 cm yr-1 (Heip et al., 1995; Figure 2). This
variability is primarily due to variability in sediment supply and availability, and
differences in hydrodynamic conditions.
Primary producers may affect sediment accumulation rates in a number of ways that
differ between macrophytes (submerged and emergent) and benthic microalgae.
Several recent studies have shown that sediment stability; hence net sediment
accumulation can be enhanced by biofilms and microbial mats (Paterson and Black,
1999). Stabilisation of sediments involves the production of extracellular
polysaccharides by benthic organisms, in particular benthic diatoms (Holland et al.,
1974; de Brouwer et al., 2000). The mud-diatom system may be of two types (Van
de Koppel et al., 2001): one with low microalgal biomass and low mud content and
one with high algal biomass and high mud content. Since nutrient contents and
availability are related directly to mud content, it is likely that sedimentary
microalgae more or less modify their environment so as to optimise nutrient
availability (ecosystem engineering, Jones et al., 1997). Parallel to the seasonality in
benthic microalgal production, the enhanced stabilisation of sediments due to
microalgae is also a seasonal phenomena and this is likely the reason for the
seasonal deposition of silty and muddy sediments on tidal flats (Fig. 3). The twoway interactions between microalgae and their hosting sediments are further
complicated through intensive grazing and sediment mixing by the infauna
(Middelburg et al., 2000; Herman et al., 2000).
Canopies of macrophyte communities are known to promote deposition of particles
and to reduce resuspension and erosion with the net result that sediment
accumulation is enhanced (Scoffin, 1970; Redfield, 1972). This effect has been
reported for mangroves, salt marshes, fresh-water marshes and seagrass beds and
has been studied in detail for seagrasses (Zostera marina and Posidonia oceanica)
and saltmarsh plants such as Spartina anglica. The presence of shoots influences the
hydrodynamics of the systems: they reduce water current velocities through
extraction of momentum and increase the thickness of the benthic boundary layer
(Paterson and Black, 1999; Jumars et al., 2001). The effects of macrophytes on
water flow and sediment deposition are reciprocal. For instance, water flow affects
the performance of the seagrasses and vice versa (Fonseca et al., 1982, 1983).
Moreover, canopies of coastal plants may also reduce wave energy (Fonseca and
Calahan, 1992) and affect water turbulence, which either may increase or decrease
depending on external conditions and canopy architecture (Jumars et al., 2001). The
consequences of this on particle deposition and resuspension are not yet fully
understood. Recently, Gacia and Duarte (2001) have shown that canopy-induced
J.J. Middelburg et al.
reduction in resuspension rates is more important than canopy-induced enhancement
of particle deposition. Canopies may also provide a habitat for suspension feeders
that cause deposition of faeces and pseudofaeces, part of which may be incorporated
in the sediment (Heip et al., 1995)
Canopies not only increase the total rate of particle accumulation, they also alter the
composition due to preferential settling and retention of fine particles within the
canopies. This preferential accumulation results in higher nutrient concentration in
sediments colonized by macrophytes compared to unvegetated sediments (see
below; Kenworthy et al., 1982; Benoy and Kalff, 1999). Finally, the roots and
rhizomes of seagrasses, marsh plants and mangrove trees stabilize the sediments and
lower erosion rates (Redfield, 1972; Woodroffe, 1992).
Figure 2. Histogram of sediment accumulation rates in unvegetated shelf sediments (based
on data summarized in Middelburg et al. (1997a)).
Plants may affect nutrient concentrations either directly through net accumulation of
plant litter or indirectly by their impact on sediment accretion or on sediment
biogeochemical conditions.
Coastal ecosystems are among the most productive ecosystems: they account for
about 20 % of total net primary production (Duarte and Cebrian, 1996). The
majority of biomass production is degraded, grazed or exported and only a small
Burial of nutrients in coastal sediments
amount remains available for accumulation (Cebrian, 2002). This refractory
accumulation is 4 fold higher for higher plants (10-17 % of net primary production)
than for algae (0.4-6 % of gross primary production; Duarte & Cebrian, 1996). This
small fraction of higher plant material being preserved can sometimes contribute
significantly to sediment accretion and carbon and nutrient burial (Middelburg et al.,
1997b). Macrophyte production contributes especially to this sediment accretion due
to litter accumulation because their intrinsic degradation rate is lower than that of
algae, their degradation rate is more sensitive to redox conditions (see below) and
part of their production is already allocated below-ground. Below-ground material
has a higher burial potential than above-ground material because most of the latter is
degraded and exported before incorporation in the sediments. However, these
below-ground organs usually have rather low nutrient contents because of structural
materials and the storage of energy-rich (COH-rich) compounds.
As canopies promote the deposition and retention of the finer, organic-matter rich
particles (see above), one often finds higher nutrient and organic matter
concentrations in vegetated relative to non-vegetated sediments (Benoy and Kalff,
1999): e.g. in seagrasses (Kenworthy et al., 1982; Gacia et al., 2002), salt marshes
(Middelburg et al., 1997b) and mangroves (Middelburg et al., 1996).
By changing the sediment biogeochemistry, plants may also impact nutrient
concentrations. Rooted macrophytes and sedimentary algae take up dissolved
nutrients and may release chelating components to solubilise mineral associated
nutrients, in particular phosphate. The ultimate result of these processes on nutrient
burial is not clear. On the one hand plant litter enriched in nutrients is more easily
degraded than nutrient poor litter (Enriquez et al., 1993; Cebrian, 1999) and nutrient
uptake might thus lower nutrient burial. On the other hand, nutrient uptake
stimulates production and thus enhances litter production and burial. Furthermore,
plants compete with heterotrophic bacteria for dissolved nutrients (Kaye and Hart,
1997) and it depends on the relative burial efficiency of plant litter and bacterial
remains whether nutrient uptake enhances or diminishes nutrient concentrations and
Rooted macrophytes living in anoxic sediments have a major influence on
sedimentary biogeochemical conditions through release of oxygen from the roots
(Sundby et al., 1997). Emergent and submerged plants have developed strategies to
supply their roots and rhizome with sufficient oxygen for respiration while the
environment is rich in oxygen demanding components (reduced iron and
manganese, sulphide, methane, etc). Emergent wetlands may either transport oxygen
by diffusion as a gas via their aerenchym or they employ an active ventilation
system based on convective flow due to pressure differentials (Van der Nat et al.,
1998). Part of the oxygen transported to the roots is respired and part is released into
the reducing environments, where it will react with many components resulting in
complex biogeochemical, coupled reactions (e.g. coupled nitrificationdenitrification, Reddy et al., 1989; chapter 10).
The oxygen released by microalgae or by roots has an influence on the availability
and concentrations of phosphorus via the iron cycle. Oxygen release results in the
oxidation of sulphide and reduced iron with the consequences that more iron oxide
is formed than in the absence of these plants (Roden and Wetzel, 1996; Van der Nat
and Middelburg, 1998). In most sediment phosphate is tightly associated with solid
phase iron oxides. Enhanced formation of iron oxides prevents escape of phosphate
J.J. Middelburg et al.
generated by decomposing organic matter and may ultimately result in higher
phosphor concentrations (Rozan et al., 2002). Phosphor speciation, i.e. the
distribution of phosphorus over various phases, is rather complex and is usually
defined in the following major pools (Ruttenberg, 1992): adsorbed P, organic P, Feoxide associated P, autogenic apatite P (within sediment formed) and refractory,
external apatite P. Oxygen release by benthic microalgae and roots of macrophytes
may modify the form under which P becomes buried.
% mud < 63 µm
Depth (cm)
June 96
September 96
March 97
June 97
September 97
Figure 3. Seasonal deposition of mud and silt on a tidal flat due to seasonal development of
microalgae. A. Temporal evolution of % mud < 63 µm along a transect on an intertidal flat,
sampled in March, June, September and December 1995. B. Depth profile of % mud at one of
these stations, sampled at five occasions (from Herman et al., 2001, printed with permission
from Elsevier).
Burial of nutrients in coastal sediments
The most studied and most dramatic effect of plant oxygen release on burial
concerns the effect on organic matter mineralisation efficiencies. The questions
whether organic matter degradation under oxic conditions is faster and more
efficient than anaerobic mineralisation has been addressed in many studies (e.g.
Kristensen, 2000). These studies revealed that redox conditions have little influence
on the degradation rate of labile organic matter, such as algae and seston (Westrich
and Berner, 1984), but that oxic conditions stimulate degradation of more refractory
organic material, be it inherently more refractory higher plant material (Kristensen et
al., 1995) or modified and aged marine organic matter (Hulthe et al., 1998).
Particularly aromatic structures and highly polymeric compounds such as lignin are
poorly degraded under anoxic conditions (Benner et al., 1984). The nature of the
primary producer can thus have a major influence on the preservation of organic
matter, hence burial efficiency. Recently, Dauwe et al. (2001) reported that not only
the origin or quality of organic matter determines the presence of an oxygen effect
on organic matter preservation. They argue that oxygen-enhanced degradation
occurs at low mineralisation levels at which bacterial biomass production becomes
limiting. Work by Aller and co-workers (Aller, 1994; Sun et al., 2002) have shown
that degradation efficiencies of organic material depend on specific patterns of redox
fluctuations. Oxygen release by benthic microalgae and macrophyte roots is
governed by illumination. Changes in light availability thus cause short-term redox
oscillations that favour extensive degradation of organic matter, hence lowering
burial potential (Aller, 1994).
The relative importance of plant litter accumulation to sediment accretion varies
among systems and for salt-marsh systems has led to the recognition of two endmember depositional facies: the organogenic and minerogenic modes (Allen, 1995).
Organic matter accumulation determines vertical accretion in sediment-starved peaty
marshes such as those found in New England, whereas mineral matter accumulation
controls accretion in marshes in NW Europe (Middelburg et al., 1997b). King et al.
(1992) have shown that mineral deposition may stimulate plant growth and thus
stimulate organic matter accumulation indicating that some feedbacks are
While local plant production may contribute and sometimes even dominate organic
matter and nutrient accumulation and burial, this is normally not the case. This has
been shown repeatedly using stable isotopes of carbon. The carbon isotopic
composition of organic matter buried in many vegetated coastal systems differs
significantly from the isotopic composition of local plant production indicating that
a large proportion of the carbon accumulating in these systems is external, i.e.
imported from adjacent coastal systems. This will likely also apply for nutrients
given the stoichiometric coupling of nutrients with carbon. A number of researchers
(Ember et al., 1987; Middelburg et al., 1997b; Boschker et al., 1999) have shown
that the organic matter accumulating in salt-marsh sediments is more depleted in
carbon isotopes and richer in nitrogen than organic matter derived from Spartina,
the dominant plant. They attributed this to trapping of allochthonous organic matter,
J.J. Middelburg et al.
input by benthic microalgae and nitrogen assimilation by heterotrophic bacteria.
Other researchers have reported similar observations for seagrass systems (Gacia et
al., 2002; Dauby, 1989; Boschker et al., 2000) and mangrove systems (Middelburg
et al., 1996). This accumulation of non-local produced plant material can be
attributed to preferential degradation of local produced material, selective
preservation of allochthonous carbon or to much larger allochthonous than
autochthonous inputs. The first two explanations do not apply to macrophyte
systems since the algal material imported is more labile than the macrophyte
material (Boschker et al., 1999), but it may apply to systems dominated by benthic
microalgae. Herman et al. (2000) have shown that the organic matter accumulating
in tidal flat sediments is significantly depleted in 13C compared to benthic
microalgae and they attributed this to preferential use of microalgal material.
While benthic primary producers often create unfavourable conditions for
denitrifiers (see chapter 10) and thus limit the natural buffering mechanism of
coastal ecosystems to enhanced nitrogen loadings, they do enhance nutrient removal
via burial.
In this chapter we have presented and discussed the factors governing burial or longterm loss of nutrients in coastal ecosystems hosting various primary producers. The
presence or absence and the nature of the primary producer have major implications
for the burial of nutrients. Benthic microalgae stabilise sediments via exudation of
polymeric substances that glues the sediment and lowers erodability. This is a
seasonal phenomenon with the result that most material being deposited during the
growing is resuspended lateron (Fig. 3). The ultimate consequences for nutrient
removal by burial are not well known, but probably minor. Seagrasses, marsh plants
and mangroves stabilise the sediments by their roots and rhizomes and their
canopies enhance deposition of particles with the result that their sediments
accumulate faster and are usually richer in nutrients. Moreover, a significant fraction
of the organic matter produced by these plants accumulates and may ultimately
become buried. These macrophyte are therefore of major importance in regional
budgets of sediment and nutrient burial. The disappearance of seagrasses due to
enhanced nutrient inputs (Duarte, 1995) may consequently result in less nutrient
removal by burial and thus a reduction in the resilience of coastal ecosystems
towards changing nutrient inputs (Valiela and Cole, 2002).
Allen, J.R.L. (1995). Salt-marsh growth and fluctuating sea level: implications of a simulation model for
Flandrian coastal stratigraphy and peat-based sea-level curves. Sedimentary Geology 100: 21-45.
Aller R.C. (1994). Bioturbation and remineralization of sedimentary organic matter: effects of redox
oscillation. Chemical Geology 114: 331-345.
Benner R., Maccubbin A.E. and R.E. Hodson (1984). Anaerobic biodegradation of the lignin and
polysaccharide components of lignocellulose and synthetic lignin by sediment microflora. Applied
Environmental Microbiology 47: 998-1004.
Burial of nutrients in coastal sediments
Benoy G.A. and J. Kalff (1999). Sediment accumulation and Pb burdens in submerged macrophyte beds.
Limnology & Oceanography 44: 1081-1090.
Berner, R.A. (1980). Early diagenesis. A theoretical approach. Princeton University press, Oxford.
Boschker, H.T.S., J.F.C. de Brouwer and T.E. Cappenberg (1999). The contribution of macrophyte
derived organic matter in microbial biomass in salt marsh sediments: stable carbon-isotope analysis
of microbial biomarkers. Limnology & Oceanography 44: 309-319.
Boschker H.T.S., Wielemaker A., Schaub B.E.M. and Holmer M. (2000). Limited coupling of
macrophyte production and bacterial carbon cycling in the sediments of Zostera meadows. Marine
Ecology Progress Series 203: 181-189.
Boudreau B.P. (1997). Diagenetic models and their implementation. Springer-Verlag, 414pp.
Canuel E.A., C.S. Martens and L.K. Benninger (1990). Seasonal variations in Be-7 activity in the
sediments of Cape Lookout Bight, North Carolina. Geochimica Cosmochimica Acta 54: 237-245
Cebrian J. (1999). Patterns in the fate of production in plant communities. American Naturalist 154: 449468.
Cebrian J. (2002). Variability and control of carbon consumption, export and accumulation in marine
communities. Limnology & Oceanography 47: 11-22.
Dauby P. (1989). The stable isotope ratios in benthic food webs of the Gulf of Calvi, Corsica. Continental
Shelf Research 9: 181-195
Dauwe, B. J.J. Middelburg and P.M.J. Herman (2001). The effect of oxygen on the degradability of
organic matter in subidal and intertidal sediments of the North Sea area. Marine Ecology Progress
Series 215: 13-22.
de Brouwer J.F.C., Bjelic S., de Deckere E.M.G.T., Stal L.J. (2000). Interplay between biology and
sedimentology in a mudflat (Biezelingse Ham, Westerschelde, The Netherlands). Continental Shelf
Research 20: 1159-1177
Duarte C.M. (1995). Submerged aquatic vegetation in relation to different nutrient regimes. Ophelia 41:
Duarte C.M. and J. Cebrian (1996). The fate of marine autotrophic production. Limnology &
Oceanography 41: 1759-1766.
Ember, L.M., Williams, D.F. and Morris, J.T. (1987). Processes that influence carbon isotopic variations
in salt marsh sediments. Marine Ecology Progress Series 36: 33-42.
Enriquez S., C.M. Duarte and K Sand-Jensen (1993). Patterns in decomposition rates among
photosynthetic organisms: The importance of detritus C:N: P content. Oecologia 94: 457-471.
Fonseca M.S., Fisher J.S., Zieman J.C. and G.W. Thayer (1982). Influence of seagrass, Zostera marina
L., on current flow. Estuarine Coastal Shelf Science 15:351-364.
Fonseca M.S., Zieman J.C., G.W. Thayer and J.S. Fisher (1983). The role of current velocity in
structuring Eelgrass (Zostera marina L.) meadows. Estuarine Coastal Shelf Science 17:376-380.
Fonseca M.S. and J.A. Calahan (1992). A preliminary evaluation of wave attenuation by four species of
seagrass. Estuarine Coastal Shelf Science 35:565-576.
Gacia E., C.M. Duarte and J.J. Middelburg (2002). Carbon and nutrient deposition in a Mediterranean
seagrass (Posidonia oceanica) meadow. Limnology & Oceanography 47: 23-32.
J.J. Middelburg et al.
Gacia E. and C.M. Duarte (2001).Sediment retention by a Mediterranean Posidonia oceanica meadow:
the balance between deposition and resuspension. Estuarine Coastal Shelf Science 52: 505-514
Heip, C.H.R. (1995). Eutrophication and zoobenthos dynamics. Ophelia 41: 113-136
Heip C.H.R., N.K. Goosen, P.M.J. Herman, J. Kromkamp, J.J. Middelburg and K. Soetaert (1995)
Production and consumption of biological particles in temperate tidal estuaries. Oceanography
Marine Biology Annual Reviews 33: 1-150.
Herman P.M.J., J.J. Middelburg, J. Widdows, C.H. Lucas and C.H.R. Heip (2000). Stable isotopes as
trophic tracers: combining field sampling and manipulative labelling of food resources for
macrobenthos. Marine Ecology Progress Series 204:79-92
Herman, P.M.J., J.J. Middelburg and C.H.R. Heip (2001). Benthic community structure and sediment
processes on an intertidal flat: results from the ECOFLAT project. Continental Shelf Research 21:
Holland, A.F., R.G. Zingmark and J.M. Dean (1974). Quantitative evidence concerning the stabilization
of sediments by marine benthic diatoms. Marine Biology 27: 191-196.
Hulthe G., S. Hulth and P.O.J. Hall (1998) Effect of oxygen on degradation rate of refractory and labile
organic matter in continental margin sediments. Geochimica Cosmochimica Acta 62: 1319-1328.
Jones C.G., J.H. Lawton and M. Shachak (1997). Positive and negative effects of organisms as physical
ecosystem engineers. Ecology 78: 1946-1957.
Jumars P.A., J.E. Eckman and E. Koch (2001). Macroscopic animals and plants in benthic flows. In: The
benthic boundary layer, edited by Boudreau and Jørgensen, p. 320-347, Oxford University Press.
Kaye J.P. and S.C. Hart (1997). Competition for nitrogen between plants and soil microorganisms. Trends
Ecology Evolution 12: 139-143.
Kenworthy W.J., J.C. Zieman and G.W. Thayer (1982). Evidence for the influence of seagrasses on the
benthic nitrogen cycle in a coastal plain estuary near Beaufort, North Carolina (USA). Oecologia 54:
King, G.M., M.J. Klug, R.G. Weigert and A.G. Chalmers (1982). Relation of soil water movement and
sulphide concentration to Spartina alterniflora production in a Georgia salt marsh. Science 218: 6163.
Kristensen E. (2000). Organic matter diagenesis at the oxic/anoxic interface in coastal marine sediments,
with emphasis on the role of burrowing animals. Hydrobiologia 426: 1-24
Kristensen E., S.I. Ahmed and A.H. Devol (1995). Aerobic and anaerobic decomposition of organic
matter in marine sediment: which is fastest? Limnology & Oceanography 40: 1430-1437.
Leonard L. A. & M.E. Luther, (1995). Flow hydrodynamics in tidal marsh canopies. Limnology &
Oceanography 40: 1474-1484.
McInerney M. and T. Bolger (2000). Decomposition of Quercus petraea litter: influence of burial,
comminution and earthworms. Soil Biology Biochemistry 32: 1989-2000.
Middelburg J.J., J. Nieuwenhuize, F.J. Slim and B. Ohowa (1996). Sediment biogeochemistry in an East
African mangrove forest (Gazi Bay, Kenya). Biogeochemistry 34: 133-155.
Middelburg J.J., K. Soetaert and P.M.J. Herman (1997a). Empirical relationships for use in global
diagenetic models. Deep-Sea Research I 44: 327-344.
Middelburg J.J., J. Nieuwenhuize, R.K. Lubberts and O. van de Plassche (1997b). Organic carbon isotope
systematics of coastal marshes. Estuarine Coastal Shelf Science 45, 681-687.
Burial of nutrients in coastal sediments
Middelburg J.J., C. Barranguet, H.T.S Boschker, P.M.J. Herman, T. Moens and C.H.R. Heip (2000). The
fate of intertidal microphytobenthos carbon: an in situ 13C labelling study. Limnology &
Oceanography 45: 1224-1234.
Paterson D.M. and K.S. Black (1999). Water flow, sediment dynamics and benthic biology. Advances
Ecological Research 29: 155-193.
Reddy K.R., W.H. Patrick and C.W. Lindau (1989). Nitrification-denitrification at the plant root-sediment
interface in wetlands. Limnology & Oceanography 34: 1004-1013.
Redfield A.C. (1972). Development of a New England salt marsh. Ecological Monographs. 42: 201-237.
Roden E.E. and R.G. Wetzel (1996). Organic carbon oxidation and suppression of methane production by
microbial Fe(III) oxide reduction in vegetated and unvegetated freshwater wetland sediments.
Limnology & Oceanography 41: 1733-1748
Rovira P. and V.R. Vallejo (2002). Mineralization of carbon and nitrogen from plant debris, as affected
by debris size and depth of burial. Soil Biology Biochemistry 34: 327-339.
Rozan T.F., M. Taillefert, R.E. Trouwborst, B.T. Glazer, S. Ma, J. Herszage, L.M. Valdes, K.S. Price and
G.W. Luther III (2002). Iron-sulphur-phosphorus cycling in the sediments of a shallow coastal bay:
Implications for sediment nutrient release and benthic macroalgal blooms. Limnology &
Oceanography 47: 1346-1354.
Ruttenberg K.C. (1992). Development of a sequential extraction method for different forms of
phosphorus in marine sediments. Limnology & Oceanography 37: 1460-1482.
Scoffin T.P. (1970). The trapping and binding of subtidal carbonate sediments by marine vegetations in
Bimini lagoon, Bahamas. Journal Sedimentary Petrology 40: 249-273.
Sun M.Y., R.C. Aller, C. Lee and S.G. Wakeham (2002). Effects of oxygen and redox oscillation on
degradation of cell-associated lipids in surficial marine sediments. Geochimica Cosmochimica Acta
11: 2003-2012.
Sundby, B., Vale, C., Cacador, I., Catarino, F., Madureira, M.-J. and M. Caetano (1998). Metal-rich
concretions on the roots of salt-marsh plants: mechanism and rate of formation. Limnology &
Oceanography 43: 245-252
Thomson J., Dyer F.M. and J.W. Croudace (2002). Records of radionuclide deposition in two salt
marshes in the United Kingdom with contrasting redox and accumulation conditions. Geochimica
Cosmochimica Acta 66: 1011-1023.
Valiela I. and M.L. Cole (2002). Comparative evidence that salt marshes and mangroves may protect
seagrass meadows from land-derived nitrogen loads. Ecosystems 5: 92-102.
Van de Koppel J., P.M.J. Herman, P. Thoolen and C.H.R. Heip (2001). Do alternate stable states occur in
natural ecosystems? Evidence from a tidal flat. Ecology 82: 3449-3461.
van der Nat F.J.W.A. and J.J. Middelburg (1998). Effect of two common macrophytes on methane
dynamics in freshwater sediments. Biogeochemistry 43: 79-104.
van der Nat F.J.W.A., J.J. Middelburg, D. van Meteren and A. Wielemakers (1998). Diel methane
emission patterns from Scirpus lacustris and Phragmites australis. Biogeochemistry 41: 1-22.
Westrich J.T. and R.A. Berner (1984). The role of sedimentary organic matter in bacterial sulphate
reduction: the G model tested. Limnology & Oceanography 29: 236-249.
Woodroffe C. (1992). Mangrove sediments and geomorphology. In: Tropical mangrove ecosystems,
edited by Robertson and Alongi, p. 7-41.
J.J. Middelburg et al.
J.J. Middelburg, K. Soetaert, P.M.J. Herman, H.T.S. Boschker & C.H.R. Heip:
Netherlands Institute of Ecology, Centre for Estuarine and Coastal Ecology, P.O.
Box 140, NL-4400 AC Yerseke, Netherlands.
In shallow coastal systems where most of the seafloor lies within the photic zone,
benthic photoautotrophy plays a key role in regulating nutrient cycling. In these
systems, production of submerged vascular plants (seagrasses), macroalgae and
benthic microalgae is high and typically exceeds that of phytoplankton (Borum and
Sand Jensen 1996). Changes in the patterns of primary production as well as in
habitat structure and trophic dynamics (Valiela et al. 1992; Nixon et al. 1996) have
been directly related to nutrient over-enrichment; this is also recognized as one of
the greatest threats to maintaining marine biodiversity in coastal regions (NRC
The widely-accepted scenario following nutrient enrichment is a shift in the
dominance of primary producers, from seagrasses and perennial macroalgae to fastgrowing green macroalgae and phytoplankton (Sand-Jensen and Borum 1991;
Duarte 1995; Valiela et al. 1997; see Chapter 3). Interestingly, only one of these
models (Sand-Jensen and Borum 1991) has included explicitly the potential
importance of benthic microalgae. Ultimately, one might expect a shift to
phytoplankton dominance in the most heavily eutrophied shallow estuaries (Duarte
1995; Valiela et al. 1997), even though total production might not change (Borum
and Sand-Jensen 1996). Given this transition, the question then becomes, how
might such a shift in primary producer dominance influence nutrient cycling in
shallow coastal waters?
The influence of primary producers on nutrient
transformations has important consequences for the role of shallow estuaries as a
buffer zone between land and sea (Fig. 1). Uptake and temporary retention of
nutrients in plant biomass, burial of recalcitrant organic matter, and the direct effects
of autotrophic metabolism on bacterially- and chemically-mediated processes all
influence nutrient cycling rates and pathways. Variations in the rates and
dominance of these processes as primary producer communities change, will
ultimately determine the fate and retention of watershed nutrients on their trajectory
to the open ocean.
S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling:
The IInfluence of Primary Producers, 231-261.
© 2004 Kluwer Academic Publishers. Printed in the Netherlands.
K.J. McGlathery, K. Sundbäck and I.C. Anderson
Fig. 1. Role of primary producers on the fate and retention of nutrients at the land-sea
In this chapter we address the influence of primary producer groups (seagrasses,
benthic microalgae, macroalgae) on nutrient cycling processes. We consider the
effects of plant metabolism on biogeochemical transformations within the sediments
and nutrient exchanges between the sediments and water column (see Chapter 10 for
discussion of denitrification). Our focus is on the retention and loss processes that
influence both the responses to nutrient enrichment and the transport of nutrients
through these shallow systems. We consider these processes in ecosystems that vary
with respect to tidal amplitude, water residence times, seasonality and sediment
type. Nutrient assimilation and turnover by benthic primary producers appear to
play a dominant role in regulating the magnitude and timing of nutrient fluxes in
these systems. This being the case, we pay particular attention to the competitive
interactions between primary producer communities and to the fate of nutrients
bound in plant tissue.
The role of benthic primary producer communities as sources or sinks of nutrients
reflects the net influences of assimilatory nutrient uptake, nitrogen fixation, leakage
by live and grazed autotrophs, oxygen release, nutrient production and losses due
mineralization, and nitrification – denitrification (Fig. 2). Nutrients can be
transferred between the sediment and water column as dissolved inorganic or
organic forms or as gases. Exchange rates are also influenced by benthic faunal
communities (e.g., Pelegri et al. 1994; Hansen and Kristensen 1997) and by physical
factors such as sediment resuspension and porewater advection. Often there are diel
Benthic N and P cycling
patterns driven by the primary producers, with an efflux of nutrients during the dark
and uptake in the light (e.g., Eyre and Ferguson 2002). The balance between
nutrient losses and gains in benthic communities also is influenced by nutrient
transfer via mobile foraging animals (see Chapter 6), and export or import of
particulate organic matter (see Chapter 4). Nutrient cycling in benthic primary
producer communities appears to be relatively conservative, with much of the
nutrient demand met by recycling within the sediments. For example, Ottosen et al.
(unpublished data) found by field 15N-labelling of sandy sediment that 90% of
labeled 15 PON disappeared from the top 3 cm of sediment after 12 weeks. This
rather long retention could be explained if remineralized N was efficiently recycled
within the microbial mat (Lomstein et al. 1998), and could be responsible for
sustaining high rates of productivity by benthic microalgae during summer in lownitrogen sandy sediments (Sundbäck et al. 2000). A similar closed recycling of both
N and P also appears to apply to benthic microalgal communities in oligotrophic
carbonate sediments of coral reefs (Miyajima et al. 2001; Suzumura et al. 2002) and
for both temperate and tropical seagrass systems (Alcoverro et al. 2000; McGlathery
et al. 2001b; Miyajima et al. 2001; Risgaard-Petersen and Ottosen 2001).
Conservative recycling of sediment N to support autotrophic communities is also the
common pattern observed in Spartina alterniflora vegetated intertidal sediments
(Anderson et al. 1997). Although the nutrient demands of seagrasses appear to be
met in large part by internal recycling of organic-bound nutrients in the sediment
(Hemminga et al. 1991; Alcoverro et al. 2000; McGlathery et al. 2001b), seagrasses
and other primary producers in the community also obtain external nutrients that are
transported by water flow. Resorption of nutrients from senescing leaves reduces the
demand for either recycled or external nutrients in seagrass (e.g., Pedersen and
Borum 1993; Stapel and Hemminga 1997; Alcoverro et al. 2000) and in salt marsh
systems (Anderson et al. 1997).
2.1 Seagrasses
Seagrasses have a major impact on sediment nutrient cycling by leakage of oxygen
and dissolved organic matter (DOM) from the roots. Both processes are
photosynthetically-driven. Oxygen produced by photosynthesis in the leaves is
transported to the roots to support aerobic respiration via a well-developed lacunal
system, a series of air channels comprising up to 60% of the total plant volume. The
oxygen that is not respired by the roots is released into the rhizosphere, creating
oxidized zones in an otherwise reducing environment. At the same time, seagrasses
release DOM from the roots, typically as simple organic carbon compounds that are
the recent products of photosynthesis (Koepfler et al. 1993; Ziegler and Benner
Bacterial productivity in the rhizosphere has been linked to these organic exudates.
Moriarty et al. (1986) showed that 11% of recent photosynthate was released within
6 hours into the rhizosphere of the tropical seagrass Halodule wrightii. Seasonal and
light-induced variation in bacterial activity and the typical sub-surface peak in
bacterial processes in the sediments where root biomass is highest, also clearly
indicate that root exudates stimulate bacterial activity (e.g., Welsh et al. 1996;
McGlathery et al. 1998; Blaabjerg et al. 1998).
K.J. McGlathery, K. Sundbäck and I.C. Anderson
N, P
N, P
Particulate Nutrients
Fig. 2. Influence of benthic primary producers on nutrient cycling processes: assimilatory
uptake and leakage, release of oxygen and dissolved organic matter (DOM) that affects
microbial activity, and mineralization and/or burial of particulate organic matter. Arrows
reflect the attenuation of light and the reduction of nutrient fluxes across the sediment-water
interface by benthic primary producer communities. DNR=denitrification, N-fix=nitrogen
For some seagrass species, including Potamogeton perfoliatus and Zostera marina,
it has been shown that oxygen released from plant roots and the creation of oxicanoxic interfaces in the rhizosphere stimulates nitrification, which produces nitrate
to support enhanced denitrification (e.g., Flindt 1994; Cornwell et al. 1999). This
stimulation may be linked to the seagrass growth cycle and to variations in oxygen
release (Caffrey and Kemp 1992; Risgaard-Petersen et al. 1998). However, recent
work using the isotope-pairing technique suggests that at least for the temperate
seagrasses Z. marina and Z. noltii, enhanced coupled nitrification-denitrification in
the rhizosphere is not generally observed (Rysgaard et al. 1996; Risgaard-Petersen
and Ottosen 2000, see Chapter 10). The mechanisms responsible for this lack of
Benthic N and P cycling
stimulation are thought to be low rates of rhizosphere nitrification and competition
for ammonium between nitrifying bacteria and benthic microalgae (Ottosen et al.
1999; Risgaard-Petersen and Ottosen 2000). This does not preclude the possibility
that enhanced denitrification may occur in sediments vegetated by other seagrass
species, in particular in tropical species such as Thalassia testudinum, which allocate
more biomass to roots and rhizomes than temperate species, and which tend to be
strongly P-limited in carbonate sediments so that competition between bacteria and
plants for N is unlikely to occur. Organic matter concentrations and bacterial
oxygen demand tend to be lower in tropical carbonates sediments, which may make
oxygenation of the rhizosphere more efficient in these systems. There are relatively
few measurements of denitrification in seagrass-vegetated carbonate sediments in
tropical/subtropical environments, but rates tend to be high (Blackburn et al. 1994;
Miyajima et al. 2001; Kemp and Cornwell 2001). The available data suggest that
both denitrification rates and nitrogen fixation rates are higher in tropical/subtropical
seagrass meadows than in temperate seagrass meadows (Table 1).
Nitrogen fixation activity is enhanced in seagrass meadows relative to unvegetated
sediments, both in temperate and tropical environments (e.g., O’Donohue et al.
1991; Welsh et al. 1996; McGlathery et al. 1998; Hansen et al. 2000). The
availability of organic substrate is a major factor controlling N fixation rates in
seagrass sediments. Many studies have shown increases of N fixation by addition of
organic compounds, and seasonal and diel variations that are consistent with the role
of photosynthetic exudates stimulating N-fixing bacteria (e.g., McGlathery et al.
1998, Hansen et al. 2000; Welsh 2000). Much of this activity (25 – 95%) has been
associated with sulfate reducers (e.g., Welsh et al. 1996; McGlathery et al. 1998),
which also are believed to be stimulated by root exudates (Blaabjerg et al. 1998;
Hansen et al. 2000). Depth profiles of N fixation activity in Z. marina sediments
showed a close association with root/rhizome biomass, with seasonal shifts in the
magnitude and depth distribution of N fixation matching the shifts in belowground
biomass (McGlathery et al. 1998). Overall, N fixation rates in sediments of seagrass
meadows are higher in tropical/subtropical systems than in temperate systems (see
references in McGlathery et al. 1998 and Welsh 2000), and this may in part be due
to increased DOM release from the greater root/rhizome biomass of many tropical
seagrass species. Autotrophic, heterocystous cyanobacterial epiphytes on seagrass
leaves also contribute to the total N fixation rates, especially in warm tropical
environments, although on an area basis the activity is lower than that of
heterotrophic bacteria in the sediments (e.g., Capone and Taylor 1977; O’Donohue
et al. 1991).
Few studies have measured N fixation and denitrification
simultaneously in seagrass beds to determine if N gains by fixation can compensate
for N losses via denitrification, and the results are equivocal (Table 1). In a tropical
seagrass meadow, Blackburn et al. (1994) observed that N fixation rates accounted
for less than half of the N loss by denitrification, whereas in a temperate seagrass
meadow, Risgaard-Petersen et al. (1998) found that N fixation more than
compensated for the N loss by denitrification.
Both oxygen and DOM release by seagrass roots influence the forms and availability
of sediment P. Oxidation in the rhizosphere results in the formation of iron oxides
that effectively bind P in the solid phase. The redox conditions in the sediment and
the formation of autogenic P minerals influence the release of P to overlying waters
(Rozan et al. 2002). In carbonate sediments that typically dominate tropical seagrass
K.J. McGlathery, K. Sundbäck and I.C. Anderson
systems, P turnover rates are generally high in the porewaters relative to N,
indicating that P is preferentially removed (Hines and Lyons 1982; McGlathery et
al. 2001b). The mechanisms responsible for this include surface adsorption onto
carbonate grains, direct precipitation of Ca-P mineral phases, and uptake by Plimited primary producers. In vegetated carbonate sediments in Bermuda, Jensen et
al. (1998) found that only 2% of the total sediment P was in the loosely-adsorbed
pool that could be readily released to the porewater; 15 to 20% was more strongly
adsorbed to the surface of the sediment, and the remainder (80%) was bound in the
mineral matrix of the sediments, probably in the form of carbonate fluorapatite and
other carbonates.
It was previously thought that carbonate sediments were a permanent sink for P in
these tropical environments and that mineral P was not available to plants.
However, recent studies suggest that seagrass metabolism facilitates the dissolution
of carbonate minerals in the rhizosphere releasing bound P (Jensen et al. 1998;
Burdige and Zimmerman 2003).
One mechanism for this is the decrease in pH resulting from root respiration or the
stimulation of bacterial respiration by root DOM release (Burdige and Zimmerman
2003). The oxidation of sulfides in the rhizosphere may also decrease pH, as has
been shown for saltmarsh grasses (Giblin and Howarth 1984). Another mechanism
is the release of organic acids from the roots that also acidify the rhizosphere and
cause the dissolution of the carbonate minerals, as is the case for terrestrial plants
(Hoffland 1992; Knight et al. 1992).
Iron associated with carbonate minerals can serve as a sorption site for inorganic P
(Jensen et al. 1998), and also can react with sediment sulfides to precipitate ironsulfide minerals that reduce sulfide toxicity to primary producers (Carlson et al.
1994; Erskine and Koch 2000). Chambers et al. (2001) found that iron additions
increased the total sediment P pool and reduced exposure of the seagrass T.
testudinum to free sulfide, but found no growth response probably because most of
the sediment P was bound in unavailable forms. One might expect a more
significant effect of iron additions on seagrasses in carbonate sediments where P is
not as strongly limiting (i.e., where P is in more readily exchangeable forms). On an
ecosystem scale, seagrass systems can be important temporary sinks for nutrients,
although there is less information than for benthic micro- and macro-algal
communities. The source-sink role will vary seasonally, depending on the growth
requirements of seagrasses. Temperate Z. marina meadows may act as strong sinks
for N from the beginning of the growing season in the spring to the time of
maximum productivity in late summer; in the fall when decomposition rates are
typically relatively high, but plant N demand low, the seagrass community may
become a source of nutrients (Risgaard-Petersen et al. 1998; Risgaard-Petersen and
Ottosen 2000). Denitrification was an important sink relative to plant uptake in
these seagrass meadows only in winter when both plant demand and decomposition
were low (Risgaard-Petersen and Otteson 2000). Since N fixation may only support
a small percentage of seagrass N demand in these systems (Welsh 2000 and
references therein), most of the nutrient influx is likely uptake of water column
nutrients by seagrass leaves and associated primary producers (benthic microalgae,
epiphytes, macroalgae) and accumulation of imported organic matter sedimented
into the meadow.
Benthic N and P cycling
Table 1. Comparison of nitrogen fixation and denitrification rates in temperate and
tropical/subtropical seagrass meadows.
Zostera marina
Zostera marina
Zostera marina
Zostera noltii
N Fixation
mmol N m-2 d-1
Vaucluse Shores,
0.07 – 0.43
0.10 – 0.41
Arcachon Bay,
Arcachon Bay
and Etang du
Prevost, France
Edmunds Bay,
Oyster Bay,
Florida Bay,
Zostera noltii
Gulf of
Gulf of
Gulf of
Moreton Bay,
mmol N m-2 d-1
0.05 – 0.14
McGlathery et
al. (1998)
RisgaardPetersen et al.
Welsh et al.
Rysgaard et
al. (1995)
Eyre &
0.7 – 4
1.6 - 3
Blackburn et
al. (1994)
Kemp &
Moriarty &
Moriarty &
Moriarty &
O’Donohue et
al. (1991)
1.1 - 3.4
0.7 – 2.9
K.J. McGlathery, K. Sundbäck and I.C. Anderson
Seagrasses enhance nutrient concentrations in the sediments, both in particulate and
dissolved forms, by an increased input of organic matter from plant biomass and
imported material. Roots and rhizomes typically decay in situ, and some
aboveground material may remain in the system rather than be exported by water
flow. However, as with salt marsh ecosystems, it appears that much, if not most, of
the organic matter buried and decomposed in seagrass sediments is imported
(Pedersen et al. 1997; Boschker et al. 2000; Gacia et al. 2002; see Chapter 8). Fine,
organic-rich particles settle out as the water flow is slowed by the seagrass canopy,
and the roots and rhizomes also help to stabilize the sediments and prevent
resuspension. As a result, organic matter and nutrient concentrations are typically
higher in vegetated than in unvegetated sediments. Nutrient accumulation may
occur in seagrass meadows over the time scale of years, but this source-sink role
may in part depend on the trophic status of the system. Pedersen et al. (1997) found
that colonization and stand development of the seagrass Cymodocea nodosa in
Mediterranean carbonate sediments resulted in a net accumulation of N, but not P,
over a 5 year period, with about half of the accumulated N accounted for by living
and dead plant material. In C. nodosa meadows where production was nutrient
limited, as is the case during stand development, Perez et al. (1997) observed a net
accumulation of N and P in slowly-degrading belowground detritus, yet in eutrophic
waters where belowground production was reduced and decay rates of leaves were
high, the seagrass meadow becomes a source of nutrients.
2.2 Benthic microalgae
The importance of microphytobenthic production in shallow coastal areas is well
documented (Underwood and Kromkamp 1999), as are the effects of microalgal
production on N cycling. Photosynthesis and respiration by benthic microalgae in
the top few mm of the sediments cause large diel variations in oxygen
concentrations and penetration depth, dissolved inorganic carbon (DIC)
concentrations and pH (see Chapter 10). In addition, assimilation of porewater
ammonium and nitrate influences the concentrations and depth distributions of
porewater N. The net effect of these changes may be diel variations in nitrification
and denitrification activity (Risgaard-Petersen et al 1994, Lorenzen et al. 1998, An
and Joye 2001, see Chapter 10), with higher rates of coupled nitrificationdenitrification in light due to stimulated nitrification activity from photosynthetic
oxygen release. However, this is not always the case (Risgaard-Petersen 2003 and
references therein), and light stimulation may only occur when there is sufficient N
to prevent N limitation (Rysgaard et al. 1995; An and Joye 2001). Recent field data
from 18 European estuaries showed that net autotrophic sediments colonized by
microalgae had lower rates of coupled nitrification-denitrification than net
heterotrophic sediments (Risgaard-Petersen 2003). This was also the case for
shallow bays in Sweden, where denitrification rates were higher (20% of
remineralized N) in net heterotrophic than in net autotrophic sediments (10% of
remineralized N) (Sundbäck and Miles 2002).
Benthic microalgal communities in both tropical and temperate sediments are
generally dominated by diatoms and cyanobacteria. Nitrogen fixation rates in
Benthic N and P cycling
surface sediments, and on seagrass leaves, tend to be high relative to those in the
water column, and to be stimulated by light (e.g., Capone and Taylor 1977; Capone
et al. 1992). Significant autotrophic N fixation occurs even in seagrass-vegetated
sediments where the leaf canopy shades the sediment surface, although areaintegrated rates generally are lower than heterotrophic N fixation (e.g., O’Neil and
Capone 1989). Rates are generally higher in tropical carbonate sands than in
temperate siliclastic sediments (e.g., Howarth et al. 1988; Capone et al. 1992).
There is currently little evidence of diurnal variations in N fixation activity in
benthic microalgal communities in submersed sediments, however, Currin et al.
(1996) found a night-time peak in N fixation in filamentous cyanobacterial mats and
a day-time peak in coccoid cyanobacterial mats in intertidal sediments.
Among the primary producers of shallow-water areas, benthic microalgae appear to
be the component whose quantitative role in nutrient cycling is least understood at
the ecosystem level. Integrating the effect of benthic microalgae at the ecosystem
scale is difficult. Although benthic microalgae are present and active throughout the
year (Rizzo et al. 1996 and references therein), their relative quantitative importance
for nutrient turnover depends on the presence, seasonal pattern and area coverage of
other primary producers, as well as physical characteristics of the embayment. For
example, McGlathery et al. (2001a) found that the contribution of benthic
microalgae to carbon production varied from 4-99% during the year in a shallow
lagoon. The influence of benthic microalgae on net autotrophy and N cycling in the
sediments appears to be particularly important in cool microtidal areas (Sundbäck et
al. 2000), where the growth of seagrasses and ephemeral floating macroalgae is
confined to the warm season (May-September). In warmer systems, macroscopic
primary producers may occur during most of the year, implying more continuous
shading and nutrient competition, although this has been observed primarily in
eutrophic systems (e.g., Sfriso et al. 1992; Viaroli et al. 1996).
Studies performed in different types of shallow-water ecosystems, ranging from cool
temperate (e.g. Rizzo et al. 1992, 1996; Cerco and Seitzinger 1997, Sundbäck and
Miles 2000; Thornton et al. 2002) to warm temperate (Eyre and Ferguson 2002) to
subtropical and tropical areas (Miyajima et al. 2001; Suzumura et al. 2002), suggest
that benthic microalgae may be an important temporary sink for nutrients. Studies
from temperate, relatively N poor areas, suggest that benthic microalgae may turn
sediments into sinks of N for the majority of the year, but particularly in winter and
spring (e.g., Sundbäck et al. 2000). In such sandy areas, N may be limiting for
benthic microalgae (Nilsson et al. 1991). Also in P-limited oligotrophic subtropical
carbonate sediments, microalgal activity has been found to enhance accumulation of
combined N (Miyajima et al. 2001).
2.3 Macroalgae
Bloom-forming macroalgae have become dominant in nutrient-enriched waters over
at least the last two decades throughout temperate and tropical regions (e.g.,
Cambridge & McComb 1984; Sfriso et al. 1992; Viaroli et al. 1996; Valiela et al.
1997; Pihl et al. 1999). These “nuisance” algae are typically filamentous or sheetlike forms -- mainly chlorophytes (e.g., Ulva, Cladophora, Chaetomorpha) -- that
can accumulate in extensive thick mats over the sediment surface or in the water
K.J. McGlathery, K. Sundbäck and I.C. Anderson
column. Variations in light availability within the mats, and in nutrient, oxygen, and
pH conditions linked to algal photosynthesis, respiration, and decomposition,
influence nutrient cycling at the sediment-water interface, and perhaps even deep
into the sediments. The presence of dense macroalgal mats can move the location of
the oxic-anoxic interface up from the sediments into the macroalgal mat since only
the upper few cm of the mat may be within the photic zone (Krause-Jensen et al.
1999; Astill and Lavery 2001).
Sediment nutrient cycling is enhanced by the presence of macroalgae, presumably
due to the input of organic matter (Trimmer et al. 2000; Tyler et al. 2003).
Increased mineralization rates cause a build up of ammonium deep within the mat
where light does not penetrate (McGlathery et al. 1997; Astill and Lavery 2001), and
may affect porewater ammonium as deep as 13 cm in the sediment (Burton et al,
unpublished data). This recycling of organic-bound nitrogen may provide an
important nitrogen source to sustain macroalgal production, particularly in areas
where tidal exchange is limited (McGlathery et al. 1997; Trimmer et al. 2000).
Increased ammonium concentrations at depth in sediments may also be partly
related to local extinction of bioturbating infauna caused by anoxia (Hansen and
Kristensen 1997). Where bioturbating infauna is present, increased oxygen diffusion
into the sediments may stimulate nitrification rates (Rysgaard et al. 1995; Hansen
and Kristensen 1997).
Compared to benthic microalgae and seagrasses, there is little information on the
influence of macroalgae on denitrification rates. In studies performed in harbors on
the south coast of England, Trimmer et al. (2000) found that both direct
denitrification, using nitrate supplied from the water column and denitrification
couple to sedimentary nitrification were low in sediments underlying macroalgal
mats. High free sulfide concentrations occur in organic-rich sediments underlying
macroalgal accumulations (Viaroli et al. 1996), and these conditions may inhibit
nitrification (Henriksen and Kemp 1998; Sloth et al. 1995) and partially account for
the low denitrification rates. In a laboratory experiment, Krause-Jensen et al. (1999)
showed that the net effect of dense macroalgal mats was to move the zone of
denitrification from the sediments up into the mat, but not to influence the rates
significantly. Denitrification rates peaked at the oxic-anoxic interface in the middle
of the mat. Presumably the same regulating mechanisms would apply over the
larger spatial scale in a macroalgal mat as have been determined in benthic
microalgal communities in the sediment. Macroalgal photosynthesis and respiration
induce diurnal variations in oxygen penetration depth within the upper productive
layers of the mat, which in turn, may cause diurnal variations in nitrification activity
and in coupled nitrification-denitrification. Mineralization of algal-bound nutrients
in the lower layers of the mat where there is insufficient light for photosynthesis
creates a diffusion gradient of ammonium up into the oxidized layers. Like benthic
microalgae, it is expected that macroalgae will compete with bacteria for ammonium
and nitrate. It appears that macroalgae do not stimulate denitrification, however,
more work is needed to address the controls on denitrification in sediments
underlying benthic macroalgal communities and within the mats themselves.
At the ecosystem scale, macroalgae can store significant quantities of nutrients. In
highly enriched waters, it is not unusual for macroalgal populations to attain peak
biomass of over 0.5 kg dry wt m-2 and for canopy heights to exceed 0.5 m. In
Waquoit Bay, Massachusetts, nitrogen stored in peak macroalgal biomass was of the
Benthic N and P cycling
same magnitude as the annual nitrogen load from the watershed (Valiela et al.
1997). Many studies have shown sediments overlain with macroalgae to be a net
sink for inorganic nutrients (e.g., Krause-Jensen et al. 1996; Tyler et al. 2001). In
Hog Island Bay, Virginia, macroalgae were a net sink for DIN throughout the year,
and at the same time a net source for organic N (Tyler et al. 2003); urea and
dissolved free amino acids (DFAA) were sequestered by macroalgae, but up to 22%
of the total nitrogen uptake was released as dissolved combined amino acids
(DCAA). On a seasonal time scale, it is common for macroalgae in eutrophic waters
to switch from being a net sink early in the growing season, to a net source of
nutrients in late summer when productivity declines due to self-shading within the
mat and high temperatures increase respiration.
High rates of macroalgal metabolism can lead to periods of oxygen supersaturation
or, following senescence, oxygen depletion throughout the entire water column (e.g.,
Sfriso et al. 1992, Valiela et al. 1992, Boynton et al. 1996). Anoxic events in
macroalgal-dominated systems tend to be episodic and to occur during the warm
summer months when macroalgal production declines. Such “boom and bust’
cycles of high production and senescence are typical of dense macroalgal
populations in eutrophic waters (e.g., McComb and Humphries 1992, Sfriso et al.
1992, Valiela et al. 1997). In Hog Island Bay, Virginia, Tyler et al. (2001) reported
dissolved inorganic and organic N release rates following the collapse of a
macroalgal bloom (primarily composed of Gracilaria tikvahiae and Ulva lactuca)
that were sufficient to result in complete mineralization of the macroalgal biomass
(up to 650 gdw m-2) within approximately 13 days. Release of plant-bound nutrients
following these “dystrophic” events may stimulate phytoplankton and bacterial
metabolism in the water column (Valiela et al. 1997; McGlathery et al. 2001a;
Lunsford 2002).
Nutrient assimilation by benthic primary producers can reduce the efflux of
remineralized nutrients from the sediments to the overlying water, effectively
decoupling nutrient turnover within the sediments from water column processes.
During times of the year when primary producer populations are most active, this
“filter” effect completely intercepts nutrient fluxes across the sediment-water
interface and thereby reduces nutrient availability for phytoplankton and bacteria, as
well as for floating macroalgal mats, in the overlying water. If the sediment nutrient
sources are insufficient to meet growth demands, there also may be a downward flux
of nutrients from the water column to the benthic community.
It is generally assumed that in littoral zone ecosystems biogeochemical processes
occurring in the benthos are closely coupled with those in the pelagic zone. For
example, primary production in the water column supports heterotrophic metabolism
in underlying sediments. In turn, organic matter remineralization in sediments
releases nutrients that support primary production in overlying water. In situations
where benthic – pelagic coupling is strong it should be possible to calculate
ammonium fluxes out of sediments based on measured rates of net ecosystem
metabolism (NEM; dissolved oxygen uptake or DIC release) (Hopkinson et al.
K.J. McGlathery, K. Sundbäck and I.C. Anderson
2001). In oxic sediments where nitrification is likely to play an important role in
transformation of ammonium, it may be the total dissolved inorganic nitrogen (DIN)
flux that is more closely related to the stoichiometry of NEM, whereas in anoxic
sediments where denitrification is important the estimated DIN flux will not equal
that predicted by NEM stoichiometry. Calculation of denitrification has often been
based upon this “missing” DIN. Using 15N isotope dilution techniques, it is possible
to measure both gross N mineralization and consumption of ammonium in
sediments amended with 15NH4+, thereby allowing determination of net
mineralization rates (Anderson et al, 2003). Net mineralization may also be
measured during incubation of unamended sediment/water cores. Where gross and
net mineralization and sediment-water column exchanges of DIN have been
measured simultaneously in autotrophic sediments, gross rates of ammonium
production generally exceed net rates by a large margin and net production of
ammonium in the sediments is often not reflected in fluxes to the water column
(e.g., Kristensen et al. 2000, Anderson et al. 2003). Benthic microalgal assimilation
is likely to be largely responsible for this (Anderson et al. 2003), although sediment
bacteria may also play a role in N immobilization depending on the C/N of the
sediment organic matter undergoing decomposition. In sediments with large
amounts of macrophyte detritus with C/N as high as 45, and with the short turnover
times typically observed for the ammonium pool (<1 d; Kristensen et al. 2000;
Anderson et al. 2003), bacterial immobilization of N would be expected. In general,
benthic microalgal uptake probably represents a major fate of mineralized N
compared to coupled nitrification – denitrification in net autotrophic sediments.
This has been observed in Hog Island Bay, Virginia, sediments (Anderson et al.
unpublished), in southern England sediments dominated by macroalgae (Trimmer et
al. 2000), in the Tagus Estuary, Portugal (Cabrita and Brotas 2000), and at sites
along the west coast of Sweden (Sundbäck et al. 2000). The DIN fluxes observed in
these littoral zone systems differ markedly from those observed in sublittoral zone
systems, which typically exhibit release of ammonium or ammonium plus nitrate,
closer to values predicted by NEM (e.g., Giblin et al. 1997; Burdige and Zheng
1998; Hopkinson et al. 1999).
The idea that benthic microalgae may significantly modify the flux of nutrients
between sediment and the overlying water column was put forward more than 20
years ago (Henriksen et al. 1980). As the density of benthic microalgae – and other
microbiota– in the top few mm of sediment is orders of magnitude higher than in the
water column, such a filter function should be expected, but had been previously
overlooked because flux measurements were traditionally made only in the dark.
More recently, several studies have shown that the presence of benthic microalgae
can indeed significantly decrease the net efflux of N, P and silica, or even change
the direction from efflux to influx (e.g., Reay et al.1995, Cerco and Seitzinger 1997;
Sundbäck et al. 2000; Suzumura et al. 2002; Anderson et al. 2003; Tyler et al. 2003;
see Table 2). This filter effect is mediated by both microalgal assimilation and
photosynthetic oxygenation of the sediment surface, which in turn creates dynamic
chemical and biological gradients. High-resolution microsensors have made it
possible to study the influence of photoautotrophic activity on N turnover in more
detail (see Chapter 10).
Benthic N and P cycling
The majority of studies on the regulatory effect of benthic microalgae on sedimentwater nutrient fluxes are from light/dark incubations of flooded sediment,
mimicking subtidal conditions.
Such incubations do not account for the effect of emersion on fluxes in intertidal
sediments that occur in many shallow estuaries (e.g., Falcao and Vale 1995).
However, Thornton et al. (1999) showed that illumination during emersion reduced
fluxes of ammonium to the overlying water after immersion and concluded that the
effect of the illumination and emersion periods should be considered when
calculating nutrient budgets for intertidal, cohesive sediments.
Like benthic microalgae, accumulations of benthic macroalgae can function
temporarily as a filter by intercepting dissolved nutrient fluxes from the sediments to
the overlying water (Thybo-Christensen et al. 1993; McGlathery et al. 1997; Tyler et
al. 2001). The effectiveness of this interception depends on the biomass and
productivity of the macroalgae as well as on the stability of macroalgal
accumulations. Unattached macroalgal mats tend to be patchy and unstable;
however, Astill and Lavery (2001) showed that nutrient and oxygen gradients
developed quickly as macroalgae accumulated (within 24 hr), suggesting that this
filter function occurs even in dynamic environments.
Table 2. Ammonium and nitrate fluxes from sediments in two temperate locations, illustrating
the effect of benthic microalgae on the magnitude and direction of nutrient fluxes. Values
represent annual means (± se) from monthly – bimonthly measurements. Fluxes from Hog
Island Bay sediments show the influence of net autotrophy (data from Tyler et al. 2003), while
those from the Swedish coastal bays show the influence of sediment type (data from Sundbäck
et al. 2000).
Hog Island Bay, Virginia
Autotrophic sediments
Heterotrophic sediments
Swedish west coast
Sandy sediments – Vallda
Ammonium Flux
µmol N m-2 d-1
Nitrate Flux
µmol N m-2 d-1
-327 ± 104
-115 ± 36
418 ± 90
384 ± 126
-78 ± 29
-80 ± 27
9 ± 47
6 ± 13
108 ± 113
145 ± 72
337 ± 124
-87 ± 64
-396 ± 325
-46 ± 130
306 ± 110
-161 ± 149
Most studies have focused on macroalgal interception of dissolved inorganic
nutrient fluxes (e.g., Thybo-Christesen et al. 1993; McGlathery et al. 1997);
however, macroalgae also may be important in regulating the flux of dissolved
organic nutrients across the sediment-water interface. Tyler et al. (in press) showed
that the chlorophyte Ulva lactuca prevented the flux of urea and DFAA from the
K.J. McGlathery, K. Sundbäck and I.C. Anderson
sediments to the overlying water. Sediments are considered to be a significant
source of nutrients to supporting production of macroalgae, since nutrient
concentrations are typically one to several orders of magnitude higher in sediment
porewaters than in the water column, even in eutrophic waters. This has been shown
in Hog Island Bay, Virginia where the efflux of DIN and urea from sediments
underlying macroalgal mats was sufficient to meet 27-75% of the macroalgal N
demand (Tyler et al. 2003). Likewise, Stimson and Larned (2000) determined that
the N efflux from the sediments of a subtropical bay was essential to meet the
growth demand of the dominant macroalgae and was at least in part responsible for
the persistence of macroalgae in an otherwise oligotrophic environment. Sediment
nutrient efflux is typically consumed in the bottom layers of dense macroalgal mats,
and diffusion from the overlying water provides a necessary nitrogen supply for the
light-saturated algae in the upper layers (McGlathery et al. 1997).
Factors that reduce macroalgal productivity, such as decreased water clarity from
suspended sediments or high phytoplankton biomass, decreased insolation, or selfshading within the lower layers of dense mats during summer also decrease
macroalgal uptake and reduce the filter effect (Sfriso et al. 1992; D’Avanzo and
Kremer 1994; McGlathery et al. 1997). Often one, or a combination of these
factors, leads to the late-summer collapse of the macroalgal population, and the
release of plant-bound nutrients to the water column. Macroalgal mats that are lightlimited may show a diurnal pattern of nutrient release, with nutrients diffusing from
the sediments through the mat to the water column in the dark (McGlathery et al.
1997). This occurs because algae growing in low-light conditions are typically Nsaturated and lack carbon reserves; as such, they are dependent on recent
photosynthate to build amino acids. This close coupling between nutrient uptake
and photosynthesis is also characteristic of N-replete phytoplankton (Turpin 1991
and references therein).
There have been fewer studies of mechanisms by which seagrass communities
influence benthic dissolved nutrient fluxes. Like benthic micro- and macro-algae,
one would expect assimilatory nutrient uptake by seagrasses to reduce the flux of
nutrients across the sediment-water interface. Risgaard-Petersen & Ottosen (2000) in
a study performed in Denmark showed a seasonal pattern of DIN exchange between
eelgrass (Z. marina) communities and the water column similar to that which has
been shown for sediments with benthic microalgal and macroalgal communities.
There was an influx of N to eelgrass-vegetated sediments in the spring and summer
when plant metabolism was highest, and a smaller influx in the colder winter
months. In the fall when plant N demand was low, but decomposition rates were
relatively high, nutrients were released to the water column. In studies performed in
Virginia, where seagrass growth is high in spring and fall and decomposition rates
high in July - August, seasonal patterns of nutrient exchange were slightly different
from those in Denmark in that there was a large efflux of DIN (primarily
ammonium) to the water column in June and August and a negligible efflux in April
and October (Anderson and Moore, unpublished data). Both studies illustrate the
important influence of assimilatory nutrient uptake by seagrasses on the direction of
nutrient fluxes at the sediment-water interface. In tropical carbonate sediments, P
fluxes are typically low (nM), in part because the P is bound in the solid phase
(Jensen et al. 1998) and in part because turnover rates in sediment porewaters are
high to meet seagrass demand (McGlathery et al. 2001b).
Benthic N and P cycling
Other primary producers associated with seagrass communities also can play a key
role in regulating the magnitude and direction of the nutrient flux. Miyajima et al.
(2001) found an influx of N in both bare and seagrass-vegetated carbonate
sediments, and attributed this at least in part to high benthic microalgal activity. The
presence of epiphytes on seagrass blades also may be important, as Eyre and
Ferguson (2002) found a net influx of N in warm-temperate Australian seagrass
communities during light incubations only when epiphytes were included.
Likewise, it has been shown that epiphytes can remove more ammonium and nitrate
from the water column than the seagrass Ruppia megacarpa even though their
biomass was significantly less (Dudley et al. 2001).
There is ample evidence to support the general hypothesis that phytoplankton and
benthic primary producers compete for resources (nutrients, light) in shallow coastal
waters (e.g., Sand-Jensen and Borum 1991; Fong et al. 1993), and that the outcome
of this competition is unstable. Likewise, competition between benthic primary
producers occurs, especially in systems that receive high external nutrient loads
(Valiela et al. 1997 and references therein). Shifts in primary producer dominance
that result from this competition have an impact on the rates and pathways of
nutrient transformations within these systems.
Most studies indicate that at certain times of the year the reduction, or elimination,
of the internal supply of nutrients to the water column from sediment regeneration
that is attributed to benthic microalgal uptake decreases nutrient availability for
phytoplankton (e.g., Sundbäck et al. 2000; Suzumura et al. 2002). However, Cerco
and Seitzinger (1997) have suggested that benthic microalgae actually enhance
phytoplankton production on an annual time scale. By sequestering mineralized
nutrients in the winter and spring, these authors suggest that benthic microalgae may
function as a temporary storage pool for regenerated nutrients; nutrients are released
in the summer when the benthic microalgae become light-limited by increased
phytoplankton biomass in the water column. The net effect is to extend the summer
phytoplankton bloom beyond the time when it would normally become nutrientlimited.
It has been suggested that uptake of mineralized nutrients in the sediment by benthic
microalgae also may inhibit, or at least defer, the initiation of benthic macroalgal
blooms in the spring. A study in shallow sheltered embayments in Sweden
suggested that benthic microalgae could successfully compete with ephemeral green
algae for the sediment pool of regenerated nutrients during the period that is critical
for the onset of the macroalgal growth (Sundbäck and Miles 2002). On a daily basis,
the benthic microalgae decreased the efflux of inorganic N by 30 – 100%, P by 70 100% and silica by 10–95%.
Dense macroalgal mats, in turn, decrease light availability at the sediment surface
(90-100% for macroalgal biomass 100-500 gdw m-2; Krause-Jensen et al. 1996;
Astill and Lavery 2001), and sediments underlying stable algal accumulations tend
to be heterotrophic (Tyler et al. 2003). Surprisingly, there have been few direct
studies of the effect of mat-forming macroalgae on benthic microalgal productivity
(Sundbäck et al. 1990, 1996a). Persistent, dense macroalgal mats would be expected
K.J. McGlathery, K. Sundbäck and I.C. Anderson
to decrease production of sediment microalgal communities by reducing light levels
and by creating anoxic conditions at the sediment surface. In Hog Island Bay,
Virginia, benthic microalgal production increased in importance following a midsummer macroalgal decline, suggesting that competition for light limited benthic
microalgal production when macroalgal densities were high (McGlathery et al.
2001a). A more recent study in this lagoon found an inverse relationship between
macroalgal biomass and both sediment chlorophyll a and sediment dissolved oxygen
production over an annual cycle (Tyler et al. 2003), indicating decreased benthic
microalgal activity in sediments underlying macroalgal accumulations. At lower
densities, or when macroalgae are floating in the water column and thereby do not
create intense shading and/or bottom water anoxia, sediment microalgal production
would not be as strongly influenced (Sundbäck et al. 1996a).
By sequestering and storing sediment-derived nutrients, dense macroalgal mats also
may outcompete phytoplankton during the spring and early summer when
macroalgal growth rates are highest (Viaroli et al. 1992; McGlathery et al. 1997;
Valiela et al. 1997). This probably accounts for the high water clarity in many
shallow macroalgal-dominated systems, despite high external nutrient loads (Valiela
et al. 1997). Macroalgae also can sustain maximum growth rates for longer periods
of time than phytoplankton because their nutrient storage capacity obviates the need
for a constant nutrient supply (Fong et al. 1993; Pedersen and Borum 1996). Stored
nutrients in phytoplankton can only sustain maximum growth for a day or less,
whereas growth of bloom-forming macroalgae such as Ulva, Cladophora and
Chaetomorpha can be sustained by stored nutrients for several days, or more,
depending on growth conditions (light and temperature) (Borum 1996). However,
the outcome of competition also can be reversed when phytoplankton blooms begin
to shade macroalgal communities. This usually happens when macroalgal
metabolism slows down in late summer and the macroalgae become a less effective
filter for the effluxed sediment nutrients. As a consequence, nutrients diffuse into
the water column and stimulate phytoplankton growth, which, in turn, further
reduces macroalgal productivity through shading. Valiela et al. (1997) observed
interannual variation in Waquoit Bay, Massachusetts, that was consistent with this
model. In the most N-enriched subestuary, phytoplankton abundance and activity
increased at the same time that macroalgal biomass declined. Many studies have
shown that low macroalgal biomass coincides with peak phytoplankton biomass
(e.g., Sfriso et al. 1992; Viaroli et al. 1992). Typically, these moderately-enriched
systems are unstable, and are characterized by shifts in dominance between benthic
macroalgae and phytoplankton. The same scenario has been suggested for systems
in which microalgae dominate the benthic community and where shading by
phytoplankton blooms can cause a temporary collapse of benthic microalgal
productivity (Blanchard and Montagna 1995; Cerco and Seitzinger 1997).
The competition between rooted macrophytes and benthic microalgae has not been
studied specifically. One might expect low benthic microalgal biomass and
production in densely-vegetated seagrass systems due to shading of the sediment by
the leaf canopy, yet many studies have shown that benthic microalgae can contribute
significantly to the total productivity of seagrass beds (e.g., Murray and Wetzel
1987; Moncrieff et al. 1992). The same has been observed in Spartina alterniflora
vegetated intertidal sediments (Anderson et al. 1997). In some seagrass systems,
benthic microalgal biomass and primary production has been found to be as high as
Benthic N and P cycling
in bare sediments (e.g., Moncreiff et al. 1992, Pollard and Kogure 1993, Kemp and
Cornwell 2001), or even higher (Miyajima et al. 2001). The contribution of this
production to nutrient retention and turnover in seagrass beds is unclear. Hansen et
al. (2000) found that uptake by benthic microalgae in a temperate seagrass bed was
the major sink for DIN in the spring. Likewise, Kemp and Cornwell (2001) found
that benthic microalgal N uptake was as important as that by seagrasses in Florida
Bay sediments in August. It is likely that the interaction between benthic microalgae
and seagrass varies seasonally with changes in light and nutrient availability as the
seagrass canopy develops.
In nutrient-enriched systems, macroalgal blooms can outcompete seagrasses,
although the outcome of this competition is reversible if macroalgal biomass
decreases. Hauxwell et al. (2000) showed that dense macroalgal canopies caused a
decrease in recruitment of new shoots and rates of leaf appearance in existing shoots
of the eelgrass Z. marina in an enriched subestuary of Waquoit Bay, Massachusetts.
Using model calculations to estimate light attenuation by the different autotrophs
(phytoplankton, epiphytes, macroalgae), they confirmed that the primary cause of
eelgrass loss was light reduction by the macroalgal canopy, in particular for newly
recruiting shoots. One might expect that some tropical seagrasses, such as Thalassia
testudinum, that have a proportionately greater allocation to belowground biomass
(and carbon reserves) may be more resilient to ephemeral shading by macroalgal
canopies. Increased ammonium concentrations within macroalgal mats (>25 µM)
also may be toxic to seagrass (van Katwijk et al. 1997), and again, this effect is
likely to be most important for newly recruiting shoots that exist entirely within the
macroalgal canopy. A decrease in redox and an increase in sediment sulfide
concentrations resulting from decomposition in the anoxic, organic-rich sediments
and decaying macroalgal layer also may reduce seagrass photosynthesis (Goodman
et al. 1995; but see Terrados et al. 1999). Decreased photosynthetic oxygen
production at all light levels also decreases the potential for oxygen translocation
and release to the rhizosphere, and creates a positive feedback that reduces sulfide
oxidation around the roots and further elevates sediment sulfide levels, which
decrease nutrient uptake and plant energy status (Pregnall et al. 1984).
Nutrient assimilation and temporary retention is probably the most important
process regulating the source-sink role of benthic communities. In the short term
(days-months), uptake and temporary assimilation influence the fluxes of nutrients
between the sediment and water column. Carbon and nitrogen that are fixed may be
rapidly re-released to the water column as DOC or DON. The DOM released by
benthic autotrophs may be processed by the microbial loop, mineralized, or
transferred to higher trophic levels. Abiotic processes such as humification are also
operative in sediments. In the long term, while most of the biomass of benthic
primary producers is decomposed, grazed, or exported from the system, some is
retained for possible burial (Cebrian 2002; see Chapter 8). Seagrass-dominated
systems have higher rates of permanent burial in the sediments because of the
K.J. McGlathery, K. Sundbäck and I.C. Anderson
inherently greater refractory content of seagrass tissue and the presence of
significant quantities of belowground material (roots and rhizomes) that are less
likely to be exported (Buchsbaum et al. 1991; Enriquez et al. 1993; Duarte et al.
1996; Klap et al. 2000; see Chapter 8). However, decay rates of seagrass leaves can
be high, and similar in magnitude as those for macroalgae (Buchsbaum et al. 1991).
Risgaard-Petersen and Ottosen (2000) found for Z. marina that the pool of tissuebound N did not accumulate in the leaves on an annual time scale. Nutrients that
were temporarily retained in leaf tissue were lost when the leaves were detached and
transported from the systems, or if retained, were nearly completely mineralized
(95-98%) within a year. Similarly, most of the fixed C and N in macroalgae is
released during senescence over a period of several months (Buchsbaum et al.
1991). Benthic microalgae also generally have high growth rates and low C/N
contents; thus a rapid turnover of N in benthic microalgae would be expected.
Calculating assimilation is difficult because benthic primary producers fix more
carbon and nutrients than are needed to meet the metabolic needs for growth and
some of the fixed C and N (and probably P) is leaked from living tissue. For
example, the release of up to 22% of DIN uptake by macroalgae as DON
compounds, means that calculations of macroalgal N assimilation based on growth
rates and tissue C:N content are severe underestimates (Tyler et al. 2003). Similarly,
calculations of benthic microalgal nitrogen demand are subject to potential errors of
either over- or underestimation. Calculations of N demand may be based either on
measurements of net primary production (NPP) (Sundbäck and Miles 2000) or gross
primary production (GPP) (Anderson et al. 2003). If NPP estimates are based on
DO or DIC fluxes measured in sediment cores, sediment community respiration will
result in an underestimation of microalgal NPP and, therefore, N demand. On the
other hand, estimation of N demand based on GPP may result in overestimation
since GPP does not take into account exudation of DOM, which may be a large
component (up to 75%, see below) of the carbon fixed. Use of 13C and 15N tracers
would provide more realistic measures of benthic autotroph N demand; however, a
complete understanding would require both inorganic and organic tracers since
benthic primary producers are known to take up DON (e.g., Nilsson and Sundbäck
1996; Tyler et al. 2003). There have been some attempts to compare the importance
of benthic microalgal and macroalgal assimilation of nutrients in coastal
embayments, in which the assimilation rate of N by benthic microalgae has been
found to be similar to, or even exceed, that of macroalgae (e.g., McGlathery et al.
2001a; Sundbäck and Miles 2002).
Carbon and nitrogen that are not used for growth may be stored in tissue, exuded as
attached extracellular polymeric substances (EPS), or released to the surrounding
water as colloidal organic matter (COM). For example, it has been observed that up
to 75% of carbon fixed by benthic diatoms may be exuded as EPS or COM (Goto et
al. 1999; Smith and Underwood 2000; Wolfstein et al. 2002). Wolfstein et al (2002)
found that the percentage of fixed C excreted appeared to vary with irradiance
(decreased excretion at higher irradiance) and with bacterial activity. In an in situ
pulse – chase 13C labeling experiment Middelburg et al. (2000) noted rapid transfer
of 13C from benthic microalgal into bacterial biomass. They suggest that exudation
of EPS may be a medium for this exchange. There are also reports of DOC leakage
from macroalgal fronds which may be as high as 39% of gross production
(Velimirov 1986). In contrast, there appears to be more limited release of organic
Benthic N and P cycling
compounds from living seagrass tissues. Of the carbon that is fixed during
photosynthesis, only 1-2% is released by the leaves (e.g., Velimirov 1986; Moriarty
et al. 1986) and up to 11% by the roots and rhizomes (Moriarty et al. 1986).
Nitrogen retention in benthic primary producer communities depends in part on the
leakage of DON from live tissues, yet there is considerably less known about this
process than for DOC. To our knowledge no studies have demonstrated exudation
of DON by benthic microalgae, although DON release has been observed for pelagic
cyanobacteria such as Synechococcus spp. (Bronk, 1999) and for the nitrogen fixing
Trichodesmium spp. (Glibert and Bronk, 1994). Rates of DON release have been
observed to vary with light availability and nutrient status, depth of the water
column, DIN species composition and concentration, and season (Ward and Bronk,
2001). The composition of DON released includes DFAA, DCAA, dissolved
primary amines, and urea (Berman and Bronk, in press). DON may be released by
passive release, excretion, sloppy feeding during grazing by copepods or direct
release by microzooplankton, and viral lysis.
Macroalgae also may release
significant quantities of DON as DCAA; Tyler et al. (in press) found benthic fluxes
of DCAA that were nearly 8-fold higher in the presence of macroalgae, and were
higher in the light, suggesting a photosynthetically-driven process (Tyler et al.
2003). We are unaware of any studies on DON release from living seagrasses.
In phytoplankton-based systems, release of DOC or DON can stimulate
water column bacteria and higher trophic levels (Azam et al. 1983). It is likely that
this also occurs in benthic primary producer communities, although it is more
difficult to show this link. In Hog Island Bay, Virginia, highest rates of water
column metabolism of DOC and DON were observed in August, the period when
the macroalgal bloom declines, suggesting that the labile DOM was derived from
macroalgal detritus (Lunsford, 2002). Rates of DOM metabolism in a small
Massachusetts phytoplankton-dominated estuary bounded by both fresh and
saltwater marshes were only 25% of those observed in the Virginia lagoon. The
DOM in the Massachusetts estuary was primarily allochthonous in origin, whereas
that in the Virginia coastal lagoon was mainly autochthonous. Studies such as this
suggest that bacterial activity in the water column can be an important link between
sediment detrital material and higher trophic levels, and may play an important role
in nutrient transformations in shallow coastal systems.
The retention time of tissue-bound nutrients also depends on the
transformations once assimilated, which can take three pathways in the benthic food
web: 1) through the macrofaunal grazing/deposit feeding chain (Asmus and Asmus
1985; Duarte and Cebrian 1996; Herman et al. 2000), 2) through the ‘small food
chain’ consisting of micro- and meiofauna (Kuipers et al. 1981; Sundbäck et al.
1996b and references therein), or 3) burial or export (Admiraal et al. 1984; Boschker
et al. 1999; Middelburg et al. 2000). Since much of the organic matter is released as
DOM by live and dead material, processing of detritus by grazers and bacteria only
accounts for a portion of the plant organic matter. For benthic microalgae, an in situ
C-labelling experiment showed that assimilated C was transferred within a few
days through the heterotrophic components in an order bacteria > macrofauna >
meiofauna (Middelburg et al. 2000). Sundbäck et al. (1996b) found by in situ duallabeling (14C-HCO3, 3H-thymidine; Montagna 1984) of sediment, that meiofauna
grazed between 2 and 12% of the microalgal biomass per day and had a significant
impact on microalgal biomass in spring and autumn, but not in summer.
K.J. McGlathery, K. Sundbäck and I.C. Anderson
The generally accepted conceptual models of eutrophication (Sand-Jensen and
Borum 1991; Duarte 1995; Valiela et al. 1997) predict that with increasing nutrient
loading, benthic primary production will decrease; sediments will become
increasingly heterotrophic, resulting in increased efflux of nutrients from the
sediment to the water column. Ultimately, this will cause a positive feedback
accelerating eutrophication through internal loading of nutrients. Some empirical
evidence supporting this hypothesized series of events has been found (e.g., Rizzo et
al. 1992; Meyercordt and Meyer-Reil 1999; Eyre and Ferguson 2002).
Initially, both seagrasses and benthic microalgae will be stimulated by increased
nutrient levels, particularly in sandy sediments (Nilsson et al. 1991; Flothman and
Werner 1992). However, at the same time the growth of filamentous macroalgae and
epiphytes will be stimulated (e.g., Nilsson et al. 1991; Havens et al. 2001), with
increased shading of the benthos. Systems dominated by fast-growing macroalgae
and phytoplankton are characterized by short-lived (days-months) bloom events and
rapid turnover of nutrients bound in plant biomass. Whether and when planktonic or
macroalgal communities outcompete benthic microalgae and seagrasses by shading
will depend on the duration and extent of bloom events. We hypothesise that, with
proceeding eutrophication and light attenuation from high phytoplankton biomass,
the negative impact on the microphytobenthic community will be more gradual and
slower than for benthic macroscopic primary producers, including seagrasses and
macroalgae. Thus, a partial beneficial “buffering” effect of benthic microalgae on
shallow sediment systems may persist even in more heavily eutrophied systems, as
has been shown by Krom et al. (1991) for a hypertrophic fish pond. This scenario is
based on the assumption that benthic microalgal communities possess, due to high
diversity and functional redundancy, a certain degree of plasticity, increasing the
overall resilience of shallow-water sediment systems after pelagic bloom events.
Benthic diatoms can survive periods of only a few % of incident light, or even
darkness, and high sulphide levels, and can rapidly resume photosynthesis when
exposed to light or after an anoxic event (Admiraal et al. 1984; Kennett and
Hargraves 1985; Sundbäck and Granéli 1988; Sundbäck et al. 1990). Similarly,
upward migrating benthic diatoms were found to rapidly restore the oxygenation of
the sediment surface after a simulated sedimentation event (Wulff et al. 1997;
Underwood and Paterson 1993). This scenario, with benthic microalgae surviving
despite deteriorating conditions, may apply particularly to cool microtidal waters,
where macroalgal bloom events last only a few months (e.g. Pihl et al. 1999;
Dalsgaard in press; McGlathery et al. 2001a), leaving the rest of the year open to
benthic microalgal primary production. In warm, eutrophic microtidal systems,
benthic microalgae can be outcompeted by shading and also by dystrophic events
when the macroalgal blooms eventually collapse (Viaroli et al. 1996).
The efflux of N and P from the sediments will increase as eutrophication proceeds
due to prolonged shading of the benthos, as suggested by experiments at varying
light intensities (Rizzo et al. 1992; Sundbäck and Granéli 1988; Cerco & Setizinger
1997). This will further stimulate pelagic primary productivity, while the sediment
system will become more heterotrophic. Several studies have specifically pointed
Benthic N and P cycling
out that the autotrophy/heterotrophy balance is a good indicator of nutrient flux from
sediments (e.g., McGlathery et al. 2001a; Eyre and Fergusson 2002; Laursen and
Seitzinger 2002; Sundbäck et al. 2003; Tyler et al. 2003), with net autotrophic
systems functioning as sinks and net heterotrophic systems as sources. As systems
become increasingly enriched with organic matter, net heterotrophy will increase
and the systems will become self-generating and demonstrate hysteresis; e.g.,
internal nutrient loading from the sediments will continue to sustain the growth of
ephemeral macroalgal mats for some period of time after nutrient loads begin to
decrease (Pihl et al. 1999; Sundbäck et al. 2003). However, measurements from
some Danish shallow fjords suggest that the release rate of the sediment N pool can
be significantly lower only a few years after the nutrient load to the water column
has been decreased (Dalsgaard et al. 1999).
When nutrient loads are decreased, a reversed scenario has been predicted, where
benthic primary producers are favored because of better light conditions, thus
accelerating the process of recovery (Christensen et al. 1998) by reoxygenation of
the sediment, interception of the nutrient flux, and by increased retention of nutrients
in the sediment. Seagrass-dominated systems will have both a longer retention time
of nutrients temporarily bound in plant material and a greater permanent removal of
nutrients by burial.
In this chapter we have reviewed the effects of seagrass, macroalgal, and benthic
microalgal communities on specific nitrogen and phosphorus cycling processes and
on the ecosystem-level role of these communities as sources or sinks of nutrients.
Photosynthetic and respiratory activities of benthic primary producers alter chemical
conditions in the sediments, causing spatial and diel variations in DO, DIC, pH and
nutrients that influence the rates and pathways of nutrient transformations. Uptake
and retention of nutrients by plant biomass and the effects of primary producers on
N losses from denitrification are key regulators of sediment-water column fluxes.
As such, benthic primary producers play an important role regulating phytoplankton
production in the overlying water column. However, nutrients bound in plant
biomass turnover rapidly and probably make little difference on annual times scales
to the overall nutrient retention within the system. It also appears that although
there are examples of both stimulation and suppression of denitrification rates by
benthic primary producer communities, there is not a consistent and significant
increase of N losses by denitrification on annual time scales caused by benthic
primary producers. This being the case, the key drivers to total estuarine nutrient
retention of external inputs are likely dependent on: 1) the amount of burial of
refractory compounds (including particulates imported to system and trapped as well
as in situ production) and 2) the water residence time (i.e., to what extent recycled
nutrients support primary and secondary production within the system before being
exported). Both of these processes will be dependent on the character of the primary
producer community. Coupling of ecological and physical processes is central to
our understanding of how shallow coastal ecosystems function as a filter of
watershed nutrient inputs, yet we know relatively little about the fate and turnover of
K.J. McGlathery, K. Sundbäck and I.C. Anderson
nutrients bound in plant biomass or the complex hydrodynamics of many shallow
coastal estuaries.
Admiraal W. (1984). The ecology of estuarine sediment-inhabiting diatoms. In: Round, F. E., Chapman,
D. J. (ed.) Progress in phycological research. Biopress Ltd, Bristol, p. 269-314
Alcoverro T., M. Manzanera and J. Romero. (2000). Nutrient mass balance of the seagrass Posidonia
oceanica: the importance of nutrient retranslocation. Marine Ecology Progress Series 194: 13-21
An S., S.B. Joye. (2001). Enhancement of coupled nitrification-denitrification by benthic photosynthesis
in shallow estuarine sediments. Limnology and Oceanography 46: 62-74
Anderson I.C., K.J. McGlathery and A.C. Tyler. (2003). Microbial mediation of ‘reactive’ nitrogen
transformations in a temperate lagoon. Marine Ecology Progress Series 246:73-84.
Anderson I.C., C.R. Tobias, B.B. Neikirk and R.L. Wetzel. (1997). Development of a process-based mass
balance model for a Virginia Spartina alterniflora salt marsh: Implications for net DIN flux. Marine
Ecology Progress Series 159, 13-27.
Asmus H. and R. Asmus. (1985). The importance of grazing food chain for energy flow and production in
three intertidal sand bottom communities of the northern Wadden Sea. Helgoländer
Meeresuntersuchungen 39: 273-301
Astill H. and P. Lavery. (2001). The dynamics of unattached benthic macroalgal accumulations in the
Swan-Canning Estuary. Hydrological Processes 15: 2387-2399
Azam R., T. Fenchel, J.G. Field, J.S. Gray, L.A. Meyerriel and F. Thingstad. (1983). The ecological role
of water column microbes in the sea. Marine Ecology Progress Series 10: 257-263.
Berman T. and D.A. Bronk. (in press). Dissolved organic nitrogen: a dynamic participant in aquatic
ecosystems. Aquatic Microbial Ecology
Blaaberg V.K., K.M. Mouritsen and K. Finster. (1998). Diel cycles of sulphate reduction rates in
sediments of a Zostera marina bed, Denmark. Aquatic Microbial Ecology 15:97-102
Blackburn T.H., D.B. Nedwell and W.J. Wiebe. (1994). Active mineral cycling in a Jamaica seagrass
sediment Marine Ecology Progress Series 110: 233-239
Blanchard G.F. and P.A. Montagna. (1995). Assessment of a brown tide impact on microalgal benthic
communities in Baffin Bay (Texas) in 1990 using a primary production simulation model.
Oceanologia Acta 18: 371-377
Borum J. (1996). Shallow waters and Land/Sea Boundaries. In Eutrophication in Coastal Marine
Ecosystem, Jørgensen, B. B. and K. Richardson, eds. American Geophysical Union, Washington, D.
C., pp. 179-204.
Borum, J. and K. Sand-Jensen. (1996). Is total primary production in shallow coastal matine waters
stimulated by nitrogen loading? Oikos 76:406-410
Boschker H.T.S., A. Wielemaker, B.E.M. Schaub, and M. Holmer. (2000). Limited coupling of
macrophyte production and bacterial carbon cycling in the sediments of Zostera spp. meadows.
Marine Ecology Progress Series 203:181-189.
Boynton W.R., J.D. Hagy, L. Murray, C. Stokes, and W.M. Kemp. (1996). A comparative analysis of
eutrophication patterns in a temperate coastal lagoon. Estuaries 19: 408-421.
Benthic N and P cycling
Buchsbaum R., I. Valiela, T. Swain, M. Dzierzeski and S. Allen. (1991). Available and refractory
nitrogen in detritus of coastal vascular plants. Marine Ecology Progress Series 72: 131-143.
Burdige D.J. and R.C. Zimmerman. (2002). Impact of seagrass density on carbonate dissolution in
Bahamian sediments. Limnology and Oceanography 47L 1761-1763.
Bronk D.A. (1999). Rates of NH4+ uptake, intracellular transformation and dissolved organic nitrogen
release in two clones of marine Synechococcus spp. Journal of Plankton Research 21:1337-1353.
Burdige D.J., S. Zheng. (1998). The biogeochemical cycling of dissolved organic nitrogen in estuarine
sediments. Limnology and Oceanography 43:1796-1813
Cabrita M.T. and V. Brotas. (2000). Seasonal variation in denitrification and dissolved nitrogen fluxes in
intertidal sediments of the Tagus Estuary. Marine Ecology Progress Series 202: 51-65
Caffrey J.M. and W.M. Kemp. (1992). Influence of the submersed plant Potamogeton perfuliatus, on
nitrogen cycling in estuarine sediments. Limnology and Oceanography 37:1483-1495
Cambridge M.L. and A.J. McComb. (1984). The loss of seagrasses in Cockburn Sound, Western
Australia. 1. The time course and magnitude of seagrass decline in relation to industrial development.
Aquatic Botany 20: 229-243
Capone D.G. (1988). Benthic nitrogen fixation: microbiology, physiology and ecology. In: Blackburn
TH, J Sørensen and T Roswall (eds). Nitrogen cycling in marine coastal environments. SCOPE
Series. J Wiley & Sons, New York, p 85-123
Capone D.G. and B.F. Taylor. (1977). Nitrogen fixation (acetylene reduction) in the phyllosphere of
Thalassia testudinum. Marine Biology 40: 19-28.
Capone D.G., S.E. Dunham, S.G. Horrigan and L.E. Duguay. (1992). Microbial nitrogen transformations
in unconsolidated coral reef sediments. Marine Ecology Progress Series 80: 75-88
Carlson P.R., L. Yarbro and T. Barber. (1994). Relationship of sediment sulfide to mortality of Thalassia
testudinum in Florida Bay. Bulletin of Marine Science 54: 733-746
Cebrian J. (2002). Variability and control of carbon consumption, export and accumulation in marine
communities. Limnology and Oceanography 47:11-22
Cerco C.F. and S.P. Seitzinger. (1997). Measured and modeled effects of benthic algae on eutrophication
in Indian River-Rehoboth Bay, Delaware. Estuaries 20: 231-248
Chambers R.M., J.W. Fourqurean, S.A. Macko and R. Hoppenot. (2001). Biogeochemical effects of iron
availability on primary producers in a shallow marine carbonate environment. Limnology and
Oceanography 46: 1278-1286
Christensen P.B. et al. (1998). The Danish marine environment: Has action improved its state. Danish
Environmental Protection Agency. Report 62:
Cornwell J.C., W.M. Kemp and T.M. Kana. (1999). Denitrification in coastal ecosystems: Environmental
controls and aspects of spatial and temporal scaling. Journal of Aquatic Ecology 33: 41-54
Currin C.A., S.B. Joye and H.W. Paerl. (1996). Diel rates of N2-fixation and denitrification in a
transplanted Spartina alterniflora marsh: implications for N-flux-dynamics. Estuarine Coastal and
Shelf Science 42: 597-616
Dalsgaard T., P.B. Christensen, S. Rysgaard and N. Risgaard-Petersen. (1999). Nitrogen removal in
Danish waters: importance and regulation (In Danish). In: A., L. B. (ed.) Havmiljøet ved
årtusindskiftet ( Marine Environments into the Millenium). Vol: Olsen & Olsen, Fredensborg, p. 101118
K.J. McGlathery, K. Sundbäck and I.C. Anderson
D’Avanzo C. and J.N. Kremer. (1994). Diel oxygen dynamics and anoxic events in an eutrophic estuary
of Waquoit Bay, Massachusetts. Estuaries 17: 131-39
Duarte C.M. (1995). Submerged aquatic vegetation in relation to different nutrient regimes. Ophelia 41:
Duarte C.M. and J. Cebrián. (1996). The fate of marine autotrophic production. Limnology and
Oceanography 41: 1758-1766
Dudley B.J., A.M.E. Gahnstrom and D.I. Walker. (2001). The role of benthic vegetation as a sink for
elevated inputs of ammonium and nitrate in a mesotrophic estuary. Marine Ecology Progress Series
219: 99-107.
Enriquez S., C.M. Duarte and K. Sand-Jensen. (1993). Patterns in decomposition rates among
photosynthetic organisms: The importance of detritus C:N: P content. Oecologia 94: 457-471.
Erskine J.M. and M.S. Koch. (2000). Sulfide effects on Thalassia testudinum carbon balance and
adenylate energy charge. Aquatic Botany 67: 275-285
Eyre B.D. and A.J.P. Ferguson. (2002). Comparison of carbon production and decompositopn, benthic
nutrient fluxes and denitrification i seagrass, phytoplankton, benthic microalgae- and macroalgaedominated warm-temperate Australian lagoons. Marine Ecology Progress Series 229: 43-59
Falcao M. and C. Vale. (1995). Tidal flushing of ammonium from intertidal sediments of Ria Formosa,
Portugal. Netherlands Journal of Aquatic Ecology 29: 239-244
Flindt M.R. (1994). Measurements of nutrient fluxes and mass balances by on-line in situ dialysis in a
Zostera marina bed culture. Verhandlungen Internationale Vereinigung Limnologie 25:2259-2264
Flothmann S. and I. Werner. (1992). Experimental eutrophication on an intertidal sandflat: effects on
microphytobenthos, meio- and macrofauna. In: Colombo G, I Ferrari, V Ceccherelli, and R Rossi
(ed.) Marine eutrophication and population dynamics. Vol: Olsen & Olsen, Fredensborg, p. 93-100
Fong P., R.M. Donohoe and J.B. Zedler. (1993). Competition with macroalgae and benthic cyanobacterial
mats limits phytoplankton abundance in experimental microcosms. Marine Ecology Progress Series
100: 97-102
Gacia E., C.M. Duarte and J.J. Middelburg. (2002). Carbon and nutrient deposition in a Mediterranean
seagrass (Posidonia oceanica) meadow. Limnology and Oceanography 47,: 23-32.
Giblin A. and R.W. Howarth. (1984). Pore water evidence for a dynamic sedimentary iron cycle in salt
marshes. Limnology and Oceanography 29: 47-63.
Giblin A.E., C.S. Hopkinson Jr and J. Tucker. (1997). Benthic metabolism and nutrient cycling in Boston
Harbor, Massachusetts. Estuaries 20:346-364
Glibert P.M. and D.A. Bronk. (1994). Release of dissolved organic nitrogen by marine diazotrophic
cyanobacteria, Trichodesmium spp. Applied and Environmental Microbiology 60: 3996-4000.
Goodman J.L., K.A. Moore and W.C. Dennison. (1995). Photosynthetic responses of eelgrass (Zostera
marina L.) to light and sediment sulfide in a shallow barrier island lagoon. Aquatic Botany 50: 37-47
Goto N., T. Kawamura, O. Mitamura and H. Terai. (1999). Importance of extracellular organic carbon
production in the total primary production by tidal-flat diatoms in comparison to phytoplankton.
Marine Ecology Progress Series 190:289-295.
Hansen K. and E. Kristensen. (1997). Impact of macrofaunal recolonization on benthic metabolism and
nutrient fluxes in a shallow marine sediment previously overgrown with macroalgal mats. Estuarine
Coastal and Shelf Science 45:613-628.
Benthic N and P cycling
Hansen J.W., J.W. Udy, C.J. Perry, W.C. Dennison and B.A. Lomstein. (2000). Effect of the seagrass
Zostera capricorni on sediment microbial processes. Marine Ecology Progress Series 199: 83-96
Hauxwell J., J. Cebrian, C. Furlong and I. Valiela. (2000). Macroalgal canopies contribute to eelgrass
(Zostera marina) decline in temperate estuarine ecosystems. Ecology 82: 1007-1022
Havens K.E., J. Hauxwell, A.C. Tyler, S. Thomas, K.J. McGlathery, J. Cebrian, I. Valiela, A.D. Steinman
and S.-J. Hwang. (2001). Complex interactions between autotrophs in shallow marine and freshwater
ecosystems: implications for community responses to nutrient stress. Environmental Pollution 113:
Hemminga, M.A., P.G. Harrison and F. van Lent. (1991). The balance of nutrient losses and gains in
seagrass meadows. Marine Ecology Progress Series 71:85-96
Henriksen K., J. Hansen and T.H. Blackburn. (1980). The influence of benthic infauna on exchange rates
of inorganic nitrogen between sediment and water. Ophelia Suppl. 1: 249-256
Henriksen K. and M.W. Kemp. (1988). Nitrification in estuarine and coastal marine sediments 201-255.
In: Blackburn TH and J Sørensen [eds.], Nitrogen cycling in coastal marine environments, SCOPE.
John Wiley & Sons Ltd
Herman P.M.J., M.J. Widdows, C.H. Licas and C.H.R. Heip. (2000). Stable isotopes as trophic tracers:
combining field sampling and manipulative labelling of food resources for macrobenthos. Marine
Ecology Progress Series 204: 79-92
Hines M. E. and W.B. Lyons. (1982). Biogeochemistry of nearshore Bermuda sediments. I. Sulfate
reduction rates and nutrient generation. Marine Ecology Progress Series 8: 87-94
Hoffland E. (1992). Quantitative evaluation of the role of organic acid exudation in the mobilization of
rock phosphate by rape. Plant Soil 140: 279-289.
Hopkinson C.S., A.E. Giblin and J. Tucker. (2001). Benthic metabolism and nutrient regeneration on the
continental shelf of Eastern Massachusetts, USA. Marine Ecology Progress Series 224:1-19.
Hopkinson C.S., A.E. Giblin, J. Tucker and R.H. Garritt. (1999). Benthic metabolism and nutrient cycling
along an estuarine salinity gradient. Estuaries 22:863-881
Howarth R.W., R. Marino, J. Lane and J.J. Cole. (1988). Nitrogen fixation in freshwater, estuarine, and
marine ecosystems. 1. Rates and importance. Limnology and Oceanography 33: 669-687
Jensen H.S., K.J. McGlathery, R. Marino and R.W. Howarth. (1998). Forms and availability of sediment
phosphorus in carbonate sand of Bermuda seagrass beds. Limnology and Oceanography 43: 799-810
Kemp W.M. and J.C. Cornwell. (2001). Role of benthic communities in the cycling and balance of
nitrogen in Florida Bay. Final report to the U.S. Environmental Protection Agency, Region 4,
Atlanta, GA. 53 pp.
Kennett D.M. and P.E. Hargraves. (1985). Benthic diatoms and sulfide fluctuations: upper basin of
Pettaquamscutt River, Rhode Island. Estuarine Coastal and Shelf Science 21: 577-586
Klap V.A. M.A. Hemminga and J.J. Boon. (2000). Retention of lignin in seagrasses: angiosperms that
returned to the sea. Marine Ecology Progress Series 194:1-11.
Knight W.G., L.M. Dudley and J.J. Jurinak. (1992). Oxalate effects on solution phosphorus in a
calcareous soil. Arid Soil Research and Rehabilitation 6: 11-20.
Koepfler E.T., R. Benner, and P.A. Montagna. (1993). Variability of dissolved organic carbon in
sediments of a seagrass bed and an unvegetated area within an estuary in southern Texas. Estuaries
16: 391-404
K.J. McGlathery, K. Sundbäck and I.C. Anderson
Krause-Jensen D., K.J. McGlathery, S. Rysgaard and P.B. Christensen. (1996). Production within dense
mats of the filamentous macroalga Chaetomorpha linum in relation to light and nutrient availability.
Marine Ecology Progress Series 134: 207-216
Krause-Jensen D., P.B. Christensen and S. Rysgaard. (1999). Oxygen and nutrient dynamics within mats
of the filamentous macroalga Chaeotomorpha linum. Estuaries 22: 31-38.
Kristensen E., F.O. Andersen, N. Holmboe, M. Holmer and N. Thongtham. (2000). Carbon and nitrogen
mineralization in sediments of the Bangrong mangrove area, Phuket, Thailand. Aquatic Microbial
Ecology 22: 199-213
Krom M.D. (1991). Importance of benthic productivity in controlling the flux of dissolved inorganic
nitrogen through the sediment-water interface in a hypertrophic marine ecosystem. Marine Ecology
Progress Series 78: 163-172
Kuipers B.R., P.A.W.J. de Wilde and F. Creutzberg. (1981). Energy flow in a tidal flat ecosystem.
Marine Ecology Progress Series 5: 215-221
Laursen A.E. and S.P. Seitzinger. (2002). Measurement of denitrification in rivers: an integrated, whole
reach approach. Hydrobiologia 485: 67-81.
Lomstein B.A., A.-G.U. Jensen, J.W. Hansen, I.B. Andreasen, L.S. Hansen, J. Berntsen and H.
Kunzendorf. (1998). Budgets of sediment nitrogen and carbon cycling in the shallow water of Knebel
Vig, Denmark. Aquatic Microbial Ecology 14: 69-80
Lorenzen J., L.H. Larsen, T. Kjaer and N.P. Revsbech. (1998). Biosensor determination of the microscale
distribution of nitrate, nitrate assimilation, nitrification, and denitrification in a diatom-inhabited
freshwater sediment. Applied and Environmental Microbiology 64: 3264-3269
Lunsford T. (2002). A comparison of the fate of dissolved organic matter in two coastal systems: Hog
Island Bay, VA (USA) and Plum Island Sound, MA (USA). Ph.D. Thesis, Virginia Institute of
Marine Science, College of William and Mary.
McComb A.J. and R. Humphries. (1992). Loss of nutrients from catchments and their ecological impacts
on the Peel-harvey estuarine system, Western Australia. Estuaries 15: 529-537.
McGlathery K.J., D. Krause-Jensen, S. Rysgaard and P.B. Christensen. (1997). Patterns of ammonium
uptake within dense mats of the filamentous macroalga Chaetomorpha linum. Aquatic Botany 59:
McGlathery K.J., N. Risgaard-Petersen and P.B. Christensen. (1998). Temporal and spatial variation in
nitrogen fixation activity in the eelgrass (Zostera marina L.) rhizosphere. Marine Ecology Progress
Series 168: 245-258
McGlathery K.J., I.C. Anderson and A.C. Tyler. (2001a). Magnitude and variability of benthic and
pelagic metabolism in a temperate coastal lagoon. Marine Ecology Progress Series 216: 1-15
McGlathery K.J., P. Berg and R Marino. (2001b). Using porewater profiles to assess nutrient availability
in seagrass-vegetated carbonate sediments. Biogeochemistry 56: 239-263
Meyercordt J. and L.A. Meyer-Reil. (1999). Primary production of benthic microalgae to two shallow
coastal lagoons of different trophic status in the southern Baltic Sea. Marine Ecology Progress Series
178: 179-191
Middelburg J.J., C. Barranguet, H.T.S. Boschker, P.M.J. Herman, T. Moens and C.H.R. Heip. (2000).
The fate of intertidal microphytobenthos carbon: an in situ 13C-labelling study. Limnology and
Oceanography 45: 1224-1234
Benthic N and P cycling
Miyajima T., M. Suzumura, Y. Umezawa and I. Koike. (2001). Microbial nitrogen transformation in
carbonate sediments of a coral-reef lagoon and associated seagrass beds. Marine Ecology Progress
Series 217: 273-286
Moncreiff C.A., M.J. Sullivan, and A.E. Daehnick. (1992). Primary production in seagrass beds of
Mississippi Sound: the contributions of seagrass, epiphytic algae, sand microflora, and
phytoplankton. Marine Ecology Progress Series 87: 161-171
Montagna P.A.. (1984). In situ measurement of meiobenthic grazing rates on sediment bacteria and
edaphic diatoms. Marine Ecology Progress Series 18: 119-130
Moriarty D.W., J.R. Iverson and P.C. Pollard. (1986). Exudation of organic carbon by the seagrass
Halodule wrightii and its effect on bacterial growth in the sediment. Journal of Experimental
Biology and Ecology 96: 115-126
Moriarty D.W. and M.J. O’Donohue. (1993). Nitrogen fixation in seagrass communities during summer
in the Gulf of Carpentaria, Australia. Australian Journal of Marine and Freshwater Research 44:
Murray L. and R.L. Wetzel. (1987). Oxygen production and consumption associated with the major
autotrophic components in two temperate seagrass communities. Marine Ecology Progress Series 38:
Nielsen L.P.. (1992). Denitrification in sediment determined from nitrogen isotope pairing. FEMS
Microbiology Ecology. 86: 357-362
Nilsson C. and K. Sundbäck. (1996). Amino acid uptake in natural microphytobenthic assemblages
studied by microautoradiography. Hydrobiologia 332: 119-129.
Nilsson P., B. Jönsson, Lindström, I. Swanberg and K. Sundbäck. (1991). Response of a marine shallowwater sediment system to an increased load of inorganic nutrients. Marine Ecology Progress Series
71: 275-290
Nixon S.W. et al. (1996). The fate of nitrogen and phosphorus at the land-sea margin of the North
Atlantic Ocean. Biogeochemistry 35: 141-180.
National Research Council. (2000). Clean coastal waters: understanding and reducing the effects of
nutrient pollution. National Academy Press, Washington, DC.
O’Donohue M.J., D.J. Moriarty and I.C. MacRae. (1991). Nitrogen fixation in sediments and the
rhizosphere of the seagrass Zostera capricorni. Microbial Ecology 22: 53-64
O’Neil J.M. and D.G. Capone. (1989). Nitrogenase activity in tropical carbonate marine sediments.
Marine Ecology Progress Series 56: 145-156
Ottosen, L.D.M., N. Risgaard-Petersen and L.P. Nielsen. (1999). Direct and indirect measurements of
nitrification and denitrification in the rhizosphere of aquatic macrophytes. Aquatic Microbial Ecology
Pedersen M.F. and J. Borum. (1996). Nutrient control of algal growth in estuarine waters. Nutrient
limitation and the importance of nitrogen requirements and nitrogen storage among phytoplankton
and species of macroalgae. Marine Ecology Progress Series 142: 261-272
Pedersen M.F., C.M. Duarte and J. Cebrian. (1997). Rates of changes in organic matter and nutrient
stocks during seagrass Cymodocea nodosa colonization and stand development. Marine Ecology
Progress Series 159: 29-36.
Pelegri S.P., L.P. Nielsen and T.H. Blackburn. (1994). Denitrification in estuarine sediment stimulated by
the irrigation activity of the amphipod Corophium volutator. Marine Ecology Progress Series
K.J. McGlathery, K. Sundbäck and I.C. Anderson
Perez M., M.A. Mateo, T. Alcoverro and J. Romero. (1997). Variability in detritus stocks in beds of the
seagrass Cymodocea nodosa. Botanica Marina 44: 523-531.
Pihl L., A. Svenson, P.O. Moksnes and H. Wennhage. (1999). Distribution of green algal mats throughout
shallow soft bottoms of the Swedish Skagerrak archipelago in relation to nutrient sources and wave
exposure. Journal of Sea Research 41: 281-294
Pollard P.C. and K. Kogure. (1993). The role of epiphytic and epibenthic algal productivity in a tropical
seagrass, Syringodium isoetifolium (Aschers) Dandy, community. Australian Journal of Marine and
Freshwater Research 44: 141-154
Pregnall A.M., R.D. Smith, T.A. Kursar and R.S. Alberte. (1984). Metabolic adaptations of Zostera
marina (eelgrass) to diurnal periods of root anoxia. Marine Biology 83: 141-147
Reay W.G., D.L. Gallagher and G.M.J. Simmons. (1995). Sediment-water column oxygen and nutrient
fluxes in nearshore environments of the lower Delmarva Peninsula, USA. Marine Ecology Progress
Series 118: 215-227
Risgaard-Petersen N. (2003). Coupled nitrification-denitrification in autotrophic and heterotrophic
estuarine sediments: On the influence of benthic microalgae. Limnology and Oceanography 48: 93105
Risgaard-Petersen N., S. Rysgaard, L.P. Nielsen and N.P. Revsbech. (1994). Diurnal variation of
denitrification and nitrification in sediments colonized by benthic microphytes. Limnology and
Oceanography 39: 573-579.
Risgaard-Petersen N., T. Dalsgaard, S. Rysgaard, P.B. Christensen, J. Borum, K.J. McGlathery and L.P.
Nielsen. (1998). Nitrogen balance of a temperate eelgrass Zostera marina bed. Marine Ecology
Progress Series 174: 281-291
Risgaard-Petersen N. and L.D.M. Ottosen. (2000). Nitrogen cycling in two temperate Zostera marina
beds: seasonal variation. Marine Ecology-Progress Series 198:93-107
Rizzo W.M., S.K. Dailey, G.J. Lackey, R.R. Christian, B.B. Berry, and R.L. Wetzel. (1996). A
metabolism-based trophic index for comparing the ecological values of shallow-water sediment
habitats. Estuaries. 19: 247-256
Rizzo W.M., G.L. Lackey, and R.R. Christian. (1992). Significance of euphotic, subtidal sediments to
oxygen and nutrient cycling in a temperate estuary. Marine Ecology Progress Series 86: 51-61
Rozan T.F., M. Taillefert, R.E. Trouwborst, B.T. Glazer, S. Ma, J. Herszage, L.M. Valdes, K.S. Price and
G.W. Luther III (2002). Iron-sulfur-phosphorus cycling in the sediments of a shallow coastal bay:
Implications for sediment nutrient release and benthic macroalgal blooms. Limnology and
Oceanography 47: 1346-1354.
Rysgaard S., P.B. Christensen and L.P. Nielsen. (1995). Seasonal variation in nitrification and
denitrification in estuarine sediment colonized by benthic microalgae and bioturbating infauna.
Marine Ecology Progress Series 126: 111-121
Rysgaard S., N. Risgaard-Petersen and N.P. Sloth. (1996). Nitrification, denitrification, and nitrate
ammonification in sediments of two coastal lagoons in Southern France. Hydrobiologia 329:133-141
Sand-Jensen, K. and Borum J. (1991). Interactions among phytoplankton, periphyton, and macrophytes
in temperate freshwaters and estuaries. Aquatic Botany 41: 137-175.
Sfriso A., B. Pavoni, A. Marcomini and A.A. Orio. (1992). Macroalgae, nutrient cycles, and pollutants in
the lagoon of Venice. Estuaries 15: 517-528
Benthic N and P cycling
Sloth N.P., H. Blackburn, L.S. Hansen, N. Risgaard-Petersen and B.A. Lomstein. (1995). Nitrogen
cycling in sediments with different organic loading. Marine Ecology Progress Series 116:163-170
Smith D.J. and G.J.C. Underwood. (2000). Exopolymer production by intertidal epipelic diatoms.
Limnology and Oceanography 43:1578-1591.
Stapel J. and M.A. Hemminga. (1997). Nutrient resorption from seagrass leaves. Marine Biology 128:
Stimson J. and S.T. Larned. (2000). Nitrogen efflux from the sediments of a subtropical bay and the
potential contribution to macroalgal nutrient requirements. Journal of Experimental Marine Biology
and Ecology 252: 159-180
Sundbäck K. and W. Granéli. (1988). Influence of microphytobenthos on the nutrient flux between
sediment and water: a laboratory study. Marine Ecology Progress Series 43: 63-69
Sundbäck K., B. Jönsson, P. Nilsson and I. Lindström. (1990). Impact of accumulating drifting
macroalgae on a shallow-water sediment system: an experimental study. Marine Ecology Progress
Series 58: 261-274
Sundbäck K., L. Carlson, C. Nilsson, B. Jönsson, A. Wulff and S. Odmark. (1996a). Response of benthic
microbial mats to drifting green algal mats. Aquatic Microbial Ecology 10: 195-208
Sundbäck K., P. Nilsson, C. Nilsson and B. Jönsson (1996b). Balance between autotrophic and
heterotrophic components and processes in the sandy sediments: a field study. Estuarine Coastal
and Shelf Science 43: 689-706
Sundbäck K. and A. Miles. (2000). Balance between denitrification and microalgal incorporation of
nitrogen in microbial sediments, NE Kattegat. Aquaict Microbial Ecology 22: 291-300
Sundbäck K. and A. Miles. (2002). Role of microphytobenthos and denitrification for nutrient turnover in
embayments with floating macroalgal mats: a spring situation. Aquat Mar Ecol. 30: 91-101
Sundbäck K., A. Miles and E. Göransson. (2000). Nitrogen fluxes, denitrification and the role of
microphytobenthos in microtidal shallow-water sediments: an annual study. Marine Ecology
Progress Series 200: 59-76
Sundbäck K., A. Miles, A. Hulth, L. Pihl, P. Engström, E. Selander and A. Svenson. (2003). Importance
of benthic nutrient regeneration during initiation of macroalgal blooms in shallow bays. Marine
Ecology Progress Series 246: 115-126
Suzumura M., T. Miyajima, H. Hata, Y. Umezawa, H. Kayanne and I. Koike. (2002). Cycling of
phosphorus maintains the production of microphytobenthic communities in carbonate sediments of a
coral reef. Limnology and Oceanography 47: 771-781
Terrados J., C.M. Duarte, L. Kamp-Nielsen, N.S.R. Agawin, E. Gacia, D. Lancap, M.D. Fortes, J. Borum,
M. Lubanski and T. Greve. (1999). Are seagrass growth and survival constrained by reducing
conditions of the sediment? Aquatic Botany 65: 175-107
Thornton D.C.O., L.F. Dong, G.J.C. Underwood and D.B. Nedwell. (2002). Factors affecting
microphytobenthic biomass, species composition and production in the Colne Estuary (UK). Aquatic
Microbial Ecology 27: 285-300
Thornton D.C.O., G.J.C. Underwood and D.B. Nedwell. (1999). Effect of illumination and emersion
period on the exchange of ammonium across the estuarine sediment-water interface. Marine Ecology
Progress Series 28: 11-20
Thybo-Christesen M., M.B. Rasmussen and T.H. Blackburn. (1993). Nutrient fluxes and growth of
Cladophora sericea in a shallow Danish bay. Marine Ecology Progress Series 100: 273-81
K.J. McGlathery, K. Sundbäck and I.C. Anderson
Trimmer M., D.B. Nedwell, D.B. Sivyer and S.J. Malcolm. (2000). Seasonal benthic organic matter
mineralisation measured by oxygen uptake and denitrification along a transect of the inner and outer
River Thames estuary, UK. Marine Ecology Progress Series 197:103-119
Turpin D.H.. (1991). Effects of inorganic N availability on algal photosynthesis and carbon metabolism.
Journal of Phycology 27: 14-20
Tyler A.C., K.J. McGlathery and I.C. Anderson. (2001). Macroalgal mediation of dissolved organic
nitrogen fluxes in a temperate coastal lagoon. Estuarine Coastal and Shelf Science 53: 155-168
Tyler A.C., K.J. McGlathery and I.C. Anderson. (2003). Benthic algae control sediment-water column
fluxes of nitrogen in a temperate lagoon. Limnology and Oceanography In press.
Underwood G.J.C. and J. Kromkamp. (1999). Primary production by phytoplankton and
microphytobenthos in estuaries. Advances in Ecological Research 29: 93-153
Underwood G.J.C. and D.M. Paterson. (1993). Recovery of intertidal benthic diatoms after biocide
treatment and associated sediment dynamics. Journal of the Marine Biology Association of the
United Kingdom 73: 25-45
Valiela I., K. Foreman, M. LaMontagne, D. Hersh, J. Costa, P. Peckol, B. DeMeo-Andreson, C.
D’Avanzo, M. Babione, S. Sham, J. Brawley and K. Lajtha. (1992). Couplings of watersheds and
coastal waters: sources and consequences of nutrient enrichment in Waquoit Bay, Massachusetts.
Estuaries 15: 443-457
Valiela I., J. McClelland, J. Hauxwell, P.J. Behr, D. Hersh and K. Foreman. (1997). Macroalgal blooms
in shallow estuaries: Controls and ecophysiological and ecosystem consequences. Limnology and
Oceanography 42: 1105-1118
Velimerov B. (1986). DOC dynamics in a Mediterranean seagrass system. Marine Ecology Progress
Series 28: 21-41.
van Katwijk M.M., L.H.T. Vergeer, G.H.W. Schmitz and J.G.M. Roelofs. (1997). Ammonium toxicity in
eelgrass Zostera marina. Marine Ecology Progress Series 157: 159-173
Viaroli P. et al. (1992). Ulva rigida growth and decomposition processes and related effects on nitrogen
and phosphorus cycles in a coastal lagoon (Sacca di Goro, Po River Delta), In: G. Colombo, et al.
(ed.), Marine eutrophication and population dynamics, 77-84. Olsen and Olsen, Fredensborg.
Viaroli P., M. Bartoli, C. Bondavalli and R.R. Christian. (1996). Macrophyte communities and their
impact on benthic fluxes of oxygen, sulphide and nutrients in shallow eutrophic environments.
Hydrobiologia. 329: 105-119
Ward B.B. and D.A. Bronk. (2001). Net nitrogen uptake and DON release in surface waters: importance
of trophic interactions implied from size fractionation experiments. Marine Ecology Progress Series
Welsh D.T., S. Bourgues, R. de Wit and R.A. Herbert. (1996). Seasonal variations in nitrogen-fixation
(acetylene reduction) and sulphate-reduction rates in the rhizosphere of Zostera noltii: nitrogen
fixation by sulfate-reducing bacteria. Marine Biology 125: 619-628
Welsh D.T. (2000). Nitrogen fixation in seagrass meadows: Regulation, plant-bacteria interactions and
significance to primary productivity. Ecology Letters 3:58-71
Welsh D.T., M. Bartoli, D. Nizzoli, G. Castaldelli, S.A. Riou and P. Viaroli. (2000). Denitrification,
nitrogen fixation, community productivity and inorganic-N and oxygen fluxes in an intertidal Zostera
noltii meadow. Marine Ecology Progress Series 208: 65-77.
Benthic N and P cycling
Wolfstein K., J.F.C. de Brouwer and L.J. Stal. (2002). Biochemical partitioning of photosynthetically
fixed carbon by benthhic diatoms during short-term incubations at different irradiances. Marine
Ecology Progress Series 245:21-31.
Wulff A., K. Sundbäck, C. Nilsson, L. Carlson and B. Jonsson. (1997). Effect of sediment load on the
microbenthic community of a shallow-water sandy sediment. Estuaries 20: 547-558
Ziegler S. and R. Benner. (1999). Dissolved organic carbon cycling in a subtropical seagrass-dominated
lagoon. Marine Ecology Progress Series 180: 149-160
K.J. McGlathery: Department of Environmental Sciences, P.O. Box 400123,
University of Virginia, Charlottesville, VI 22904, USA.
K. Sundbäck: Gothenburg University, Department of Marine Ecology, P.O. Box
461, S-40530 Gothenburg, Sweden.
I.C. Anderson: School of Marine Science, Virginia Institute of Marine Science,
College of William and Mary, Gloucester Point, VI 23602, USA.
As nitrogen passes through the estuarine environment it undergoes a complex series
of transformations. Inorganic N is incorporated into pelagic primary producers, and
a large fraction reaches the seafloor either through sedimentation of algal cells or via
grazing followed by deposition of faecal pellets and dead organisms. At the seafloor
the organic bound nitrogen is either buried permanently or undergoes decomposition
resulting in release of NH4+ or urea. Alternatively, inorganic N is taken up directly
by the sediment system and incorporated into benthic primary producers or used as
substrate in microbial metabolic processes.
Remineralized nitrogen may be liberated from the sediment to the water column,
often through a complex series of microbiologically mediated metabolic processes
that determine the form in which N is liberated and consequently the degree to
which remineralized N is made available for new primary production.
Sediment nitrification and denitrification are central microbiological processes in
this context. Nitrification is the aerobic conversion of NH4+ to NO3-, the most
oxidized form of nitrogen. The process is mediated by bacteria of the family
Nitrobacteraceae that grow chemolitho-autotrophically, meaning that they derive
energy from the oxidation of inorganic compounds (N) and use CO2 as sole carbon
source. The oxidation of NH4+ to NO3- proceeds in two steps and is carried out by
two types of organisms that make use of different metabolic pathways: The
oxidation of NH4+ to NO3- is performed by ammonium oxidizing bacteria while the
oxidation of NO2- to NO3- is performed by nitrite oxidizing bacteria.
Denitrification is the anaerobic conversion of oxidized nitrogen (NO3- or NO2-) to
reduced gaseous compounds (N2O or N2). The process is mediated by facultative
aerobes capable of utilizing oxidized N as terminal electron acceptor in a
dissimilatory process when O2 becomes limiting. The ability to perform this process
is found in a variety of taxonomic groups including litotrophic and phototrophic
bacteria. Most denitrifiers grow organotrophically, however (Zumft 1992), meaning
that they are able to derive energy from the oxidation of reduced organic carbon.
Denitrification is an important factor determining the N-availability in estuarine
environments, since only a limited number of marine organisms can utilize the end
product N2 as an N-source (Howarth, et al. 1988).
S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling:
The IInfluence of Primary Producers, 263-280.
© 2004 Kluwer Academic Publishers. Printed in the Netherlands.
N. Risgaard-Petersen
Table 1. Recent measurements of denitrification rates in European estuaries.
Rörtungen , Sweden
Denitrification rate
(mmol N m-2 d-1)
Randers Fjord, Denmark
Targus Estuary, Portugal
Løgstør Bredning, Denmark 0.08-0.26
River Thames Estuary, UK
Bassin d’Arcachon, France
et al. 2000
et al. 2001
Cabrita and Brotas
et al. 1998
et al. 2000
et al. 2000
Rates of nitrification and denitrification have been measured intensively in European
estuaries during the last decades. Sediment denitrification and nitrification are
regulated by a complex matrix of biological, physical and chemical factors, and
actual in situ activity may vary orders of magnitude (Table 1) depending on season,
nutrient load and the composition of the ecosystem. In the following section major
factors regulating the processes will be discussed with special reference to responses
to organic loads, nutrient concentrations and benthic primary production.
Basically the processes of nitrification and denitrification are located in the upper few
mm of the sediment. Figure 1 shows examples of porewater profiles of NO3-, NH4+
and O2 in unvegetated sediment as measured with microsensors. On the basis of these
profiles the distribution of nitrification and denitrification can be estimated (Berg et
al. 1998).
Being a strictly aerobic process, nitrification takes place in the oxic zone, as indicated
in the figure, and the substrate NH4+ is supplied mainly by mineralization processes in
the anoxic sediment strata while O2 is supplied by the water column. Denitrification
takes place just below the oxic/anoxic interface, and apart from ensuring a supply of
NO3- from the water column, the close proximity of the nitrification and
denitrification processes facilitates an efficient coupling between the two processes
through transport of nitrification products (NO2- /NO3-) across the oxic/anoxic
interface. This coupling implies a loss of remineralized N from the system, and
nitrification therefore plays a key role in determining the fate of organic bound
nitrogen that is deposited on the seafloor.
In order to understand the regulation of denitrification in estuarine sediments it is
useful to look at the dependency on the NO3- source: The dependency of
denitrification on bottom-water NO3- and on NO3- produced by sediment nitrification,
respectively. Both of these dependencies are measurable with the isotope pairing
technique (Nielsen 1992) and can also be deduced from interpretation of NO3porewater profiles (Meyer et al. 2001).
O2 (µM)
0 20 40 60 80 100 -0.2 -0.1 0.0 0.1 0.2
NO3 /NH4 (µM)
Production (nmol cm s )
Figure 1. Porewater profiles of NO3-, NH4+ and O2 as measured with ion selective sensors
(Ref) (left) and volume specific NO3- production (right). Positive production indicates
nitrification, while negative production indicates denitrification. Data from Berg et al. (1998).
A prerequisite for denitrification of bottom-water NO3- is the transport of NO3- from
the water column to the anoxic sediment strata. Fich’s laws of diffusion imply that the
diffusive transport is dependent on both the thickness of the oxic zone and the
concentration of NO3- in the water column. The rate of denitrification of bottomwater NO3- is therefore ideally inversely proportional to the O2 penetration depth and
directly proportional to the concentration of NO3- in the water column. This ideal
relationship was demonstrated experimentally by Rysgaard et al. (1994) via
manipulations of bottom-water O2 and NO3- concentrations.
In estuarine systems oxygen penetration depths and bottom-water NO3concentrations may vary independently. The O2 penetration depth is determined by
the bottom-water O2 concentration and in sediments without benthic photosynthesis
N. Risgaard-Petersen
by the rate at which O2 is consumed in the sediment (Revsbech et al. 1980). Thus, an
increase in O2 consumption will lead to a decrease in O2 penetration depth. On
seasonal scale maximum sediment O2 consumption rates and minimum O2
penetration depths are found mainly during the late summer months, while the NO3load is largest in the winter and early spring months. Therefore, the correlation
between bottom-water NO3- and denitrification observed in seasonal studies can be
poor (Dong et al. 2000 a. o.).
Christensen et al. (1990) developed a simple model to predict the rate of
denitrification of bottom-water NO3- from the bottom-water O2 and NO3concentrations and the sediment O2 consumption:
D( NO3−) C NO3 1
) − 1⎥ (9.1)
Dw = FO2 α ⎢ (1 +
D (O 2 ) C O α
Where Dw is the rate of denitrification of NO3- from the water column and FO2 the
sediment O2 consumption rate. D(NO3-) and D(O2) represent the diffusion
coefficients of NO3- and O2, respectively, while α is the ratio between depth-specific
denitrification and O2 consumption activity (app. 0.8 (Christensen et al. 1990)).
The model assumes that the depth-specific activities of denitrification and O2
consumption are constant and that diffusion is the major transport process and,
furthermore, that denitrification is the sole NO3- consuming process. This implies that
the model is not useful in sediments with significant algal activity or in sediments
where bacteria capable of reducing NO3- to NH4+ dominate NO3- reduction. In such
sediments the model typically overestimates denitrification. In sediments where the
model assumptions are fulfilled, however, the model predicts the rate of
denitrification well. Data from station 75, Randers Fjord (Nielsen et al. 2001), is a
good example of application of the model (Fig. 2).
Rysgaard et al. (1995) demonstrated an equally good agreement between modelled
and measured data in the Danish estuary Kertinge Nor.
The model obviously predicts that an increase of the NO3- load will result in an
increase in denitrification activity. The model also predicts that increasing the organic
load to sediments (for instance through sedimentation of phytoplankton cells
following an algal bloom) and a resultant increase in the sediment O2 consumption
rate (Sloth et al. 1995) will lead to elevated removal of bottom-water NO3- via
denitrification. This relationship has been demonstrated through in situ measurements
as well as experimentally. Jensen (1988) thus observed increasing denitrification
activity following the spring phytoplankton bloom in the Bay of Aarhus. Caffrey et
al. (1993) and Sloth et al. (1995) likewise demonstrated experimentally that
increasing the organic load to sediments would result in elevated denitrification of
bottom-water NO3-.
Measured Dw (µmol N m h )
Model estimate (µmol N m h )
Figure 2. Measured vs. modelled rates of denitrification of NO3- supplied from the water
column (DW). The model accurately predicts the measured values (r2 =0.9289, p<0.0001).
Data from Nielsen et al. (2001).
The critical step in the removal of remineralized N from sediments through coupled
nitrification-denitrification is generally the nitrification process and its vertical
distribution in the oxic zone. The distribution of the processes shown in Fig. 1 is
typical of a stable sediment environment where bacterial populations have established
themselves in zones with an optimal substrate supply, and in sediments such as this a
very efficient coupling between nitrification and denitrification can be achieved. In
more unstable environments, e.g. sediments exposed to frequent resuspension events,
bioturbation or benthic primary production, a more even distribution of nitrification
in the oxic zone can be observed (Meyer et al. 2001). In this type of sediment the
coupling between nitrification is less efficient and a higher proportion of the NO3produced through sediment nitrification will escape denitrification and thus become
available for new primary production.
Being an aerobic litotrophic process nitrification is dependent on the supply of both
NH4+ and O2. Therefore, O2 deficit in the water column will obviously cause
nitrification to cease and with it the coupled nitrification-denitrification (Rysgaard et
al. 1994).
N. Risgaard-Petersen
In estuarine sediments NH4+ produced through mineralization of organic bound
nitrogen is the major NH4+ source for nitrification since NH4+ concentrations in the
water column are generally low. However, mineralization of organic bound nitrogen
results in consumption of O2 either directly or indirectly via the oxidation of reduced
compounds such as H2S, Mn+ and Fe++ produced via anoxic mineralization. Thus, the
biogeochemical processes responsible for the production of substrate for nitrification
at the same time compete with the nitrifying bacteria for O2.
Blackburn (1990) demonstrated this antagonistic dependency of nitrification and
coupled nitrification-denitrification on the sediment mineralization processes through
model simulations. Based on these simulations it was concluded that increasing the
organic load to the sediment and thereby promoting mineralization does not result in
a parallel stimulation of the removal of nitrogen via coupled nitrificationdenitrification. Beyond a certain point nitrification becomes O2 limited due to oxygen
consumption resulting from increased carbon availability, and this reduces the loss of
nitrogen via coupled nitrification-denitrification. Experimental evidence for this idea
was provided by Sloth et al. (1995) who measured nitrification and denitrification in
sediment cores loaded with different amounts of organic material (Fig. 3).
Nitrification and denitrification rates were highest in the moderately loaded sediments
because the availability of NH4+ was greater in these cores than in the unloaded
control cores.
DN (mmol N m d )
30 g dw
100 g dw
100 surf
Figure 3. Coupled nitrification-denitrification in sediment cores loaded experimentally with
yeast. 30 g DW and 100 g DW designate mixing of 30 g DW and 100 g DW of organic material
into the sediment, respectively, while 100 surf designates the addition of 100 g of organic
material to the sediment surface. Redrawn from Sloth et al. (1996).
In cores heavily loaded with organic material nitrification and consequently coupled
nitrification-denitrification was significantly reduced or absent because of O2
limitation and probably also due to irreversible inhibition of nitrification by H2S
(Henriksen and Kemp 1988; Joye and Hollibaugh 1995) during the initial phases of
the experiment.
A case story from Kertinge Nor, Denmark, reported by Dalsgaard et al. (1999) is a
good example of how organic loading to the estuaries may influence N loss from
sediments via coupled nitrification-denitrification. The external nutrient load to the
estuary was reduced from the year 1990, but two years later, in 1992, the estuary was
still hypereutrophic and exhibited a high sediment O2 demand (2705 ton O2 y-1) due
to deposition of planktonic alga cells fuelled by a large pool of nutrients that had
accumulated in the sediment. In 1995 the sediment nutrient pool was markedly
reduced, and this resulted in reduced pelagic primary production and a decrease in the
organic load reaching the sediment. Sediment O2 consumption was reduced to 1578 t
O2 y-1 and coupled nitrification-denitrification increased from 6.1 t N y-1 in 1992 to
17.1 t N y-1 in 1995.
Dissimilatory NO3- reduction to NH4+ (DNRA) represents an alternative NO3reduction pathway which may compete with denitrification. Bacteria capable of
reducing NO3- to NH4+ may dominate anaerobic NO3- reduction in carbon rich
environments low in NO3- (Tiedje 1982, Tiedje 1988). Gilbert et al. (1997) thus
reported a capacity for DNRA in sediments exposed to high organic loading from
mussel farms in a Mediterranean lagoon.
In a study in the Danish estuary Horsens Fjord (Christensen et al. 2000) it was
observed that the relative importance of DNRA decreased with increasing distance
from fish cages deployed in the fjord. Just below the cage approximately 95% of
NO3- reduction proceeded via the ammonification pathway, while denitrification was
the dominant NO3- reducing process 100 m from the cage. The dataset suggested a
strong positive correlation between C-mineralization (measured as sediment O2
consumption) and the relative importance of DNRA (Fig. 4). Thus, with increasing
mineralization activity DNRA becomes relatively more important than denitrification.
Despite the fact that DNRA has been known to take place in decades there is still a
lack of understanding of the role of the process in estuaries, and systematic
measurements of the distribution of the process relative to denitrification are very
much needed.
N. Risgaard-Petersen
O2 consumption (mmol m d )
Figure 4. Proportion of denitrification and DNRA to total NO3- reduction as function of the
sediment O2 consumption rates. Data from Christensen et al. (2000).
Benthic microalgae are concentrated at the sediment/water interface, and their activity
may significantly alter the chemical microenvironment in the surface sediment.
Microphytobenthic photosynthesis and respiration may result in large diurnal
oscillations in both O2 concentrations and the concentrations of inorganic carbon as
well as in pH (Revsbech 1983). Likewise, through assimilatory nutrient uptake, algae
may have a significant impact on the porewater concentrations of NO3- and NH4+ in
the surface sediment.
Examples of diurnal oscillations in O2 and pH in estuarine sediment colonized by
microalgae are given in Figure 5. During illumination (200 µmol photons m-2 s-1)
photosynthesis resulted in an O2 concentration of 500 µM in the most active layer
approximately 0.2 mm below the surface. The O2 penetration depth was 2.5 mm. At
the depth of maximum activity pH was 9.5 due to intense microalgal carbon
assimilation. In darkness respiration resulted in a decrease in both sediment O2
concentrations and pH. The O2 penetration depth in darkness was only 25% of the
depth measured in light.
Examples of porewater profiles of NO3-, NH4+ and O2 in freshwater sediment
colonized by benthic microalgae are given in Figure 6. During illumination NH4+ was
not detectable in the oxic zone and a significant reduction in NO3- concentration was
observed in the 0.5-mm depth interval. The zone within the oxic layer exhibiting the
lowest NO3- concentration corresponded to the zone in which the O2 concentration
peaked, and this suggests that microphytobenthic NO3- uptake was responsible for the
observed decrease in NO3-. In darkness no indications of microphytobenthic N uptake
were seen. Ammonium penetrated the oxic zone and was released to the water
column, while NO3- was consumed exclusively in the anoxic zone, a fact that can be
attributed to denitrification.
[O2] µM
[O2] µM
Figure 5. Porewater profiles of O2 and pH measured with microsensors in a microalgaecolonized sediment. Original (data).
The dramatic changes in porewater chemistry of the surface sediment induced by
benthic microalgae must be expected to have a significant impact on the nitrification
and denitrification processes. The presence of benthic microalgae may lead to diurnal
variations in nitrification and denitrification activity and to a reduction of the overall
capacity of the sediment to remove nitrogen via denitrification.
Nielsen et al. (1990) investigated the effect of photosynthesis and assimilation on
denitrification of bottom-water NO3- in stream sediment exposed to high NO3concentrations (300 µM). The authors found that the overall rate of denitrification
was reduced by 70% in light. This reduction was ascribed to the O2 penetration depth
being greater in light than in the dark, leading to an increase in diffusion distance for
NO3- from the water column to the anoxic sediment strata. Calculated porewater NO3profiles suggested that microalgal NO3- assimilation did not influence denitrification.
Risgaard-Petersen et al. (1994) demonstrated a similar mechanism in a study on
nitrification and denitrification in estuarine sediments. Using NO3- biosensors (Larsen
N. Risgaard-Petersen
et al. 1997) Meyer et al. (2001) showed that at low NO3- concentrations (4 µM)
microalgae are efficient sinks for water-column NO3-, however, and are able to
significantly impede the flux of NO3- toward anoxic sediment layers through nutrient
[O2] µM
[O2] µM
20 40 60 80 100 120
NH4 /NO3 (µM)
0 20 40 60 80 100120
NH4 /NO3 (µM)
Figure 6. Porewater profiles of NO3-, NH4+ and O2 measured with ionselective sensors (NO3/NH4+) in an alga-colonized sediment. Original data.
The diurnal variations in O2 penetration depth may also induce diurnal variations in
nitrification activity and consequently in coupled nitrification-denitrification activity.
Risgaard-Petersen et al. (1994), Lorenzen et al. (1998) and An and Joye (2001)
observed higher rates of coupled nitrification-denitrification in light than in darkness.
This stimulation was attributed to a photosynthetically mediated increase in O2
penetration depth leading to increased nitrification activity (Risgaard-Petersen et al.
1994). However, several field studies have indicated that light-induced stimulation of
nitrification is not a general phenomenon (Rysgaard et al. 1995; Cabrita and Brotas
2000; Sundback and Miles 2000). It has been suggested that stimulation of coupled
nitrification-denitrification in light only occurs when competition between algae and
nitrifiers and denitrifiers for DIN is not limiting the bacterial processes (Rysgaard et
al. 1995; Dong et al. 2000). Although attractive this hypothesis is unsupported by
experimental evidence at present. In a review on nitrification Henriksen and Kemp
(1988) presented experimental data showing that the potential for nitrification in the
uppermost sediment strata was reduced in sediments exposed to light/dark cycles for
6 weeks as compared to sediments incubated in the dark. It was suggested that light
and a combination of factors linked to algal activity (i.e. high O2 concentrations, high
pH and induction of NH4+ and CO2 limitation) were responsible for the observed
µmol N m h
Heterotrophic sediments
Autotrophic Sediments
Figure 7. Coupled nitrification-denitrification rates in bare sediments with and without
microphytobenthic photosynthesis. Sediments are classified as heterotrophic if O2 fluxes in
light ≤ O2 fluxes in the dark and <0 (n=200) Sediment is classified as autotrophic if O2 fluxes
in light > O2 fluxes in the dark and >0 (n=944). Data from Dalsgaard et al. (2000).
Based on their observations the authors proposed that this reduced ability of nitrifying
bacteria to survive in alga-colonized sediment would lead to a reduced loss of
nitrogen via coupled nitrification-denitrification from the system as compared to
sediment without benthic microalgae.
Results from the EU-funded NICE project (Nitrogen Cycling in Estuaries (Dalsgaard
et al. 2000)) have shown that mean coupled nitrification-denitrification rates in
sediments colonized by benthic microalgae are significantly lower (ca. 50%, Fig. 7)
than rates measured in heterotrophic sediments.
Experimental studies indicate that a major factor responsible for the reduction in
coupled nitrification-denitrification in alga-colonized sediments is the generation of
unfavourable growth conditions for nitrifying bacteria in these systems. Furthermore,
induction of nitrogen limitation of nitrifying bacteria seems to be a major controlling
mechanism in alga-colonized sediments, while factors associated with photosynthesis
(e.g. high pH and O2 concentrations) are of minor importance (Risgaard-Petersen,
2003). Benthic microalgae may enhance the flux of inorganic N from the water
column toward the sediment through nutrient assimilation and are likewise able to
reduce the efflux of remineralized nitrogen from the sediment. The suppression of
nitrification and denitrification in these sediments suggests a highly conservative
environment for nitrogen characterized by efficient coupling between sediment N
regeneration and N assimilation.
N. Risgaard-Petersen
Rooted aquatic angiosperms may alter the sediment environment and thereby the
niches occupied by nitrifying and denitrifying bacteria. The plants may release O2 to
the rhizosphere via radial O2 loss from the roots if the respiratory demand of the roots
is lower than the O2 supply or if O2 consuming processes in the rhizosphere take
place closer to the O2 rich cortex tissue compared to O2 consuming processes in the
root tissue (Armstrong 1994). Consequently, free O2 may be present in the otherwise
anoxic rhizosphere if the reductant formation through chemical and microbial
processes is sufficiently slow (Caffrey and Kemp 1992; Pedersen et al. 1995 a.o.).
This introduction of O2 into the rhizosphere creates oxic/anoxic interfaces in an
otherwise reduced environment.
Aquatic plants also assimilate and incorporate inorganic nitrogen into organic matter
and a significant part of this nitrogen may be supplied from the sediment (Short and
McRoy 1984; Caffery and Kemp 1992; Pedersen and Borum 1992). Thus, the plants
compete with nitrifying and denitrifying bacteria for nitrogen, and recent results have
shown that angiosperm roots may be better competitors than nitrifiers for sediment
NH4+ (Verhagen et al. 1995). Several studies on freshwater systems have
demonstrated the ability of aquatic plants to stimulate nitrogen loss through O2
release in the rhizosphere (Reddy et al. 1989; Christensen et al. 1986; RisgaardPetersen and Jensen 1997; Arth et al. 1998). In the case of marine species the dataset
is more limited. Direct evidence for rhizosphere-associated coupled nitrificationdenitrification has only been presented for a few species and the activity and the
ability of the plants to stimulate coupled nitrification-denitrification via O2 release is
apparently strongly linked to the growth cycle of the plants and therefore displays a
distinct seasonal pattern. Caffrey and Kemp (1992) observed that denitrification
activity in Potamogeton perfolitiatus-colonized sediments was higher than the
activity measured in bare sediments in July when the plant biomass was high. On the
other hand, Risgaard-Petersen et al. (1998) observed that in Zostera marina surfaceand rhizosphere-associated coupled nitrification-denitrification was only measurable
in the spring. In the summer months a high sediment O2 demand probably prevented
the formation of oxic/anoxic interfaces in the rhizosphere and with it root-associated
It has been suggested that denitrification activity is enhanced in seagrass-colonized
sediments compared to the activity in bare sediments due to the ability of the plants to
stimulate coupled nitrification-denitrification via root O2 release (Flindt 1994).
However, direct measurements using 15N isotopes have shown that the ability of for
instance Zostera marina and Z. noltii to stimulate coupled nitrification-denitrification
via root O2 release is generally low (Rysgaard et al. 1996; Risgaard-Petersen et al.
1998; Ottosen et al. 1999; Welsh et al. 2000). A comparison between Danish
sediments with and without benthic macrophytes presented by Risgaard-Petersen and
Ottosen (2000) showed that coupled nitrification-denitrification on average
corresponded to 24% of the activity reported from various unvegetated sediments in
Danish fjords
Depth (cm)
Activity (nmol N cm h )
Figure 8. Depth profiles of coupled nitrification-denitrification measured in light and in the
dark in Z. marina-colonized sediment. In April maximum activity was measured in the upper
sediment strata in light and denitrification was measurable until a depth of 8 cm. Coupled
nitrification-denitrification activity in the deeper sediment strata was caused by O2 release
from the roots. In darkness the activity was below detection limit. In August coupled
nitrification-denitrification activity was undetectable. Redrawn from Risgaard-Petersen et al.
This suggests that eelgrass-colonized sediment is an unfavorable environment for
nitrifying and denitrifying bacteria, possibly due to the effect of competition with the
plants for nitrogen as shown in for instance Rumex species (Engelaar et al. 1991).
Alternatively, the unfavorable conditions may be a result of enhanced organic loading
to the sediment induced by plants (Hemminga et al. 1991) or enhanced sediment
SO42- reduction activity (Blaaberg et al. 1998; Holmer and Nielsen 1997). These
conditions may inhibit nitrification (Henriksen and Kemp 1988; Sloth et al. 1995) and
improve conditions for bacteria capable of dissimilatory NO3- reduction to NH4+
(Tiedje 1988). The latter process has recently been shown to play a major role in
seagrass-colonized sediments (Rysgaard et al. 1996). Just like microphytobenthos
colonized sediments the seagrass meadows may thus represent a highly conservative
environment for nitrogen, where the N-cycle is dominated by the primary
productivity of the plant community and the associated assimilatory demand for fixed
N to support this productivity.
N. Risgaard-Petersen
mmol N m d
Bare sediment
Fig 9. Coupled nitrification-denitrification rates in Z. marina-colonized sediments (n=23) and
bare sediments from Danish estuaries (n=71) exposed to N-concentrations below 12 µM. Data
from Risgaard-Petersen and Ottosen (2000).
In this chapter the process of denitrification has been discussed in relation to
responses to organic loads, nutrient concentrations and benthic primary production.
On the basis of this discussion I propose the following scenarios regarding the
process in relation to changes in nutrient load to the estuarine environment:
When nutrient loading to estuaries is increased and with it the concentration of
nitrogen in the water column, denitrification will increase in response to the elevated
NO3- concentrations. Given that increased N loading results in elevated pelagic
primary production (Borum and Sand-Jensen 1996) and a concurrent increase in
deposition of organic matter at the seafloor, sediment mineralization will increase and
the NH4+ availability for nitrifying bacteria improve. This will, to a certain point, lead
to enhanced nitrogen removal via coupled nitrification-denitrification. However, as
mineralization increases, nitrification will become O2 limited leading to a reduced
loss of nitrogen via coupled nitrification-denitrification. If water-column O2
concentrations decline to anoxia as often happens below the pycnocline when the
organic load is high, coupled nitrification-denitrification will cease and remineralized nitrogen will escape from the sediment as NH4+. At high nutrient loads
denitrifying bacteria may eventually become out-competed by bacteria capable of
reducing NO3- to NH4+. This may further contribute to the establishment of a positive
feedback loop between pelagic primary production and sediment NH4+ generation.
When nutrient loading to estuaries is reduced the amount of NO3- available for
denitrification decreases. Furthermore, as light conditions at the seafloor improve, a
shift from pelagic to benthic primary production is to be expected (Borum and SandJensen 1996). A shift such as this will significantly alter the benthic N cycle. Benthic
primary producers will assimilate a greater amount of water-column NO3- before it
becomes available for denitrification, and may likewise reduce the transport of NO3from the water column to the anoxic sediment strata via a photosynthetically
mediated increase in O2 penetration depth. Moreover, benthic primary producers
seem to create unfavorable growth conditions for nitrifying bacteria resulting in
inhibition of coupled nitrification-denitrification. Thus, reduced nutrient loading will
lead to reduced loss of N via denitrification and a temporary accumulation of N in
plant biomass.
An, S.M. and S.B. Joye (2001). Enhancement of coupled nitrification-denitrification by benthic
photosynthesis in shallow estuarine sediments. Limnology and Oceanography 46:62-74
Arth, I., P. Frentzel and R. Conrad (1998). Denitrification coupled to nitrification in the rhizosphere of rice.
Soil. Biology and Biochemistry 40:509-515
Berg, P., N. Risgaard-Petersen and S. Rysgaard (1998). Interpretation of measured concentration profiles in
sediment pore water. Limnology and Oceanography 43:1500-1510
Blackburn, T.H. (1990). Denitrification model for marine sediments 323-337. In: Revsbech, N. P. and J. S.
Sørensen [eds.], Denitrification in soils and sediments.Plenum Press
Blaaberg, V., K. Mouritsen and K. Finster (1998). Diel cycles of sulphate reduction rates in sediments of a
Zostera marina bed, Denmark. Aquatic . Microbial . Ecology l 15:97-102
Borum, J. and K. Sand-Jensen (1996). Is total primary production in shallow coastal matine waters
stimulated by nitrogen loading? Oikos 76:406-410
Cabrita, M.T. and V. Brotas (2000). Seasonal variation in denitrification and dissolved nitrogen fluxes in
intertidal sediments of the Tagus estuary, Portugal. Marine Ecology Progress Series 202:51-65
Caffrey, J.M. and W.M. Kemp (1992). Influence of the submersed plant Potamogeton Perfuliatus, on
nitrogen cycling in estuarine sediments. Limnology and Oceanography 37:1483-1495
Caffrey, J.M., N.P. Sloth, H.F. Kaspar, and T.H. Blackburn (1993). Effect of organic loading on
nitrification and denitrification in a marine sediment microcosm. FEMS Microbial Ecology. 12:159167
Christensen, P.B., L.P. Nielsen, J. Sørensen and N.P. Revsbech (1990). Denitrification in nitrate-rich
streams: Diurnal and seasonal variation related to benthic oxygen metabolism. Limnology and
Oceanography 35:640-651
Christensen, P.B., S. Rysgaard, N.P. Sloth, T. Dalsgaard and S. Schwaerter (2000). Sediment
mineralization, nutrient fluxes, denitrification and dissimilatory nitrate reduction to ammonium in an
estuarine fjord with sea cage trout farms. Aquatic Microbial Ecology 21:73-84
N. Risgaard-Petersen
Dalsgaard, T., P.B. Christensen, S. Rysgaard and N. Risgaard-Petersen (1999). Kvælstoffjernelse i danske
kystnære farvande: betydning og regulering 102-118. In: Aa., L. B. [eds.], Havmiljøet ved
årtusindskiftet.Olsen &Olsen
Dalsgaard, T., L.P. Nielsen, V. Brotas, P. Viaroli, G.J.C. Underwood, D.B. Nedwell, K. Sundbäck, S.
Rysgaard, A. Miles, M. Bartoli, L.F. Dong, D.C.O. Thornton, O.L.D.M, G. Castaldelli and N.
Risgaard-Petersen (2000). Final Scientific report for NICE Nitrogen Cycling in Estuaries Commision
of the European Communities DC XII, Brussels
Dong, L.F., D.C.O. Thornton, D.B. Nedwell and G.J.C. Underwood (2000). Denitrification in sediments of
the River Colne estuary, England. Marine Ecology Progress Series 203:109-122
Flindt, M.R (1994). Measurements of nutrient fluxes and mass balances by on-line in situ dialysis in a
Zostera marine bed culture. Verhandlungen Internationale Vereinigung Limnologie 25:2259-2264
Gilbert, F., P. Souchu, M. Bianchi and P. Bonin (1997). Influence of shellfish farming activities on
nitrification, nitrate reduction to ammonium and denitrification at the water-sediment interface of the
Thau Lagoon, France. Marine Ecology Progress Series 151:143-153
Hemminga, M.A., P.G. Harrison and F. van Lent (1991). The balance of nutrient losses and gains in
seagrass meadows. Marine Ecology Progress Series 71:85-96
Henriksen, K. and M. Kemp (1988). Nitrification in estuarine and coastal marine sediments 201-255. In:
Blackburn, T. H. and J. Sørensen [eds.], Nitrogen cycling in coastal marine environments, SCOPE.
John Wiley & Sons Ltd
Holmer, M. and S.L. Nielsen (1997). Sediment sulfur dynamics related to biomass-density in Zostera
marina (eelgrass) beds. Marine Ecology Progress Series. 146:163-171
Howarth, R.W., R. Marino, J. Lane and J.J. Cole (1988). Nitrogen-Fixation in Fresh-Water, Estuarine, and
Marine Ecosystems .1. Rates and Importance. Limnology and Oceanography 33:669-687
Jensen, M.H., T.K. Andersen, and J. Sørensen (1988). Denitrification in coastal bay sediment: regional and
seasonal variation in Aarhus Bight, Denmark. Marine Ecology Progress Series. 48:155-162
Larsen, L.H., T. Kjær and N.P. Revsbech (1997). A microscale NO3- biosensor for environmental
applications. Analytcal Chemistry. 69:3527-3531
Lorenzen, J., L.H. Larsen, T. Kjaer and N.P. Revsbech (1998). Biosensor determination of the microscale
distribution of nitrate, nitrate assimilation, nitrification, and denitrification in a diatom-inhabited
freshwater sediment. Applied and Environmental Microbiology 64:3264-3269
Meyer, R.L., T. Kjaer and N.P. Revsbech (2001). Use of NOx- microsensors to estimate the activity of
sediment nitrification and NOx- consumption along an estuarine salinity, nitrate, and light gradient.
Aquatic Microbial Ecology 26:181-193
Nielsen, K., N. Risgaard-Petersen, B. Somod, S. Rysgaard and T. Bergo (2001). Nitrogen and phosphorus
retention estimated independently by flux measurements and dynamic modelling in the estuary,
Randers Fjord, Denmark. Marine Ecology Progress Series 219:25-40
Nielsen, L.P. (1992). Denitrification in sediments determined from nitrogen isotope pairing. FEMS
Microbial. Ecology. 86:357-362
Nielsen, L.P., P.B. Christensen, N.P. Revsbech and J. Sørensen (1990). Denitrification and photosynthesis
in stream sediment studied with microsensor and whole-core techniques. Limnology and
Oceanography 35:1135-1144
Ottosen, L.D.M., N. Risgaard-Petersen and L.P. Nielsen (1999). Direct and indirect measurements of
nitrification and denitrification in the rhizosphere of aquatic macrophytes. Aquatic Microbial Ecology
Risgaard-Petersen, N 2003. Coupled nitrification-denitrification in autotrophic and heterotrophic estuarine
sediments: On the influence of benthic microalgae. Limnology and Oceanography 48:93-105.
Risgaard-Petersen, N. and L.D.M. Ottosen (2000). Nitrogen cycling in two temperate Zostera marina beds:
seasonal variation. Marine Ecology Progress Series 198:93-107
Risgaard-Petersen, N., S. Rysgaard, L.P. Nielsen and N.P. Revsbech (1994). Diurnal variation of
denitrification and nitrification in sediments colonized by benthic microphytes. Limnology and
Oceanography 39:573-579
Risgaard-Petersen, N., T. Dalsgaard, S. Rysgaard, P.B. Christensen, J. Borum, K. McGlathery and L.P.
Nielsen (1998). Nitrogen balance of a temperate eelgrass Zostera marina bed. Marine Ecology
Progress Series 174:281-291
Rysgaard, S., P.B. Christensen and L.P. Nielsen (1995). Seasonal variation in nitrification and
denitrification in estaurine sediment colonized by benthic microalgae and bioturbating infauna. Marine
Ecology Progress Series. 126:111-121
Rysgaard, S., N. Risgaard-Petersen and N.P. Sloth (1996). Nitrification, denitrification, and nitrate
ammonification in sediments of two coastal lagoons in Southern France. Hydrobiologia 329:133-141
Rysgaard, S., N. Risgaard-Petersen, N.P. Sloth, K. Jensen and L.P. Nielsen (1994). Oxygen Regulation of
Nitrification and Denitrification in Sediments. Limnology and Oceanography 39:1643-1652
Sloth, N.P., H. Blackburn, L.S. Hansen, N. Risgaard-Petersen and B. Lomstein, Aa. (1995). Nitrogen
cycling in sediments with different organic loading. Marine Ecology Progress Series. 116:163-170
Sundback, K. and A. Miles (2000). Balance between denitrification and microalgal incorporation of
nitrogen in microtidal sediments, NE Kattegat. Aquatic Microbial Ecology 22:291-300
Sundback, K., A. Miles and E. Goransson (2000). Nitrogen fluxes, denitrification and the role of
microphytobenthos in microtidal shallow-water sediments: an annual study. Marine Ecology Progress
Series 200:59-76
Tiedje, J.M., Sexstone,A..J.,Myrold,D.D. and Robinson,J.A. (1982). Denitrification: ecological niches,
competition and survival. Antonie van Leeuwenhoek 48:569-586
Tiedje, J.M. (1988). Ecology of denitrification and dissimilatory nitrate reduction to ammonium 179-244.
In: Zender, A. J. B. [eds.], Ecology of anaerobic microorganisms. John Wiley and Sons
Trimmer, M., D.B. Nedwell, D.B. Sivyer and S.J. Malcolm (2000). Seasonal benthic organic matter
mineralisation measured by oxygen uptake and denitrification along a transect of the inner and outer
River Thames estuary, UK. Marine Ecology Progress Series 197:103-119
Welsh, D.T., M. Bartoli, D. Nizzoli, G. Castaldelli, S.A. Riou and P. Viaroli (2000). Denitrification,
nitrogen fixation, community primary productivity and inorganic-N and oxygen fluxes in an intertidal
Zostera noltii meadow. Marine Ecology Progress Series 208:65-77
Zumft, W. G. (1992). The denitrifying prokaryotes. 544-582 in: A. Balows,The prokaryotes : A handbook
on the biology of bacteria: Ecophysiology, isolation, identification, applications. Springer
N. Risgaard-Petersen
N. Risgaard-Petersen: Department of Marine Ecology, National Environmental
Research Institute, Vejlsøvej 25, DK-8600 Silkeborg, Denmark.
The overall purpose of this book has been to summarize and evaluate the mutual
interactions between nutrients and vegetation in shallow coastal ecosystems. Several
books and scientific papers have focused on how nutrient richness modify the
structure and dominance patterns of marine plant communities, while less attention
has been given to the evaluation of how different types of marine plants may affect
nutrient dynamics within the systems they live. The major topic of this book is
therefore to review how plants affect the overall nutrient dynamics (i.e. transport,
recycling, temporary or permanent retention) in shallow coastal ecosystems and to
judge whether or not such effects may vary systematically with dominance of
different plant types – “plants” in this context includes algae.
Shallow coastal ecosystems are, in contrast to the ocean where phytoplankton is the
only important primary producer, inhabited by taxonomically and functionally
diverse plant communities. The plant communities in shallow coastal ecosystems are
typically made up by different plant types including rooted plants (e.g. seagrasses,
salt marsh plants etc.), slow-growing, perennial macroalgae, fast-growing, ephemeral
macroalgae and benthic, epiphytic and pelagic microalgae that each may contribute
significantly to total plant biomass and production (Chapter 1). Comparative studies
show that these plant types differ systematically from each other in a number of
morphologically, physiological, functional and ecological properties as discussed by
Sand-Jensen and Nielsen in Chapter 2. These variations led us to expect that a shift in
dominance from seagrasses and/or slow-growing macroalgae to more or less
complete dominance by phytoplankton would increase the importance of within
system remineralization and recycling of nutrients and reduce the importance of
nutrient retention and removal through permanent burial and other processes. But is
this so?
1.1 Eutrophication and plant responses
A large number of abiotic and biotic factors contribute to determine the exact
structure of plant communities in shallow coastal ecosystems, but the role of nutrients
has received considerable attention in the last couple of decades due to the increasing
problems related to coastal eutrophication. It is now clear that increasing nutrient
richness of estuaries and other coastal marine areas – known as cultural
eutrophication – leads to structural shifts in the plant assemblage, although it does not
S. Nielsen, G. Banta and M. Pedersen (eds.), Estuarine Nutrient Cycling:
The IInfluence of Primary Producers, 281-292.
© 2004 Kluwer Academic Publishers. Printed in the Netherlands.
S.L. Nielsen, M.F. Pedersen and G.T. Banta
necessarily lead to an increase in total primary production (Borum 1995; Borum &
Sand-Jensen 1996; Schramm and Nienhuis 1996).
The composition of coastal plant communities and the structural changes that follow
eutrophication is dealt with in detail by Hauxwell and Valiela in Chapter 3.
Seagrasses and large, slow-growing, perennial macroalgae (mainly brown algae) tend
to dominate plant communities in shallow coastal ecosystems under pristine
conditions. Although the total plant biomass in such systems may be completely
dominated by slow-growing, perennial macrophytes, total primary production is often
much more “diverse”, receiving significant contributions from other plant types such
as ephemeral macroalgae, benthic microalgae and phytoplankton. This divergence
emerges as a result of the inherent variation in physiological and ecological rates
brought about by the different plant types - seagrasses and large, perennial
macroalgae have lower specific growth rates and thus a lower productivity than
ephemeral macroalgae and most types of microalgae (Borum & Sand-Jensen 1996;
Nielsen et al., 1996).
One of the primary effects of eutrophication in coastal marine areas is an increase in
the contribution of ephemeral macroalgae, microphytobenthos and phytoplankton to
total plant biomass and production, whereas the importance of slow-growing,
perennial macrophytes decrease (Sand-Jensen & Borum, 1991). The mechanisms
behind these structural changes are now relatively well understood - seagrasses and
slow-growing, perennial macroalgae are superior competitors for nutrients due to the
low nutrient requirements and a high capacity to store nutrients and these plants
therefore dominate the plant community under nutrient poor conditions (Pedersen &
Borum, 1996). Fast-growing macroalgae and microalgae have, in contrast, higher
nutrient demands per unit biomass and time and are therefore stimulated by high
nutrient supply. These plant types are superior competitors for light due to their thin
thalli (Agustí et al., 1994) and because of their epiphytic, free-floating or pelagic
nature, and these attributes allow them to overgrow benthic and slower growing
primary producers in situations where nutrients are no longer limiting.
Light limitation caused by shading from drifting macroalgae, epiphytes and/or
phytoplankton seem to be the most important cause for the observed decline of
seagrasses and other benthic macrophytes during eutrophication, but direct nitrogen
toxicity (Burkholder et al., 1992; Touchette & Burkholder, 2002) and unfavorable
changes in the biogeochemical environment of the seagrasses (e.g. Greve et al., 2003;
Holmer & Bondgaard, 2001) have lately been brought forward as additional potential
causes. Whether eutrophication of a shallow coastal ecosystem leads to final
dominance of ephemeral macroalgae or by phytoplankton may depend on the
resulting nutrient regime but also on the water residence time of the system in
question since low residence time seems to favor dominance by ephemeral
macroalgae whereas longer residence time is claimed to favor dominance by
Although the overall relationship between eutrophication and changes in plant
composition is well documented through a large number of studies it should be
remembered that these relationships are not as simple and predictable as often
suggested in the literature. Several processes may either affect the composition
directly or indirectly through feedback effects on the availability of nutrients.
Substantial losses of seagrasses and other benthic macrophytes may, for example,
increase the frequency and importance of resuspension events (Pedersen et al., 1995),
which may in turn enhance the release of nutrients from the sediment and thus
stimulate the productivity of fast-growing plant types even further (Duarte, 1995).
Other processes, such as grazing, can delay or even prevent the accumulation of fastgrowing ephemeral plants and thus act as a buffer against one of the primary
responses to nutrient enrichment. Selective grazing on specific plant types may,
however, at the same time increase rates of remineralization and thus contributes to
the stimulation of growth of other plant types that are not grazed to the same extent.
Intense grazing on phytoplankton by benthic filter feeders, for example, may
stimulate the productivity of ephemeral macroalgae and thus contribute to the
regulation of the plant assemblage. The primary response to eutrophication may
finally be influenced by the hydrography of the system in question since low water
residence time and/or high physical exposure to waves may stimulate the export of
certain plant types. The exact nature of these processes, and the extent to which they
contribute to modify the structure of marine plant communities, needs further
Marine plants and algae are often highly abundant in shallow coastal ecosystems and
these systems rank among the most productive biomes in the world (Chapter 1). The
preliminary analysis carried out in Chapter 1 showed that most of the nutrients
received by coastal ecosystems from terrestrial and atmospheric sources are
assimilated by the primary producers regardless of the structural composition of the
plant community. Total assimilation seems therefore to depend mostly on total
nutrient availability (i.e., loadings). The further fate of these nutrients must therefore
be closely coupled to the fate of the primary production. Since the fate of the
produced plant material depends to a large extent on the plant type in question, it can
be expected that dominance of specific plant types influence whether nutrients are
“lost” from the ecosystem through export and burial or whether they are
remineralized and recycled to support further primary production within the system.
2.1 Mass transport and export
While the massive assimilation of nutrients by plants may delay the horizontal
transport and subsequent export of nutrients from the ecosystem, losses of plantbound nutrients through mass transport of plants themselves may prevent plant
material and associated nutrients from accumulating within the system. The export of
nutrients bound in plants and algae seems to be an important, but usually overlooked
process. Flindt et al. (Chapter 4) reports a series of studies undertaken to elucidate the
importance of nutrient losses from coastal ecosystems through mass transport of plant
matter. These studies show that total export of plant-bound nutrients is highly
variable among different systems but also that it can contribute substantially to the
export of particle bound nutrients. Export of plant-bound nutrients thus represented
63% of the total N-export from Venice Lagoon, Italy, 21% of the total N-export from
Roskilde Fjord, Denmark and less than 5% of the total N-export from Mondego
estuary in Portugal. The substantial export of plant-bound nutrients observed in some
systems suggests that this export may represent a significant percentage of the
S.L. Nielsen, M.F. Pedersen and G.T. Banta
nutrients originally received by the system from external sources or assimilated by
the primary producers. Unfortunately, very few studies allow such comparisons (see,
however, Salomonsen et al., 1999 and Bergamasco et al., 2003 for examples).
The absolute size of the export depends to a great extend on the hydrographical
conditions of the system in question but may also depend upon the dominance
patterns of the plant assemblages. Strong river influence and near unidirectional ,
flow of water seaward will, other things being equal, increase the importance of mass
transport, although net transport will also depend upon the morphometry of the
recipient. Plant material that is exported from the system will invariably be lost if the
system opens on a deeper water body where the plant matter sink to greater depths,
while some of the material may return when the tide turns or if the wind changes
when the system opens on a shallow water body, thus reducing the net export.
Different plant types show different patterns of transport. Attached macrophytes with
roots or strong hold-fasts (i.e. seagrasses and most perennial, slow-growing
macroalgae) are less susceptible to mass transport than most ephemeral macroalgae,
which are easily detached from their substrate, and phytoplankton, which is
suspended in the water column. Even strongly attached plants and algae may,
however, occasionally become detached by storms, grazing and other types of
physical disturbances. Furthermore, seagrass leaves are lost continuously as old
leaves are shed as part of the growth process. When detached, most macroalgae
moves as bedload along the bottom and different types of macroalgae appear to have
more or less the same threshold for transport, measured as water velocity needed to
keep them in suspension. The transport of macroalgae is therefore especially
dependent upon current speed and direction. Sloughed seagrass leaves move, in
contrast, primarily at the water surface, probably because of aerenchymatic air and
the transport is therefore affected by both currents and wind speed and direction.
All these observations taken together suggest that the absolute loss of nutrients due to
mass transport should increase with eutrophication since this should lead to a decline
of benthic, attached vegetation and an increase in the amount of free-floating,
ephemeral macroalgae and suspended microalgae that are more prone to mass
transport. The few studies available do, however, not allow a test of the hypothesis
that fast-growing plant types lose a larger proportion of their net primary production
than slow-growing, perennial plant types, since parallel estimates of production and
export losses are rare. Comparative studies that compare the proportions of
assimilated nutrients lost from coastal ecosystems dominated by different plant types
should therefore be conducted in the future.
2.2 Grazing losses – consequences for nutrient mineralization
The shift from dominance by seagrasses and other slow-growing perennial
macroalgae to dominance by ephemeral macroalgae and phytoplankton during
eutrophication is expected to affect the flux of nutrients to higher trophic levels since
these changes also represent a shift from production of low quality food to production
of high quality food (i.e. higher nutrient concentrations and lower content of
structural compounds and defense chemicals). Since ecosystem processes such as
grazing and decomposition, rely partly on these characteristics (e.g. Cebrián, 1999), it
is expected that remineralization of nutrients through grazing and decomposition will
change concomitantly with changes in plant composition.
Petersen and Cebrián deal extensively with the potential role of grazing upon
phytoplankton and benthic plants in Chapters 5 and 6, respectively. Benthic
suspension feeders – by far the most important grazers on phytoplankton in shallow
coastal ecosystems - can exert a strong top-down control on phytoplankton and
essentially strip the water column for suspended primary producers. For example, the
mean residence time for algal cells in the water above a dense populations of benthic
suspension feeders can be as low as 1 hour (ranging from 1 hour to 2 days) in shallow
areas. Whether or not this potential grazing impact is reached is partly under
hydrographical control since benthic filter feeders will be able to control the
phytoplankton biomass if their clearance time is lower than the water residence time,
whereas they are unable to control the phytoplankton when the system is stratified or
if the water residence time is lower than the clearance time.
The grazing pressure on benthic primary producers is less intensive than on pelagic
primary producers in most coastal ecosystems. Relative grazing intensity (i.e. percent
of biomass or production grazed) is partly controlled by the nutrient content of the
primary producers and grazing intensity is therefore lower for nutrient poor, perennial
macrophytes than for ephemeral macroalgae and microalgae, which have more
nutrient rich tissues. This pattern is similar to that found when grazing intensity is
compared across a wider range of primary producers including both terrestrial
vegetation as well as phytoplankton (Cebrián, 1999).
The variations in grazing intensity indicate, at first glance, that more plant biomass is
channeled to higher trophic levels when fast-growing, nutrient rich plants dominate
the plant community, but Cebrián shows elegantly that this is not the case - the
absolute amount of biomass (or carbon) lost to grazers per unit area and time is
approximately the same for different plant groups. The explanation is that absolute
grazing losses are positively correlated to absolute production and that slow-growing
perennial macrophytes tend to accumulate larger biomass and produce more per unit
area and time than faster growing plant types. Thus, the absolute amount of biomass
(or carbon) that is grazed per unit area and time remains therefore more or less the
same when compared across plant communities dominated by different plant groups.
Eutrophication leads to a replacement of nutrient poor, slow-growing, perennial
macroalgae by nutrient rich, ephemeral macroalgae and microalgae and grazing
intensity therefore increases with eutrophication, whereas the total flux of biomass
(carbon) remains approximately the same. Some field observations have, however,
shown that that the flux of carbon to herbivores may decrease under severe
eutrophication. This phenomenon may be due to some of the side effects of
eutrophication, i.e., the replacement of seagrasses and large, perennial macroalgae by
ephemeral macroalgae and microalgae which may represent a destruction of habitats
to many benthic herbivores and the increasing production of easy decomposable
organic matter which may increase the frequency of anoxia events. Both of these
changes can have detrimental effects on herbivore density and activity.
The absolute flux of nutrients that are channeled to higher trophic levels may,
however, increase with eutrophication, although this aspect is not addressed
specifically by Cebrián in Chapter 6. The N and P content of phytoplankton and
benthic microalgae are typically 2-5 fold higher than in slow-growing macrophytes
(Duarte, 1992; Cebrián, Chapter 6), so the flux of these nutrients to higher trophic
levels and, hence, the absolute amount of remineralized nutrients, may increase
during eutrophication, even if the total flux of carbon remain constant or even
S.L. Nielsen, M.F. Pedersen and G.T. Banta
decreases slightly. Whether or not nutrient fluxes from primary producers to
herbivores vary systematically across systems dominated by different plant groups
warrants further investigation in the future. It is also unclear whether higher nutrient
fluxes to herbivores will lead to higher rates of mineralization - grazing may either
lead to mineralization and thus increase nutrient availability that can support primary
production or herbivores may assimilate and store nutrients as biomass. Storage of
nutrients in biomass is of course temporary, unless the herbivores in question are
economically important and therefore removed from the system by harvest (e.g.
mussels, oysters, prawns) or migrate themselves out of the ecosystem. Little is known
about the balance between mineralization and assimilation but a larger proportion of
nutrients should, other things being equal, be mineralized when nutrient-rich food
items (i.e. ephemeral macroalgae and microalgae) are consumed. More studies that
specifically address these questions are needed before firm conclusions regarding the
quantitative role of grazing on different plant types on overall nutrient cycling can be
2.3 Decomposition and permanent burial
Decomposition of plant litter represents yet another process that may affect the
balance between accumulation and recycling of nutrients within shallow coastal
ecosystems. Banta et al. show in Chapter 7 that patterns of decomposition differ
markedly between different groups of primary producers although decomposition
patterns do vary substantially within each group. Slow-growing macrophytes
(especially seagrasses) decompose more slowly and less completely when compared
to most macro- and microalgae. Especially important seems to be the proportion of
the detritus that is resistant to fast decomposition (i.e. the refractory component) –
this fraction is typically much larger in seagrasses, and especially so in their roots and
rhizomes, than in marine macro- and microalgae. Surprisingly, detritus from
phytoplankton also contain a significant refractory portion, although this observation
may be in part an artifact due to methodological differences (i.e., litter-bag studies not
being well-suited for the study of phytoplankton decomposition). The mineralization
rates and patterns of N and P (relative to carbon decomposition) appear to be
controlled by the initial nutrient status of the litter with nutrients that are limiting for
decomposer organisms (usually N) being preferentially retained relative to carbon
and non-limiting nutrients.
Systematic variations in decomposition and mineralization rates and in the amount of
refractory detritus among different plant groups may have important consequences
for the retention of detritus and associated nutrients at the ecosystem level. Low rates
of decomposition lead to long residence times for litter in the detritus pool, which
suggests that detritus and associated nutrients from slow-growing macrophytes are
more prone to permanent burial than detritus from fast-growing plants. Permanent
burial is defined as the long-term (annual to decades) removal of nutrients from the
pelagic component by accumulation of sediments (Middelburg et al., Chapter 8).
Comparative studies based on carbon flows have shown that 10-17% of the net
primary production from rooted, slow-growing macrophytes are buried whereas much
less (about 1-5%) of the net primary production of algae are buried permanently.
Although this difference may seem significant, it is worth remembering that the
difference may be much smaller when expressed in units of nutrients (N or P)
because slow-growing macrophytes typically contain fewer nutrients per unit biomass
than fast-growing algae.
Banta et al. present such a comparison of nutrient budgets for estuaries dominated by
different primary producers at the end of Chapter 7. Their model simulations based
on data from two shallow estuaries show that accumulation of detritus-bound
nutrients is markedly higher in systems dominated by slow-growing macrophytes
than in systems dominated by fast-growing macro- and microalgae. This is because
total primary production is larger in systems dominated by the former plant group and
because a larger proportion of that production ends up as detritus. Furthermore, the
turnover rate of that detritus is much slower and less complete than for fast-growing
plant types resulting in greater nutrient retention. Even when seagrasses dominate,
however, permanent nutrient retention in detrital pools is only a minor fraction (about
10%) of the nutrients entering such systems. This retained fraction falls towards zero
when faster growing plants dominate the community.
One would immediately think that the main contribution of local primary producers
to permanent burial in an ecosystem would be their own biomass, or rather,
necromass. This is not necessarily the case, however. Studies using stable carbon
isotopes have shown that a large proportion of the carbon that accumulates in many
shallow coastal ecosystems is in fact of allochthonous (i.e., external) origin. The main
contribution of local primary producers to permanent burial of plant detritus and
associated nutrients may therefore rather be to reduce resuspension, dampen water
movement and, thus, stimulate net sedimentation of allochthonous particulate matter.
This process may have little effect on nutrient cycling in coastal ecosystems as these
nutrients were already assimilated in detrital pools before they entered the system and
were therefore unavailable for plant uptake.
Benthic microalgae stabilize sediments through exudation of polymeric substances
that may lower erodability and help keep detritus particles bound to the sediment
surface. This effect is, however, seasonal and organic matter “trapped” by these
polymeric substances may become resuspended later. The role of benthic microalgae
for burial of organic matter and associated nutrients is therefore most probably of
minor importance over annual time scales. Perennial, rooted macrophytes, on the
other hand, have a stronger and longer-lasting stabilizing effect on the sediment. In
addition, rooted macrophytes increase sedimentation of suspended particulate matter.
These effects of reducing resuspension and increasing deposition lead to enhanced
accumulation of sediments including organic matter and associated nutrients within
plant meadows. Structural changes in the plant community that lead to the loss of
rooted macrophytes may therefore have important consequences for the long-term
retention of major nutrients through burial and preservation.
2.4 Benthic nutrient cycling
Risgaard-Petersen and McGlathery et al. provide excellent analyses of the potential
effects of different plant groups on benthic nutrient cycles in Chapters 9 and 10.
Benthic processes are important because they affect the concentration of porewater
nutrients and, thus, the benthic-pelagic flux of nutrients. The flux of nutrients to the
water phase may, in the long run, determine whether nutrients released through
decomposition remain “trapped” in the sediment and thus become prone to permanent
burial or, whether they are released again to the overlying water phase and become
S.L. Nielsen, M.F. Pedersen and G.T. Banta
available for assimilation and/or export once again. Benthic processes may finally
influence whether nitrogen becomes lost permanently from the ecosystem through the
process of denitrification. Four processes seem especially important for regulating the
availability of nutrients in sediments: 1) mineralization, where dissolved inorganic
nutrients are “produced” from organic detritus, 2) coupled nitrification and
denitrification where ammonia is transformed via nitrate to free N2 that is lost to the
atmosphere, 3) N-fixation, by which gaseous nitrogen is incorporated into organic
matter and thereby made biologically available and 4) chemical binding of P to iron
ferrihydroxides and in other complexes that may reduce porewater concentrations of
dissolved P and fluxes to the water phase and may stimulate the permanent burial of
P. Risgaard-Petersen and McGlathery et al. (Chapters 9 and 10) show through a large
number of examples that most benthic, marine plant groups can affect these
processes, but the question is whether some plant types assert stronger effects than
others and, what the net effects are.
Mineralization is typically larger when slow-growing, perennial plants dominate
because absolute production, deposition (from allocthonous sources) and
decomposition of detritus is larger per unit area and time in systems dominated by
these plant types, despite the fact that mineralization rates per unit organic matter are
lower (Banta et al. Chapter 7). Rooted macrophytes may further stimulate
decomposition of slowly decomposable detritus through leakage of oxygen in the
rhizosphere and rates of mineralization are therefore often higher in sediments
covered by seagrasses when compared to bare sediments. This does not necessarily
mean that the flux of nutrients to the water is higher in seagrass meadows, however,
because mineralization is often closely coupled to assimilation by plant roots in
productive seagrass beds. A close coupling between mineralization and assimilation
also exists in dense populations of benthic microalgae and within dense mats of
floating macroalgae although rates are typically lower.
Nitrification and denitrification affect the concentrations of N species and may lead to
permanent losses of nitrogen from the ecosystem. Both nitrification and
denitrification are profoundly affected by the presence of plants. Several studies have
shown that rooted plants and benthic microalgae both can stimulate nitrification
through exudation of oxygen to anoxic sediments. It has been suggested that the
presence of seagrasses should stimulate coupled nitrification-denitrification, but more
recent studies have shown that this is rarely the case. Rates of coupled nitrificationdenitrification are in fact typically lower in areas vegetated by benthic microalgae and
seagrasses due to plant uptake. Both benthic microalgae and seagrasses seem to be
better competitors for ammonium than nitrifying bacteria and coupled nitrificationdenitrification may thus be reduced by 50-75% in the presence of benthic microalgae
and seagrasses. A positive effect of benthic microalgae and seagrasses on coupled
nitrification-denitrification may, however, occur under N-replete conditions where
plants and nitrifiers do not compete for ammonium.
Little is known about the effect of macroalgae on nitrification-denitrification but the
presence of dense mats of floating macroalgae often leads to anoxic conditions in the
sediment and the zone of nitrification/denitrification may move up into the algal mat
temporarily. In such cases, nitrification-denitrification is usually reduced again due to
competition for nutrient uptake with the macroalgae
There is little consensus about the potential effects of different primary producers on
N-fixation. Some studies have shown that N-fixation rates can be substantially higher
in sediments vegetated by seagrasses while other studies have not been able to
demonstrate such an effect. The relatively few reported rates on N-fixation in
seagrass beds show that rates are approximately within the same range as rates of
denitrification. Rates of N-fixation and denitrification are, however, relatively low
when compared to other N-transport rates in shallow coastal ecosystems, so the
effects of seagrasses on these processes have relatively little ecological significance.
The release of oxygen to the sediment by benthic microalgae and rooted plants, or the
depletion of oxygen by dense mats of floating macroalgae or benthic microalgae (in
the dark), may dramatically affect redox conditions and thus benthic P-dynamics.
Stimulation of oxidized conditions in the uppermost sediment favors the binding of P
to oxidized iron compounds and thus lowers both the concentrations of dissolved P in
the porewater and reduces the flux of P to the water column. In contrast, lower redox
conditions during oxygen depletion have the opposite effect and may stimulate the
release of P to the water phase.
The strongest effect that plants may exert on benthic nutrient dynamics may be
through their direct interference with nutrient fluxes across the sediment-water
interface. Seagrasses and benthic microalgae stabilize the sediment and thus reduce
the release of sediment nutrients to the water via resuspension events. Both these
plant groups are further able to intercept the flux of dissolved nutrients from the
sediment to the water through their own assimilation. Especially benthic microalgae
seem potentially important in this respect, since dense mats of these algae can prevent
benthic-pelagic fluxes of nutrients completely. The efficiency of benthic microalgae
to assimilate nutrients may vary substantially, however, over diurnal and longer time
scales since the community must be net autotrophic in order to be able to assimilate
enough nutrients to prevent diffusive transport of nutrients to the water. There are
many temporarily variable factors that affect benthic algal production and thus their
ability to assimilate nutrients and function as a filter at the sediment-water interface.
The temporal and spatial variations in the importance of these algae for nutrient
release require further studies. Seagrasses may also reduce the flux to the water and
lower porewater nutrient concentrations through assimilative uptake of nutrients via
the roots, but again, no studies allow a direct comparison between the effect of
seagrasses and other plant types. Finally, dense mats of floating macroalgae may have
the same nutrient filtering effect as benthic macroalgae. Experimental studies have
shown that dense mats of macroalgae are able to adsorb nutrients released from
sediments completely, but such mats are often highly “ephemeral” in time and space,
and therefore most probably not very efficient filters when evaluated over larger
temporal and spatial scales.
The different marine plant groups each have the potential to influence benthic
nutrient cycling though a number of mechanisms, but quantitative studies that allow
direct comparisons of the net effects of theses different plant types on an ecosystem
basis are sorely lacking.
S.L. Nielsen, M.F. Pedersen and G.T. Banta
When we first embarked on the work behind this book (see Chapter 1), we were
rather convinced that the observed changes in dominance by different plant types,
observed as a consequence of coastal eutrophication, would lead to predictable
changes in the cycling and ultimately the fate of nutrients in these ecosystems. We
were also convinced that these relationships could be used to predict changes in
cycling and fate of nutrients caused by changes in the composition of coastal marine
primary producers brought about by other factors than eutrophication. After our own
studies and with the contributions of our colleagues, we realize that the world may be
slightly more complicated. We had hoped to be able to give typical rates for the
various main processes (export, grazing, decomposition etc.) under dominance by
seagrasses, slow and fast growing macroalgae and phytoplankton, but this is hardly
possible, due to a lack of data. Instead, we can attempt to assess the constancy and
importance of these processes. Mass transport of plant-bound nutrients out of the
estuarine systems (export) is highly variable between different systems, but can
undoubtedly be of huge importance in systems where export of plant-bound nutrients
constitute a large percentage of the total nutrient export. Unfortunately, comparative
studies where export is measured together with primary production are largely absent
and it is therefore difficult, if not impossible, to assess the significance of export of
plant-bound nutrients relative to the primary production. It is equally difficult to
conclude on how export may change as a consequence of changes in the plant
community. Grazing has been shown to be of a constant magnitude, when measured
in biomass or carbon units. As ephemeral macroalgae and phytoplankton have higher
contents of nutrients in their tissue than seagrasses and slow-growing macroalgae,
however, this should mean that the amount of nutrients going through the grazingfood chain will increase if ephemeral macroalgae or phytoplankton replace seagrasses
or fucoid macroalgae. Unfortunately, data to confirm this are largely lacking.
Probably most surprising to us were the results relating to the decomposition of
various plant types. That below-ground parts of seagrasses decomposed slowly and
with a large refractory component was to be expected, but that phytoplankton also
had a large refractory component was more surprising, given their chemical
composition, although this may be a experimental artifact and needs further
consideration. Large differences between different types of macroalgae existed, but
their overall contribution to nutrient retention via refractory pools were minor.
Refractory detrital pools can ultimately give rise to a loss of nutrients from the system
through permanent burial, and indeed data show that the percentage of the primary
production that is lost to permanent burial is significantly larger when the system is
dominated by slow growing primary producers. However, the percentage lost to
permanent burial is always a minor fraction of the primary production and as such of
minor significance in overall nutrient budgets. Another loss process is denitrification.
Data indicate that denitrification varies with several orders of magnitude between
different estuarine systems, and that primary producers have a tendency to lower the
rates of denitrification through competition with the bacteria for substrate (i.e., NO3-).
Unfortunately data do not allow comparisons between systems dominated by
different types of primary producers to allow us to asses the importance of this loss
term in the context of now marine plants alter nutrient cycling
At the onset of this work, we were convinced that the composition of the plant
community was paramount in determining cycling and fate of nutrients in the coastal
marine ecosystems. This work has shown, however, that other factors are of equal
importance. Most important of these factors are undoubtedly the hydrography and the
nutrient loading of the system in question. Hydrography (morphometry, water
residence time) can be decisive in determining which plant type will dominate, the
degree of export of plant-bound nutrients and the importance of various grazers.
Nutrient loading in itself is probably of equal significance as the plants in determining
the various fluxes as well as the ultimate fate of the nutrients. However, the plant
component of coastal ecosystems is still very important as virtually all nutrients pass
through the primary producers at least once on their way from the terrestrial to the
oceanic ecosystems. On the other hand, the role of the estuary as a filter is
questionable. Our data show that from 1 to 17 % of the primary production becomes
permanently buried, depending on the primary producers that dominate. These
numbers, however, assume steady-state conditions, and the question is to which
extent steady-state conditions prevail in costal marine ecosystems, that could be seen
as going through cycles of eutrophication or other detrimental anthropogenic
influence, followed by restoration initiatives.
As it can be inferred from the above, many data are still lacking, but we are now in a
better situation than before, as we are now able to describe and model the individual
processes involved in coastal nutrient cycling (the arrows in our conceptual model,
Fig. 4 in Chapter 1) mechanistically correct. Models of these individual processes
should now be synthesized in dynamic ecosystem models and validated by empirical
studies in comparable ecosystems, where all individual processes are compared and
related to total primary production and the contributions from the different plant
types. In these studies nutrient loading and hydrography should be included as
important controlling parameters.
Agustí, S., Enríquez, S., Frost-Christensen, H., Sand-Jensen, K., & Duarte, C. M. (1994). Light harvesting
among photosynthetic organisms. Functional Ecology, 8, 273-279.
Bergamasco, A., De Nat, L., Flindt, M.R. & Amos, C.L. (2003). Interactions and feedbacks among
phytobenthos, hydrodynamics, nutrient cycling and sediment transport in estuarine ecosystems.
Continental Shelf Research 23, 1715-1741.
Burkholder, J. M., Mason, K. M., & Glasgow, H. B. (1992). Water-column nitrate enrichment promotes
decline of eelgrass Zostera marina: evidence from seasonal mesocosm experiments. Marine
Ecology Progress Series, 81, 163-178.
Cebrián, J., & Duarte, C. M. (1994). The dependence of herbivory on growth rate in natural plant
communities. Functional Ecology, 8, 518-525.
Cebrián, J., & Duarte, C. M. (1995). Plant growth-rate dependence of detrital carbon storage in ecosystems.
Science, 268, 1606-1608.
Cebrián, J., & Duarte, C. M. (1998). Patterns in leaf herbivory on seagrasses. Aquatic Botany, 60, 67-82.
Cloern, J. E. (2001). Our evolving conceptual model of the coastal eutrophication problem. Marine
Ecology-Progress Series, 210, 223-253.
Duarte, C. M. (1995). Submerged aquatic vegetation in relation to different nutrient regimes. Ophelia, 41,
Duarte, C. M., Benavent, E., & Sanchez, M. d. C. (1999). The microcosm of particles within seagrass
Posidonia oceanica canopies. Marine Ecology Progress Series, 181, 289-295.
Enríquez, S., Duarte, C. M., & Sand-Jensen, K. (1993). Patterns in decomposition rates among
photosynthetic organisms: The importance of detritus C:N:P content. Oecologia, 94, 457-471.
S.L. Nielsen, M.F. Pedersen and G.T. Banta
Greve, T. M., Borum, J., & Pedersen, O. (2003). Meristematic oxygen variability in eelgrass (Zostera
marina). Limnology and Oceanography, 48(1), 210-216.
Hauxwell, J., Cebrían, J., & Valiela, I. (2003). Eelgrass Zostera marina loss in temperate estuaries:
relationship to land-derived nitrogen loads and effect of light limitation imposed by algae.
Marine Ecology Progress Series, 247, 59-73.
Holmer, M., & Bondgaard, E. J. (2001). Photosynthetic and growth response of eelgrass to low oxygen and
high sulfide concentrations during hypoxic events. Aquatic Botany, 70, 29-38.
Kennedy, V. S. (Ed.). (1984). The estuary as a filter. Orlando: Academic Press.
Krause-Jensen, D., McGlathery, K., Rysgaard, S., & Christensen, P. B. (1996). Production within dense
mats of the filamentous macroalga Chaetomorpha linum in relation to light and nutrient
availability. Marine Ecology Progress Series, 134, 207-216.
Littler, M. M., & Littler, D. S. (1980). The evolution of thallus form and survival strategies in benthic
marine macroalgae: Field and laboratory tests of a functional form model. American Nauralistt,
116(1), 25-44.
Nielsen, S. L., Enríquez, S., Duarte, C. M., & Sand-Jensen, K. (1996). Scaling maximum growth rates
across photosynthetic organisms. Functional Ecology, 10, 167-175.
Nielsen, S. L., & Sand-Jensen, K. (1990). Allometric scaling of maximal photosynthetic growth rate to
surface/volume ratio. Limnology and Oceanography, 35, 177-181.
Pedersen, M. F., & Borum, J. (1996). Nutrient control of algal growth in estuarine waters. Nutrient
limitation and the importance of nitrogen requirements and nitrogen storage among
phytoplankton and species of macroalgae. Marine Ecology Progress Series, 142, 261-272.
Pedersen, O., Borum, J., Duarte, C. M., & Fortes, M. D. (1998). Oxygen dynamics in the rhizosphere of
Cymodocea rotundata. Marine Ecology Progress Series, 169, 283-288.
Pedersen, O. B., Christiansen, C., & Laursen, M. B. (1995). Wind-induced long term increase and short
term fluctuations of shallow water suspended matter and nutrient concentrations, Ringkøbing
Fjord, Denmark. Ophelia, 41, 273-287.
Salomonsen, J., Flindt, M.R., Geertz-Hansen, O. & Johansen, C. (1999). Modelling advective transport of
Ulva lactuca (L) in the sheltered bay, Møllekrogen, Roskilde Fjord, Denmark. Hydrobiologia
397, 241-252.
Sand-Jensen, K., & Borum, J. (1991). Interactions among phytoplankton, periphyton, and macrophytes in
temperate freshwaters and estuaries. Aquatic Botany, 41(1-3), 137-175.
Schramm, W., & Nienhuis, P. H. (1996). Introduction. In W. Schramm & P. H. Nienhuis (Eds.), Marine
benthic vegetation. Recent changes and the effects of eutrophication. (pp. 1-6). Berlis,
Heidelberg, New York: Springer.
Terrados, J., Duarte, C., Fortes, M., Borum, J., Agawin, N., Bach, S., et al. (1998). Changes in community
structure and biomass of seagrass communities along gradients of siltation in SE Asia.
Estuarine, Coastal and Shelf Science, 46, 757-768.
Touchette, B. W., & Burkholder, J. M. (2002). Seasonal variations in carbon and nitrogen constituents in
eelgrass (Zostera marina L.) as influenced by increased temperature and water-column nitrate.
Botanica Marina, 45, 23-34.
Worm, B., L., L. H., Boström, C., Engkvist, R., Labanauskas, V., & Sommer, U. (1999). Marine diversity
shift linked to interactions among grazers, nutrients and propagule banks. Marine Ecology
Progress Series, 185, 309-314.
S.L. Nielsen, M.F. Pedersen & G.T. Banta: Department of Life Sciences and
Chemistry, Roskilde University. P.O. Box 260, DK-4000 Roskilde, Denmark
Advective transport;2;8;13;17;31;112;146;159;345
Algal blooms;77;83;103;154;190;193;196;197;207;253
Algal mats;90;93;94;195;197;200;208;209;264;302;304;340
Amino acids;69;282;286
Anthropogenic influence;76;344
Benthic diatoms;62;258;266;291;294;305
Bottom-up control;153
Brown tide;78;88;104;106;296
Carrying capacity;70;151;160;161;163;171;172
Catchment area;5
Clearance rate;149;151;165
Clearance time;162;163;164;336
Colonial algae;37
Dissolved Inorganic Carbon;279;283;291;295
Dissolved Inorganic Nitrogen;133;134;141;142;282;283;286;289;291;292;296;321
Dissolved Organic Carbon;290;292;305
Dissolved Organic Matter;273;274;275;276;290;291;292;293
Dissolved Organic Nitrogen;138;290;291;292;306
Dry weight;5;201;215;217;227;240
Faecal pellets;308
Filter effect;284;286
Filter feeders;149;172;173;174;333;336
Food limitation;151;160
Gene flow;24
Genetic diversity;25
Ghost tissue;119;123
Green tides;76;101
Hydrodynamic forces;124
Labile matter;215
Land use;5;83;177
Light attenuation;47;49;50;51;56;58;63;88;89;90;92;100;177;289;294
Linear dimension;27;28;30;40
Multi-G models;215;216
Net Primary Production;291
Nitrogen fixation;69
Nutrient assimilation;6;7;215;253;271;283;290;322
Nutrient retention;10;144;218;219;230;234;240;241;242;248;289;295;331;338;343
Nutrient supply;20;26;33;40;76;153;161;288;332
Organic acids;11;56;276
Particle load;149
Perennial macroalgae;215;234;240;270;331;332;335;336
Photic zone;31;48;49;50;51;54;57;270;281
Physical mixing;156;166;170
Refractory pool of detritus
Relative surface area;1
Reynolds number;30
Rooted macrophytes;vii;1;7;10;11;110;112;131;242;289;339
Salt marshes;68;106;135;142;143;144;146;258;260;268;298
Sampling strategy;108
Sea urchins;98;103;105;178;181;183;206;207;210;211
Seagrass decline;75;76;77;83;85;98;177;187;189;195;297
secondary production;150;160;174;195;202;204;295
Sediment-water interface;10;59;274;281;283;286;300;305;341
Sinking rate;2;8
Terrestrial plants;24;43;44;215;276
Threshold velocity;109;114;116
Top-down control;164;178;179;336
Trophic levels;172;290;292;335;336;337
Unicellular algae;1;23;37;43;44;59
Water motion;22;26;30;61
Water residence time;6;75;77;83;85;87;97;161;162;163;164;271;295;333;336;344
AQUATIC ECOLOGY BOOK SERIES,11855,4-40109-69-33114010-0,00.html
Series Editor:
Prof. Robert G. Wetzel, Department of Environmental Sciences and Engineering,
The University of North Carolina, Chapel Hill, North Carolina 27599-7431, USA.
E-mail: [email protected]
Current volumes in this series
Volume 1
Ecology of a Glacial Flood Plain
J.V. Ward, U. Uehlinger
Hardbound, ISBN 1-4020-1792-8, December 2003, 320 pp.
Volume 2
The Influence of Primary Producers on Estuarine Nutrient
S.L. Nielsen, G.T. Banta, M.F. Pedersen
Hardbound, ISBN 1-4020-2638-2, 2004
Forthcoming volumes:
Harmful Cyanobacteria
J. Huisman, H. C.P. Matthijs, P. M. Visser
Ecological History of the Lowland River Basins of Rhine and
P.H. Nienhuis
Ecology of Russian Rivers
A.V. Zhulidov, V.V. Khlobystov, R.D. Robarts, A.A. Ischenko
Dr. J. Headley
Biology of Polluted Rivers
T.E.L. Langford
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