Marine eutrophication and benthos - BES journal - Wiley

1 downloads 0 Views 342KB Size Report
fl 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830. 814 JACQUES ..... tions, mass mortality of some benthic species generally occurs ..... of nutrients–such as C (Dame & Patten, 1981; Doering ..... North Sea–Estuaries Interacions (eds Mclusky DS, de Jonge V .... Graneli E), Elsevier Science Publishing Co,.
Global Change Biology (2002) 8, 813±830

Marine eutrophication and benthos: the need for new approaches and concepts J A C Q U E S G R A L L and L A U R E N T C H A U V A U D Institut Universitaire EuropeÂen de la Mer, LEMAR UMR-CNRS 6539, Place Copernic, 29280, Plouzane, Brittany, France

Abstract In this review, using examples drawn from field observations or experimental studies, our goals are (i) to briefly summarize the major changes, in terms of species composition and functional structure, occurring in phyto and zoobenthic communities in relation to nutrient enrichment of the ecosystems; particular interest is given to the major abiotic and biotic factors occurring during the eutrophication process, (ii) to discuss the direct and indirect influences of benthic organisms on eutrophication and whether the latter can be controlled or favoured by benthos; most benthic species play a major role in the process of benthic nutrient regeneration, affecting primary production by supplying nutrients directly and enhancing rates of pelagic recycling; experimental studies have shown that the impact of benthic fauna on benthic±pelagic coupling and nutrient release is considerable. Thus, once the eutrophication process is engagedÐthat is, high organic matter sedimentationÐit may be indirectly favoured by benthic organisms; benthos should always be considered in eutrophication studies, (iii) to evaluate the limits of our observations and data, highlighting the strong need for integrated studies leading to new concepts. Coastal ecosystems and benthic communities are potentially impacted by numerous human activities (demersal fishing, toxic contaminants, aquaculture. . .); in order to design strategies of ecosystem restoration or rehabilitation, we have to better understand coastal eutrophication and develop tools for quantifying the impacts; in order to achieve this goal, some possible directions proposed are: integrated studies leading to new concepts, model development based on these concepts and finally comparison of various ecosystems on a global scale. Keywords: benthos, eutrophication, nutrients and carbon fluxes, phytobenthos, suspension feeders, zoobenthos Received 13 September 2001; revised version received and accepted 15 February 2002

Introduction Marine eutrophication of estuaries or coastal waters is considered to be a direct result of increasing population densities along coastlines and/or the use of fertilizers for agriculture (Cloern, 2001). Higher nutrient loading in coastal waters leads to increased primary production (either pelagic or benthic) and results in organic matter input to the sediment that is directly available for heterotrophic organismsÐbacteria and meio and macrofaunal grazers or detritus feedersÐ(Heip, 1995). In this paper, an increase in the supply rate of organic matter to the benthos is termed benthic eutrophication (Nixon, 1995). Correspondence: Jacques Grall, fax ‡33 2 98 49 86 45, e-mail: [email protected] ß 2002 Blackwell Science Ltd

However, organic inputs to the bottom may have many various sources and the organic input does not necessarily originate from in situ primary production. Sediments may contain organic matter that has both autochthonous originsÐthat is, organic matter derived from fixation by in situ primary producersÐor allochthonous originsÐ organic matter not originating from the benthosÐ(Paerl et al., 1998). Benthic communities are highly sensitive to eutrophication and hypoxia (Jorgensen & Richardson, 1996); therefore, benthic species have been used in several studies as indicators of organic enrichment (Pearson & Rosenberg, 1978; Majeed, 1987; Grall & GleÂmarec, 1997; Mucha & Costa, 1999). The extreme effects of eutrophication, hypoxia and subsequent mortality of most organisms are well-described (Gray, 1992). However, in order 813

814 J A C Q U E S G R A L L & L A U R E N T C H A U V A U D to monitor the health of coastal ecosystems it is crucial that we are able to identify and measure what are the initial effects of eutrophic stress on coastal and estuarine benthic communities (Gray et al., 1990). The wide variety of functional roles, feeding types and sizes of benthic organisms in these communities makes it difficult to predict quantitatively and accurately the consequences of eutrophication (Bonsdorff & Pearson, 1999). In addition, the sources of organic matter in benthic food chains must also be described and quantified (Heip, 1995) along with hydrologic parametersÐsalinity, temperature, tidal currentsÐall of which can vary between estuaries and bays. Other important factors structuring benthic communities include potential anthropogenic impacts on the benthic systemÐsuch as bottom fishing, sewage dumping and invasive species. Benthic eutrophication is a complex problem and we will attempt neither to describe the detailed processes involved here nor to give a comprehensive review of the problem because this has been done by several authors (Gray, 1992; Diaz & Rosenberg, 1995; Heip, 1995). Using examples from field observations and laboratory experiments, we propose to examine the state of the knowledge on the effects of eutrophication on coastal and estuarine benthic communities. Our objectives are three-fold in this study: to briefly have an overview of the major changes of species composition and functional structure that occur in benthic communities in relation to organic matter inputs to the sediment; to assess the direct and indirect influences of benthic organisms on eutrophication and whether eutrophication can be controlled or favoured by benthos and to evaluate the limits of our observations and data, highlighting the need for integrated studies leading to new concepts.

Eutrophication and benthos dynamics Eutrophication can have both positive and negative effects on phyto and zoobenthic communities. Increasing nutrient concentrations can stimulate benthic primary production, resulting in higher food availability for grazers, whereas increased sedimentation of pelagic production benefits filter-feeding and deposit-feeding macrobenthos. However, increased sedimentation of organic matter is harmful to some benthic fauna through siltation, habitat modification and oxygen (O) depletion caused by higher decomposition rates.

Microphytobenthos and eutrophication Benthic microalgal production has been studied extensively in intertidal and estuarine habitat. Nutrient concentrations are not controlling microphytobenthic production (Sundbaeck & Joensson, 1988; Paterson &

Crawford, 1986; Kromkamp et al., 1995) because other environmental parametersÐlight, sediment parameters, space availabilityÐare more much limiting factors (Collos, 1987; Rizzo, 1990; Irigoien & Castel, 1997). However, benthic primary production reduces nitrate benthic fluxes of 20% (Graneli & Sundbaeck 1985; Rizzo, 1990) and a positive relationship between benthic microalgae biomass increase and increased nutrients loading to coastal water has been established (de Jonge 1990; Philippart & CadeÂe, 2000).

Macrophytobenthos and eutrophication Macroalgal blooms have been linked to a variety of conditions in eutrophic coastal areas around the globe. The increased import of nutrients increases macrophyte biomass (CadeÂe, 1984), but is accompanied by changes in macrophytobenthic species composition, distribution, and thus can permit growth and dominance of the ecosystem by a relatively small number of taxa (Valiela et al., 1997). We summarize the abiotic and biotic factors affecting macrophytobenthic communities in the sections below.

Abiotic factors controlling macroalgal blooms Nutrient availability is the major factor controlling macroalgal blooms. Soft, delicate, annual seaweed species have higher rates of nutrient uptake than perennial, long-lived species (Kautsky, 1991) and they rapidly assimilate the higher amount of nutrients available under eutrophic conditions (Wallentinus, 1984). Thus nutrient enrichment may lead to the loss of submerged coastal macrovegetation as the result of shading by rapidly proliferating epibionts (Burkholder et al., 1994; Heck et al., 2000). Changes in nutrient concentration increase the amount of vegetation biomass and modify competition between species, creating conditions that favour annual opportunistic species over perennial forms (Baden et al., 1990) (Fig. 1). The chemical composition of the dissolved and particulate organic matter in the ecosystem has been investigated for elements that may limit macroalgal growth. Nitrogen (N) was previously thought to limit macroalgal growth in the sea (Odum, 1970; Ryther & Dunstan, 1971; Burkholder et al., 1994; Valiela et al., 1997) and phosphorus (P) has also been considered as a limiting nutrient (Piriou & Menesguen, 1992; Lapointe, 1999). Other essential elementsÐsuch as iron (Fe)Ðmay also be involved (Taylor et al., 1995). By contrast, Smith (1984) argued that considering a single nutrient as the limiting factor is an unacceptable oversimplification. More recent studies have argued that the interactive effects of N and P were greater than those of P alone (Hughes et al., 2000), and ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

M A R I N E E U T R O P H I C A T I O N A N D B E N T H O S : A R E V I E W 815

Fig. 1 Schematic presentation of the changes occurring in the dominance of primary producers during increasing phases of eutrophication. (After Schramm, 1999).

Lapointe (1999) introduced the `nutrients set threshold' concept for macroalgal blooms. Nutrient concentrations and turnover rates in coastal water are influenced by hydrologic conditionsÐsuch as temperature, water residence time (Piriou & Menesguen, 1992), tidal or wind-forced currents (Burkholder et al., 1994) and riverine fluxes (Schramm, 1999). Shallow-water bays with long water residence times are particularly susceptible to developing macroalgal blooms when nutrient loads increase (Schramm, 1999). Strong water currents usually prevent eutrophication by flushing nutrients out of the ecosystem. There are some sitesÐsuch as the Bay of St BrieucÐwhere nutrients can accumulate, despite the presence of strong currents, and conditions favourable to macroalgal blooms are created (Piriou & Menesguen, 1992). In general, eutrophication effects are more severe in habitats with reduced flushing, where temperature fluctuates more and nutrient loadings are frequent and concentrated, than in areas with shorter water residence times (Burkholder et al., 1994). Light becomes the controlling factor for macroalgal bloom development as nutrient loads increase. High pelagic biomass produces cell-shading, resulting in decreased light penetration. Decreasing the light intensity affects the species composition and depth distribution of macrophyte assemblages (Valiela et al., 1997). Macroalgae are initially outcompeted by phytoplankton or ephemeral macroalgae; this leads to shifts in the species composition and to decreases in the depth distribution of ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

the macrophytes in response to their minimum light requirement (Taylor et al., 1995). Shifts from seagrass to ephemeral algae occur because perennial benthic macrophyte production is light-limited, whereas phytoplankton production and opportunistic macroalgal production are nutrient-limited (McClelland & Valiela, 1998). In cases of high eutrophication, the very high amounts of pelagic production and subsequent seston sedimentation decrease incident light to levels below the requirement of all benthic algal species. This results in the disappearance of seaweeds from the bottom (Valiela et al., 1997; Schramm, 1999) and thus primary production occurs only in the pelagic environment (Fig. 1).

Biotic factors controlling macroalgal blooms Small invertebrate mesograzers control epiphytic algal abundance and may even prevent seagrass decline (Heck et al., 2000). Lapointe (1999) argued that decreasing grazing pressure alone cannot explain heavy macrophyte production. They showed that this phenomenon is controlled by nutrient concentrations rather than by large herbivorous benthic species. In any case, a better understanding of the factors controlling grazer abundance is required in order to fully understand the changes in macroalgal populations under eutrophication (Heck et al., 2000). Grazers are more likely to consume soft, ephemeral algal species than perennial speciesÐ such as seagrasses. Preferential consumption of the former may be a factor maintaining dominance of

816 J A C Q U E S G R A L L & L A U R E N T C H A U V A U D seagrasses in waters under relatively moderate nutrient inputs. However, with increasing nutrient concentrations competitive advantages and grazer abundance are likely to change (Valiela et al., 1997). Another animal±plant interaction in the eutrophication process is competition for space; increased primary production may support the development of dense benthic suspension-feeder populations; that is, barnacles or mussel, and high densities of these animals may strongly affect recruitment of the seaweeds to the substratum (Everett, 1991). Thus, the effects of eutrophication on macroalgae are complex and largely site-specific (Valiela et al., 1997; Schramm, 1999). Eutrophic algal communities modify habitat characteristics (Valiela et al., 1997) and have a strong influence on macroalgae-associated fauna as many invertebrates depend on vegetation for substratum, food or shelter (Gee & Warwick, 1994).

Zoobenthos and organic enrichment Natural or anthropogenic-derived organic inputs have the same effects on zoobenthos (Weston, 1990). The quality and the condition of the organic particles (either sedimenting particulate organic matter, phytoplankton or macroalgal detritus) are apparently more important for determining how various processes in the benthic ecosystem are favoured under eutrophication.

Macroalgal blooms and zoobenthos At the outset of the eutrophication process, increased macroalgal biomass may increase available living space, food and shelter from predation, and as a result increase faunal diversity and abundance (Gee & Warwick, 1994). However, fauna associated with macroalgal blooms may already differ from faunal assemblages of nonbloom conditions (McClelland & Valiela, 1998). Changes in population structure have been studied by Lillebo et al. (1999) along an eutrophication gradient. The authors showed that the influence of macrovegetation abundance and quality on populations of gastropods and amphipods is to decrease production and cause shifts from K to r reproductive strategies of the dominant species of the benthic communities as eutrophication increases. To a certain extent, the increase of macroalgal biomass may favour small crustacean populations, but extensive blooms can lead to local extinctions (Pardal et al., 2000). Moreover, Valiela et al. (1997) argued that carbon fixed by soft, ephemeral algae would move through trophic webs faster than that fixed by seagrasses or vascular plants. Thus, under conditions of eutrophication the shift from perennial macrophytes to opportunistic species affects the ecosystem in a way that may

propagate changes in community assemblages throughout the entire food web.

Drifting algal mats and zoobenthos Delicate, ephemeral algal speciesÐwhich are described as more fragile than perennial species (Schramm, 1999)Ð can be more easily torn off from the substratum by strong currents, storms or swells and may result in massive drifting algal mats along beaches or shallow basins. These drifting mats have strong impacts on sediment zoobenthos, considerably altering the functioning of this normally uncovered bottom structure (Trush, 1986; Bonsdorff, 1992). In the sublittoral zone, high macrofaunal abundances have been recorded in drifting algal mats; these mats have been shown to function as an alternative habitat for mobile benthic macrofauna (Norkko et al., 2000). An experimental study by Hull (1987) showed an increase in abundance of some opportunistic species (Capitella capitata complex; Nereis diversicolor) and substantial changes in the amphipod Corophium volutator population size-structure. These observations coincided with siltation and the rapid appearance of anoxia under the mats. Field studies have shown that habitat is affected through high organic load, siltation and/or stratification of the sediment, which in turn affects population structure, production and recruitment of macrofauna (Bonsdorff, 1992; Peterson et al., 1994). These effects will depend on the spatial and temporal extent of the algal mats; a patchy occurrence may be beneficial to macrofaunal populations, whereas extensive or even total coverage impoverishes the species composition and structure of benthic communities (Holmquist, 1997; Raffaelli et al., 1998).

Particulate organic matter sedimentation and zoobenthos Correlation between the structure of macrofaunal communities (species richness, dominance and total abundance) and organic enrichment has been reported in numerous studies (Hily, 1983; Majeed, 1987; Weston, 1990). These observations led to useful empirical models that describe the spatial and temporal successions of macrofauna in relation to organic load of the sediment (Pearson & Rosenberg, 1978; Gray, 1982, 1992) (Fig. 2). Typically, the response to moderate eutrophication is an increase in benthic species richness, numerical abundance and biomass compared with background conditions. If organic loading increases, species richness decreases below background, and the benthic community assemblage shifts toward higher densities and total abundances of only a few opportunistic species. Biomass fluctuates according to the density of the opportunistic species populations (Fig. 2); Gray et al. (1990) ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

M A R I N E E U T R O P H I C A T I O N A N D B E N T H O S : A R E V I E W 817

Fig. 2 Pearson and Rosenberg's SAB (Species, Abundance, Biomass) model on effects of increased organic matter loadings on communities of benthic macrofauna. (After Gray, 1992).

hypothesized that after a disturbance some species take advantage of the changed environmental conditions and increase in abundance, to the detriment of less opportunistic species populations. In the final state of the descriptive model, very high concentrations of organic matter drive the benthos into an afaunal state (Ferraro et al., 1991). Although this model was supported by field evidence from long-term eutrophication studies (Rosenberg & MoÈller, 1979; Cederwall & Elmgren, 1980; Pearson et al., 1985; Reise et al., 1989; Gray et al., 1990; Beukema & CadeÂe, 1997), the results from experimental studies on the response of benthos to nutrient enrichment have been less convincing (Heip, 1995) although most of these studies exhibited at least some of the expected responses. Benthic community biomass responded weakly to nutrient enrichment in the MERL experimental ecosystem long-term study (Widbom & Elmgren, 1988), whereas other studies (Grassle et al., 1985; Hull, 1987) showed an increase in the abundance of opportunistic species after organic matter was added to the sediments. In general, a positive linear correlation is observed between nutrient inputs and productivity (Dauer & Conner, 1980; Nixon, 1992). The Pearson and Rosenberg model (1978) described other effects of organic enrichment on benthic populationsÐsuch as reduced mean size of individuals, shallower distribution of the fauna in the sediment column and an impoverishment of the functional structure of the community. Small-sized invertebrates have relatively large surface areas compared with their volume that would allow them to survive in hypoxic environments; ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

Weston (1990) and Gray et al. (1990) showed that interspecific measures of animal size actually decreased with increasing organic matter input. In addition, Weston (1990) also showed that intraspecific measures of animal size were positively correlated with organic enrichment, suggesting the potential for enhanced growth in the species capable of living under higher organic concentrations. This finding has been corroborated by several studies (Hily, 1987; Grall & Chauvaud, in prep). It has yet been impossible to relate accurately the response of the benthic system to a particular level of organic matter input (Heip et al., 1995). Local and temporal variations caused by abiotic factors (such as currents, tide, winds. . .) may disturb the relationships between organic matter supply and consumer abundance (Beukema & CadeÂe, 1997) because the increased amount of food supplied to the benthos may exert positive and negative effects. Further investigations, integrating sites-specific characteristics, are thus necessary to better understand the effects of a given quantity of organic matter on benthic communities. Most macrobenthic species are present at, or near, the surface of the sediment independently of the amount of organic enrichment. Thus, the shallower vertical distribution of the fauna in the sediment described by the Pearson±Rosenberg model actually reflects the absence of deep-burrowing species, an observation which would have little effect on the total abundance, but would change the biomass profile (Weston, 1990). In contrast to Weston's observation (1990) that there are no clear trends in the simplification of the trophic organization along an

818 J A C Q U E S G R A L L & L A U R E N T C H A U V A U D 100 suspensivores

% Polychaete biomass

herbivores

carnivores

surface deposit feeders subsurface deposit feeders

0 Increasing organic input Fig. 3 Relative dominance of the major polychaete trophic groups with increasing levels of organic matter input to the benthos (After Weston, 1998).

organic enrichment gradient, Weigelt (1991) observed that functional groups of nonselective deposit feeders were favoured over suspension feeders and selective deposit feeders (Fig. 3). These considerations suggest the need for experimental and long-term benthic eutrophication studies and operational models that would enable identification of the important processes in community modification during and after a eutrophication event.

Hypoxia/anoxia Sediment stratification and algal detritus decomposition raise the biological oxygen (O) demand and induce hypoxic or anoxic conditions (Norkko & Bonsdorff, 1996; Valiela et al., 1997) in poorly mixed waters. The dissolved O concentration is a function of temperature, salinity, the hydrodynamics of the ecosystem, and the quantity and quality of the deposited organic matter (Diaz & Rosenberg, 1995). Hypoxia directly affects metabolic processes by altering mobility, and indirectly by changing feeding and species interactions of the benthic organisms (Diaz & Rosenberg, 1995). As some benthic species rise from the sediment attempting to reach oxygenated waters, they become more vulnerable to predation, and therefore low-O conditions in the overlying waters are likely to facilitate transfer of benthic production to higher trophic levels (Nestlerode & Diaz, 1998). Even if many macrobenthic species survive short-term hypoxia through behavioural or physiological adaptations, mass mortality of some benthic species generally occurs, depending on the magnitude of the O depletion. Larger, long-lived species are eliminated first, then communities shift toward dominance by smaller, short-lived,

often opportunistic species in successive stages that depend on the frequency and intensity of hypoxia (Diaz & Rosenberg, 1995). The disappearance of functionally important speciesÐsuch as bioturbatorsÐfrom the benthic ecosystem may change the balance of the community dynamics, altering not only its functioning but also its composition in unforeseen ways. Oxygen concentrations, in the water and in the organisms, are therefore a major factor in the structuring and functioning of benthic communities. In the case of permanent anoxia, the macrofauna is eliminated from benthic communities. The resilience of macrofauna communities depends on the life cycles and natural histories of the constituent species and on the complexity of the preeutrophic community. Recovery of `simpler' assemblages is completed faster than in the case of communities with higher organization (Diaz & Rosenberg, 1995). Complete recovery of a system that has suffered severe anoxia takes 5±8 years and the `mature' community that develops after recovery may not resemble the original community. A greater understanding of the processes underlying these faunal patterns is needed in order to describe accurately how functional changes could, for example, alter rates of nutrient regeneration and bioturbation in benthic systems during and also after eutrophication (Weston, 1990).

The fate of benthic nutrients Are benthic suspension-feeding populations likely to deplete suspended food in the overlying water column? Oxygen deficiency in temperate coastal waters has led to an increased awareness of vertical particle flux to bottom waters. In recent years, particular attention has been paid to coupling and energy transfer between the benthos and plankton. It has been postulated that suspension-feeding communities can self-organize in order to enhance the efficiency of food capture and thus establish boundary systems capable of successfully exploiting the less structured planktonic system (Gili & Coma, 1998). Other studies demonstrated that in temperate coastal waters the most prominent event in the annual flux of organic material to the benthos is usually the spring diatom bloom (Smetacek, 1984; Harding et al., 1987; Wassmann, 1991; Olesen & Lundsgaard, 1995). In coastal (and also temperate and polar) seas this rapidly sinking phytoplankton is often dominated by diatoms that reach the sea floor relatively intact without being ingested by zooplankton (see Smetacek, 1985 for review, Alldredge & Gostschalk, 1989). Seasonally sedimented phytoplankton blooms are a major source of nutrients that are processed rapidly through the benthic system in open coastal areas (Graf et al., 1982; Gili & Coma, 1998). Information from field studies supports ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

M A R I N E E U T R O P H I C A T I O N A N D B E N T H O S : A R E V I E W 819 the hypothesis that suspension feeders ingest a wide spectrum of particle sizes (Ribes et al., 1998; Coma et al., 1995; Pile et al., 1996). Many suspension feeders are capable of utilizing any type of food, and are limited only by morphological constraints (Riisgard & Manriquez, 1997). For the case of suspension-feeding bivalves, food quality and quantity have been shown by Willdish & Kristmanson (1997) and Rosenberg & Loo (1988) to be a limiting factor. For example, infaunal bivalve growth was shown to be positively correlated with both the chlorophyll input to the sediment and the diatom availability in the near-bottom waters (MacDonald & Thompson, 1985; Thompson & Nichols, 1988; see Wildish & Kristmanson, 1997 for review). It has been widely recognized that benthic suspension feeders, which are among the main contributors to the biomass of benthic communities of coastal and estuarine ecosystems worldwide, benefit directly from pelagic primary production in the overlying water column (Graf et al., 1982; Christensen & Kanneworff, 1986). Thus, suspension feeders are responsible for a considerable share of the energy flow from the pelagic to the benthic system, in addition to secondary production in benthic environments (Petersen & Black, 1987; Gili & Coma, 1998). There are strong indications that phytoplankton biomass may be severely reduced (Fig. 4) or regulated by active suspension feeders in shallow ecosystems (Cloern, 1982; Officer et al., 1982; Smaal et al., 1986; Hily, 1991; Chauvaud et al., 2000). From these studies, it can be concluded that the secondary trophic level is dominated by the benthic ecosystem where active suspension feeders can even regulate pelagic primary production

when the water body is shallow, the residence time is long, and the suspension-feeding biomass is high (Cloern, 1996). Figure 5 illustrates the fact that suspension feeders are able to consume potentially high amounts of suspended food where the clearance time is shorter than the residence time of the water in the bay. Rates of suspension-feeding are a function of food supply for a variety of taxa, but as the process of filter-feeding uses a variety of various particle-trapping mechanisms and ranges from passive to active filter feeding, the relation between suspension feeders abundance and distribution is highly complex. Hydrodynamic factors may be critical in determining the food supply. The availability of food depends on the three-dimensional nature of the fluid and particulate fluxes to the benthic ecosystem (FreÂchette et al., 1989). Herman & Scholten (1990) proposed three reasons why benthic suspension feeders may have a stabilizing influence on benthic ecosystems. First, benthic suspension feeders are a stable component of the ecosystem. Secondly, the filtration rate of suspension feeders does not stabilize as food availability increases. Finally, the biomass of suspension feeders has a slow turnover rate. Together these three observations suggest that suspension feeders may limit the intensity and duration of phytoplankton blooms. By contrast, the zooplankton biomass increases in response to phytoplankton blooms and thus does not reach an effective filtration level during the bloom period. Mussel or clam beds and oyster reefs may also supply nutrients in high amounts to the overlying water column and promote phytoplankton production. Thus suspension feeders are not only important in terms of direct

Fig. 4 An example of the control of pelagic primary production by an introduced suspension feeder Potamocorbula amurensis in San Francisco Bay. (After Carlton et al., 1990). ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

820 J A C Q U E S G R A L L & L A U R E N T C H A U V A U D 10000

No regulation

Clearance time (days)

1000

DB CBP

100

AS NB WW

10

RA

SY

ES MO

BB

1 NI

CBO Regulation

SSF

0.1 0.1

1

10

100

1000

10000

100000

Residence time (days) Fig. 5 Graphical comparison of water mass residence time and clearance time in suspension feeders-dominated ecosystems. Ecosystems situated in the shaded area are potentially regulated by suspension feeders where clearance time is shorter than residence time. AS: AskoÈ bay; SSF: South San Francisco Bay; ES: Oosterschelde; CBO: Chesapeake Bay, Past; BB: Bay of Brest; MO: Marennes-OleÂron Bay; RA: Ria de Arosa; WW: Western Wadden Zee; NI: North Inlet; NB: Narangasett Bay; SY: Sylt, Eastern Wadden Zee; DB: Delaware Bay; CBP: Chesapeake Bay, Present. (In Dame, 1996).

Nutrients

Advection Quality ? R. PP Quantity

PP

Grrazing + Biodeposits

Nutrients resuspension

Bioturbation Fig. 6 Schematic representation of the nutrient fluxes and pelagic primary production dynamics in a suspension feeder-dominated ecosystem. PP ˆ Primary production, RPP ˆ Regenerated primary production.

control, but also affect nutrient recycling and sedimentation or resuspension of organic particulate matter (Fig. 6). When the amount of food filtered by bivalves exceeds the needs of the individual, pseudofaeces are produced that incorporate the excess particulate organic matter. The formation of these pseudofaeces facilitates the sedimentation of POM and thus increases the overall sedimentation rate. Biodeposits are also a source of food for benthic organismsÐsuch as bacteria, meiofauna and

macrofauna (Graf, 1992). Faeces and pseudofaeces contribute to enhance bacterial activity on a day-scale basis, whereas meiofauna and macrofauna populations rather respond on a week- and month-scale basis, respectively (Grenz et al., 1990; Smaal & Prinz, 1993). Fauna may feed directly either on the organic matter of the biodeposits or on bacteria, which contribute to secondary production in the benthos but also to increase the turnover of nutrients (Graf, 1992, see below). Kautsky & Evans (1987) argued ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

M A R I N E E U T R O P H I C A T I O N A N D B E N T H O S : A R E V I E W 821 that the role of suspension feeders in energy transfer may be minimal compared with the role they have in carbon and nutrient cycling in coastal ecosystems. In areas controlled and stabilized by suspension feeders and subject to increasing nutrient loads, the ecosystem is then highly vulnerable to changes in the suspension-feeder community (see also, Chauvaud et al., 2000).

Nutrient absorption/regeneration and zoobenthic communities The importance of the carbon (C) flux in the sediment is strongly determined by the origin as well as by the potential degradability of the organic matter arriving on the sea floor. Organic matter entering coastal marine sediments originates outside the basin (allochthonous) as phytoplankton bloom sedimentation or from decaying macrophyte (either macroalgae or vascular plants) detritus transported in by water or wind currents. Low specific gravity particles generally have a higher organic content than high specific gravity particles. Organic matter produced by primary or heterotrophic production (such as benthic grazers) inside the basin is termed autochthonous. The mineralization of organic matter begins soon after it reaches the bottom sediments; benthic microbial organisms and meiofauna respond quickly (on the order of days), whereas macrofauna respond more slowly the sediment column (Somerfield et al., 1995). The degradation rate of the organic material strongly depends on its quality; high quality, living phytoplankton are easily and quickly mineralized by microorganisms, meiofauna (copepods) or macrofaunal suspension and surface deposit feeders, whereas detritus from macroalgae or phanerogams is more refractory and thus more likely to be incorporated into macrofaunal deposit feeders (Widbom & Elmgren, 1988). The structure of zoobenthic assemblages may strongly be influenced by the quality of the organic matter input to the sediment (Heip, 1995). When high-quality organic matter dominates (through the sedimentation of phytoplankton or fresh organic matter), most organisms are found at shallow sediments depth (, 2 cm) and are mainly surface-deposit feeders or suspension feeders. On the other hand, when poorquality organic matter dominates, assemblages are dominated by deep burrowing (up to 20 cm) depositfeeding fauna (Dauwe et al., 1998). The processes of organic matter mineralization are also a function of the temperature, geochemistry, and the extent and is related to the intensity of bioturbation in the benthic ecosystem. A temperature increase of 10  C may augment marine coastal heterotrophic processes by a factor of 2±3 times; in addition, there is an inverse correlation between dissolved O concentrations and ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

temperature (Graf, 1992). The mineral assemblage and particle grain size of the sediments affect the equilibrium concentrations of nutrients in porewaters (Graf, 1992). For example, the sorption of P is a function of the surface area and mineralogical assemblage of the sediments; in general sandy sediments have a lower capacity for sorption than silt- and clay-rich sediments (De Casabianca et al., 1997). Water exchange is enhanced by feeding and burrowing behaviours (Fig. 6), affecting the efficiency of both heterotrophic activity and solute fluxes. Several studies have shown experimentally that degradation of organic matter was faster and more efficient in sediments with macrofauna populations. It has been suggested by various authors that the proportion of nutrient export from the benthic ecosystem to the water column because of macrofaunal activity supplies a significant proportion of the phytoplankton requirements (Asmus & Asmus, 1991; Smaal & Prinz, 1993). However, there is no clear agreement on the amount; published estimates range from 0 to 100%, with a mean of 28±30% (Doering, 1989). A recent study by Grenz et al. (2000) showed that during a bloom, nutrient fluxes from the sediment represented 20% silicon (Si), 16% (P) and 9% (N) of the primary production demand. In situ rates of nutrient flux from sediments have been shown to be much higher below bivalve±aquaculture tables (De Casabianca et al., 1997). Experimental studies in mesocosms with an intact benthos have showed that measured annual apparent primary production increased by 23% relative to mesocosms lacking benthos. Furthermore, in another experimental study, Christensen et al. (2000) showed that nutrient release to the water column depends on the species function (both behaviour and biology) in the system; NH 4 and NO 3 release were 3 times higher in sediments which had the polychaeteÐ Nereis diversicolorÐas suspension feeder, and were just 1.5 times higher in sediments which included Nereis virens as a deposit feeder. The presence of Nereis in the sediment increased both NO2 and silicate fluxes by two orders of magnitude. Magni et al. (2000) showed that macrofauna rapidly and efficiently recycle the inorganic forms of N and P, thus playing a role in the process of nutrient regeneration. Using an open flow-through system, Asmus & Asmus (1991) have shown that the potential for primary production induced by the nutrient release of an intertidal mussel bed was higher than the uptake of phytoplankton by the mussel bed itself. Thus, if benthic communities are able to reduce significantly the phytoplankton biomass, then they also have the potential to maintain eutrophication (De Casabianca et al., 1997). Oviatt et al. (1986) note that as primary production increases, respiration and thus recycling of nutrients in the water column also increases. The pelagic production

822 J A C Q U E S G R A L L & L A U R E N T C H A U V A U D fuelled by nutrients derived from the benthos could then result in higher pelagic recycling rates. De Casabianca et al. (1997) argued that the concentration of dissolved inorganic N did not vary much as a function of the presence or absence of a benthos, but rather in recycling of nutrients either in bottom sediments or in the water column. Thus, the effects of the benthos on pelagic production are two-fold: supplying nutrients directly, and indirectly increasing regeneration rates in the overlying water column.

Relationship between benthic macrophytes and nutrient cycling Dense assemblages of macrophytes covering sediments are likely to intercept nutrients from the water column or from the sediments, and in some cases this uptake is sufficient enough to maintain reduced nutrient concentrations in the overlying waters (Fig. 6). Porewater concentrations of inorganic N and P may be 10±30 times higher than in the water column and, if an adequate flux rate is present, the porewaters can provide a portion of the dissolved nutrients required by submerged benthic macrophytes. Lee & Dunton (2000) concluded, based on seagrass growth response to an increase in porewater NH+4 observations, that sediment N availability limits seagrass productivity. Macrophytes would then have a preference for absorbing their nutrients from the porewaters exclusively, or use both sources (porewater plus the water column) of nutrients (Flindt et al., 1999). Moreover, Valiela et al. (1997) argued that the water quality in macrophyte-dominated shallow-water systems may be much better than in a phytoplankton-dominated system assuming similar nutrient loadings. However, some species create canopies of foliage that result in a net storage of C and N that would otherwise remain in the water column. Several authors have shown that macroalgal biomass may, when important in an estuarine system, act as a major sink of nutrients during growth and as a source during decay (Risgaard-Petersen et al., 2000). In some ecosystems, benthic macrophytes may even store N on the same order of magnitude as the annual N load coming from freshwater inputs, and thus must be accounted for in nutrient and C-cycling budgets. Valiela et al. (1997) have speculated that C flux through trophic webs, and thus nutrient regeneration, would be faster in sediments containing macroalgae detritus than in sediments containing phanerogam detritus. Macrophyte populations which grow in eutrophic waters show higher biomass peaks and variations, which suggest the macrophytes may also sustain the eutrophic conditions in which they are living (Fig. 6) (De Casabianca et al., 1997). Once the eutrophication process is engaged, it may be indirectly favoured by the benthic

organisms and communities, and thus the role of the benthos is important to evaluate in ecosystem-level eutrophication studies.

Limits of our observations Cloern (2001) in describing the `evolving conceptual model of the coastal eutrophication problem' highlighted a need to analyse ecosystem changes as responses to multiple stressors. Coastal ecosystems are highly complex (Nixon, 1992) and Cloern (2001) demonstrated that we have to consider the myriad potential interactions among those stressors, some of which are discussed in more detail below.

Selective mortality following HAB The harmful algal bloom (HAB) is an indirect result of human activities (Fig. 7). There is little dispute that the number of HABs, the economic losses from them, the types of resources affected and the number of toxins and toxic species have all increased in recent years at sites worldwide (Smayda, 1990; Hallegraeff, 1993; Anderson, 1995). However, the reasons for this expansion have not been well-documented and many various explanations have been proposedÐsuch as natural species dispersal, increase in aquaculture operations, destruction of fisheries and discharge of ship ballast water. Another explanation for the increased incidence of HAB outbreaks is that increased pollution and nutrient loading in coastal waters have resulted in coastal ecosystem conditions favourable to HABs (Radach et al., 1990; Smayda, 1990). Massive mortality of the majority of benthic invertebrates is often described following an HAB (of the toxic dinoflagellate, Gyrodinium aureolum) and selective mortality of littoral fauna was also reported for adult populations in Ireland (Southgate et al., 1984) in Norway (Olsgard, 1993) and for postlarval populations in Brittary (Chauvaud 1998). Chauvaud & Thouzeau (2001) have suggested that toxic nonsiliceous phytoplankton can alter recruitment of benthic postlarvae. In the Bay of Brest, specific changes in the composition of benthic communities (Fig. 7) could result from the occurrence of toxic phytoplankton species which may cause differential larval and/or postlarval mortality, and not from organic matter enrichment (Pearson & Rosenberg, 1978; Gray, 1982; Gray et al., 1990; Grall & GleÂmarec, 1997) or mass adult mortality. A trophic routing and impact model was proposed by Smayda (1992) who showed possible linkages between HAB and various marine food-web compartments. However, the nature of the interactions between the quality and quantity of phytoplankton and long-term changes in benthos at all life stages (larvae, postlarvae, juveniles) have not yet been investigated and quantified. ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

M A R I N E E U T R O P H I C A T I O N A N D B E N T H O S : A R E V I E W 823

Toxic phytoplankton events

Larvae Mortality Adults

Benthic populations

Postlarvae

Mortality Juveniles

Modifications of recruitment

Sensitive species: Pecten maximus, Aequipecten opercularis, Chlamys varia... Indifferent species: Crepidula fornicata, Rissoa parva... Opprtunistic species: Jassa pusilla, Musculus spp.... Fig. 7 Specific changes in the composition of benthic communities resulting from the occurrence of toxic phytoplankton blooms (After Chauvaud, 1998).

Pollutants The effects of chemical pollutants (toxic metals, pesticides, herbicides, oil spills) may be to selectively inhibit some classes or species of algae, and indirectly promote population growth of less-sensitive taxa (Cloern, 2001). The effects of such pollutants on the ecosystem may be massive mortalities of several benthic populations (for example copper, Castilla & Nealler, 1978) or may be more selective (such as the sterilization of Nucella lapillus by TBT, Huet et al., 1996) ultimately causing the extinction of a single species or family (see Gibbs & Bryan, 1996). Macrofaunal communities, owing to the relative ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

longevity of their component species, may have a physiological process that allows the organism to sequester, and thus neutralize, some contaminants by incorporating the contaminant into a part of their organism (Somerfield et al., 1995). By contrast, meiofaunal communities do not integrate contaminants into their organism. Therefore, the effects of a contaminant may be observed for years after an eventÐsuch as an oil spillÐin benthic populations by surveying not only individuals for preserved contaminant signatures but also the species assemblage present in the benthic community (Poggiale & Dauvin, 2001). Gunnarson & SkoÈld (2000) suggested that direct consequences of eutrophicationÐsuch as increasing

824 J A C Q U E S G R A L L & L A U R E N T C H A U V A U D phytoplankton productionÐmay induce bioaccumulation of organic pollutants in benthic fauna at higher trophic levels.

Diseases Mass mortalities as a result of disease outbreaks have recently affected marine benthic populations (kelp, seagrasses, coral, sponges, starfish, abalone, oysters or scallops; see Harvell et al., 1999 for review). The fact that these species may be endangered should encourage further investigations on the potential interactions between marine diseases and nutrient enrichment. The case of the Chesapeake Bay (11 500 km2) illustrates the consequences of what occurs when a commercially important shellfish population was decimated by disease. The bay was the most important commercial source of the American oysterÐCrassostrea virginica; however, since 1970 a virus (Perkinsus marinus) which proliferates when water quality decreases, has nearly pushed the oyster stock to extinction in the bay (Newell, 1988; Ulanowicz & Tuttle, 1992). In the summer season, the population of oysters in Chesapeake Bay prior to 1870 had the potential to filter the entire bay's water in one week, whereas the currently depleted population requires more than 46 weeks. The decline in filtering capacity decreases the grazing pressure on the phytoplankton population and could increase the flux of sedimenting organic matter, thus providing the fuel for creating anoxic conditions in the bottom waters of the bay (Kemp & Boyton, 1984). The introduction of a disease into an ecosystem in addition to the existence of other stressors (overfishing, prior or concurrent presence of one or more toxic contaminants, habitat destruction, etc.) can destroy populations. For example, decreased resistance of oysters to P. marinus after sublethal exposure to tributyltin oxide (Fisher et al., 1999) or other pollutants (Chu & Hale, 1994) has also been reported. The intensity of the viral oyster disease (P. marinus) is also influenced by El Nino/La Nina climatic shifts (Kim & Powell, 1998) and Kim et al. (1999) have demonstrated the influence of climate on the temporal variability in body burden of most organic, and some metal, contaminants in oyster populations in the Gulf of Mexico.

Transport of organisms Invasive species often outcompete native fauna for food and space, in the process altering the interactions between multiple components in the affected ecosystem (Chauvaud, 1998). The most famous example is the invasion of San Francisco Bay by the Asian clamÐ Potamocorbula amurensisÐin 1986 (Carlton et al., 1990).

Within 2 years, the clam had proliferated to more than 10 000 individuals m 2, and accounted for up to 95% of the benthic biomass in colonized areas (Nichols et al., 1990). In the northern basin of the bay, this dense population directly controls the phytoplankton biomass (Cloern, 1982, 1996; Lucas et al., 1999) and affects the zooplanckton community by predation or by food competition (Kimmerer et al., 1994). The appearance and proliferation of this species have altered the food web of the entire ecosystem (Nichols et al., 1990; Thompson, 1999). Field studies on another introduced speciesÐ Crepidula fornicata, an introduced species in the Bay of Brest (France)Ðdemonstrated that the population density is high enough to directly control the export rate of biogenic silica to the open ocean, and thus also controls the specific composition of phytoplankton blooms in the bay (Chauvaud et al., 2000). These examples show how spreading of exotic species in the benthic ecosystem and eutrophication can be interconnected and how these species can alter the ecosystem functioning when the population is high enough to control primary production in the overlying water column (Cloern, 1982; Chauvaud et al., 2000), and alter the cycling of nutrientsÐsuch as C (Dame & Patten, 1981; Doering et al., 1987), O (Effler & Siegfried, 1994), N (Dame, 1993), P (Asmus et al., 1995) or silica (Chauvaud et al., 1998). Benthic community compositions have been profoundly modified by the proliferating invading species, threatening the benthic biodiversity at the ecosystem scale (Chauvaud et al., 2000). Stachowicz et al. (1999) demonstrated experimentally that increased species richness significantly decreases invasion success, thus the impact of mild eutrophication on biodiversity should be included in further studies.

Climate changes Climate is believed to be responsible for many aspects of temporal variability of marine communities (Ware, 1995), including cyclical patterns of variation in some populations (Gray & Christie, 1983). In such a context, working on soft-bottom macrobenthic communities on the Swedish west coast, Tunberg & Nelson (1998) suggested that climatic variability may be a more basic causative factor for benthic disturbance than eutrophication and other possible stressors that have been proposed previously. Even if the climate±eutrophication linkage is well-established for the pelagic system (see Cloern, 2001), it remains poorly defined for benthic±pelagic coupling and the benthos. Indeed, if the role of climatic variation is not considered, it may lead to incorrect conclusions regarding the relative contribution of anthropogenic vs. naturally generated perturbations in regulating marine biological systems (Tunberg & Nelson, 1998). ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

M A R I N E E U T R O P H I C A T I O N A N D B E N T H O S : A R E V I E W 825

Physical disturbance of the sediments Hydrodynamic processes responsible for sediment movements and physical disturbances caused by human activities (e.g. bottom fishing, dredging and gravel-extraction operations) can deeply modify benthic communities (Hall, 1994). The importance of tidal currents, waves and storms on benthic communities patterns and dynamics have been pointed out by Emerson & Grant (1991). These authors have argued that these natural perturbations can control population demography as well as the age structure of populations in temperate areas. Bottom fishing is one of the most widespread sources of anthropogenic disturbances of seabed communities (Kaiser et al., 1998). Ways in which bottom fishing directly affects benthos can be classified as ploughing, sediment resuspension, disruption to surface physical features and removal of both target and nontarget species. Other effects include mortality, the attraction of scavenger populations (Collie et al., 1997), and long-term changes to the benthic community structure, with possible reductions in benthic biodiversity (Kaiser et al., 1998; Hall-Spencer & Moore, 2000). Harbour and channel dredging is usually an infrequent event and sediments are then recolonized by benthic organisms, with successional stages as described by Van der Veer et al. (1985), although communities may never return to their predredging structure. Continuous gravel extraction on one site usually leads to complete defaunation, whereas surrounding areas may suffer from siltation. Dredging and bottom fishing occur along some of the most densely populated coastlines of the world (North Sea, Wadden Zee. . .); places that have been shown to be under the threat of eutrophication as well, emphasizing the difficulty of describing and quantifying benthos dynamics and building useful, predictive models of the function of coastal benthic ecosystems.

Concluding remarks Benthic communities are highly sensitive to the flux and refractory quality of organic matter; the community metabolism, composition and structure are all affected. An increase in the flux of high-quality organic matter to the sediments can result in increased biomass and shifts in the assemblages towards the dominance of r-selected species which have higher productivity per biomass ratios. However, when organic loading exceeds certain site-specific thresholds, hypoxic conditions in the overlying waters of the sediment may appear that ultimately lead to the total disappearance of the macrofauna. Hecky & Kilham (1988) speculated on the possibilities of managing species composition of phytoplankton communities to increase the secondary productivity of ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

valuable food species, including benthos as target species or food for predatory fish. But an agricultural model cannot be applied to coastal marine ecosystems: 'No farmer feeds his animals by broadcasting their fodder on the wind' (Nixon, 1986). There are strong interactions between nutrient loading, environmental forces and the benthos. Benthic communities are highly variable in response to environmental factors and may show natural variability that exceeds the responses to anthropogenic stressors. Yet, long-term studies have shown evidence of profound modifications in benthic structures during the last century (Nixon, 1986). In this review, we have tried to highlight the importance of considering the effects of eutrophication on benthos together with other anthropogenic disturbances which may simultaneously influence ecosystems at comparable scales. It is now clear that benthos influence a significant part of the supply of nutrients for primary production. The discrimination between pelagos on the one hand and benthos on the other hand is now obsolete. We have to consider the aquatic ecosystem as a single entity. The challenge is to objectively quantify and describe the detailed processes and ramifications of benthic eutrophication at the ecosystem scale, taking into account all the potential stressors. We need to consider many more factors, standardize methods of ecological survey and compare eutrophic ecosystems on a global scale. Ultimately, building models of these ecosystems that can be used for not only estimating the sensitivity of the ecosystems to various factors but also assessing the risks to them is crucial.

Aknowledgements The authors would like to thank Dr Jason Hall-Spencer (UMBS Millport), Prof Geoff Moore (UMBS Millport) and Dr Jennifer Guarini for English correction. Dr Jennifer Guarini, Dr Jean Marc Guarini, Prof Michel Glemarec and Dr Jason HallSpencer, together with two anonymous referees, gave valuable comments for improvement of the manuscript. This paper was presented as plenary address of the `Response of benthic organisms to eutrophication' session of the 31st Annual Symposium of the Estuarine and Coastal Science Association, Bilbao, Spain, July 2000. The work was partly funded by the ACI-PECTEN French government program.

References Alldredge AL, Gostschalk CC (1989) Direct observations of the mass flocculation of diatom blooms: characteristics, settling velocities and formation of diatom aggregates. Deep-Sea Research, 36, 159±171. Anderson DM (1995) Toxic red tides and harmful algal blooms: a practical challenge in coastal oceanography. Reviews of Geophysics, 1, 1189±1200.

826 J A C Q U E S G R A L L & L A U R E N T C H A U V A U D Asmus RM, Asmus H (1991) Mussel beds: limiting or promoting phytoplankton? Journal of Experimental Marine Biology and Ecology, 148, 215±232. Asmus H, Asmus RM, Zubillaga GF (1995) Do mussel beds intensify the phosphorus exchange between sediment and tidal waters? Ophelia, 41, 37±55. Baden SP, Loo L-O, Pihl L et al. (1990) Effects of eutrophication on benthic communities including fish: Swedish west coast. Ambio, 19 (3), 113±122. Beukema JJ, CadeÂe GC (1997) Local differences in macrozoobenthic response to enhanced food supply caused by mild eutrophication in a Wadden Zee area: Food is only locally a limiting factor. Limnology and Oceanography, 42(6), 1424±1435. Bonsdorff E (1992) Drifting algae and zoobenthos ± effects on settling and community structure. The Netherlands Journal of Sea Research, 30, 57±62. Bonsdorff E, Pearson TH (1999) Variation in the sublittoral macrozoobenthos of the Baltic seas: a functional group approach. Australian Journal of Ecology, 24, 312±326. Burkholder JM, Glasgow HB, Cooke JE (1994) Comparative effects of water column enrichment on eelgrass Zostera marina, shoalgrass Halodule wrightii and widgeongrass Ruppia maritima. Marine Ecology Progress Series, 105, 121±138. Carlton JT, Thompson JK, Schemel LE et al. (1990) Remarkable invasion of San Francisco Bay (California, USA) by the Asian clam Potamocorbula amurensis. I. Introduction and dispersal. Marine Ecology Progress Series, 66, 81±94. Castilla JC, Nealler E (1978) Marine environmental impact due to mining activities of El Salvador copper mine, Chile. Marine Pollution Bulletin, 9, 67±70. Cederwall H, Elmgrenn R (1980) Biomass increase of benthic macrofauna demonstrates eutrophication of the Baltic sea. Ophelia supplement, 1, 287±296. Chauvaud L (1998) La coquille Saint Jacques en rade de Brest: un modeÁle biologique d'eÂtude des reÂponses de la faune benthique aux fluctuations de l'environnement. TheÁse de doctorat, universite de Bretagne Occidentale, Brest. Chauvaud L, Jean F, Ragueneau O et al. (2000) Long-term variation of the Bay of Brest ecosystem: benthic-pelagic coupling revisited. Marine Ecology Progress Series, 200, 35±48. Chauvaud L, Thouzeau G, Paulet YM (1998) Effects of environmental factors on the daily growth rate of Pecten maximus juveniles in the bay of Brest (France). Journal of Experimental Marine Biology and Ecology, 227, 83±111. Christensen H, Kanneworff E (1986) Sedimentation of phytoplankton during a spring bloom in the Oresund. Ophelia, 26, 109±122. Christensen B, Vedel A, Kristensen E (2000) Carbon and nitrogen fluxes in sediment inhabited by suspension-feeding (Nereis diversicolor) and non-suspension-feeding (N. virens) polychaetes. Marine Ecology Progress Series, 192, 203±217. Chu FL, Hale RC (1994) Relationship between pollution and susceptibility to infectious disease in the eastern oyster, Crassostrea virginica. Marine Environmental Research, 38, 243±256. Cloern JE (1982) Does the benthos control phytoplankton biomass in south San Francisco Bay? Marine Ecology Progress Series, 9, 191±202.

Cloern JE (1996) Phytoplankton blooms dynamics in coastal ecosystems; a review with some general lessons from sustained investigation of San Francisco Bay, California. Review of Geophysics, 34, 127±168. Cloern JE (2001) Our evolving conceptual model of the coastal eutrophication problem. Marine Ecology Progress Series, 210, 235±265. Collie JS, Escanero GA, Valentine PG (1997) Effects of bottom fishing on the benthic megafauna of Georges Bank. Marine Ecology Progress Series, 155, 159±172. Collos Y (1987) Simultaneous uptake of nitrate, nitrite, ammonium and urea by microphytobenthos in the Lauzieres Channel during summer. In 2nd Soviet±French symposium on production and trophic relationships, YALTA 27 OCTOBER 2 NOVEMBER 1984. Actes de Colloques IFREMER, 5, 215±220. Coma R, Gili JM, Zabala M (1995) Trophic ecology of a benthic marine hydroid Campanularia everta. Marine Ecology Progress Series, 119, 211±220. Dame RF (1993) The role of bivalve filter feeder material fluxes in estuarine ecosystems. In: Bivalve Filter Feeders in Estuarine and Coastal Ecosystem Processes (ed. Dame RF). Springer-Verlag, Berlin, 245±269. Dame RF (1996) Ecology of marine bivalves: an ecosystem approach. CRC Marine Science Series, New York, 254 pp. Dame RF, Patten BC (1981) Analysis of energy flows in an intertidal oyster reef. Journal of Experimental Marine Biology and Ecology, 5, 115±124. Dauer DM, Conner WG (1980) Effects of moderate sewage input on benthic polychaete populations. Estuarine, Coastal and Shelf Science, 10, 335±346. Dauwe B, Herman PMJ, Heip CHR (1998) Community structure and bioturbation potential of macrofauna at four North Sea stations with contrasting food supply. Marine Ecology Progress Series, 173, 67±83. De Casabianca ML, Laudier T, Marinho-Sauriano E (1997) Seasonal changes of nutrients in water and sediment in a Mediterranean lagoon with shellfish farming activity (Thau lagoon, France). ICES Journal of Marine Science, 54, 905±916. Diaz RJ, Rosenberg R (1995) Marine benthic hypoxia: a review of its ecological effects and the behavioural responses of benthic macrofauna. Oceanography and Marine Biology: An Annual Review, 33, 245±303. Doering PH (1989) On the contribution of the benthos to pelagic production. Journal of Marine Research, 47, 371±383. Doering PH, Kelly JR, Oviatt CA et al. (1987) Effects of the hard clam Mercenaria mercenaria on benthic fluxes of inorganic nutrients and gases. Marine Biology, 94, 377±383. Effler SW, Siegfried C (1994) Zebra mussel (Dreissena polymorpha) Populations in the Seneca River, New York: Impact on oxygen resources. Environmental Science and Technology, 28, 2216±2221. Emerson CW, Grant J (1991) The control of soft-shell clam (Mya arenaria) recruitment on intertidal sandflats by bedload sediment transport. Limnology and Oceanography, 36, 1288±1300. Everett RA (1991) Intertidal distribution of infauna in central California lagoon: the role of seasonal blooms of macroalgae. Journal of Experimental Marine Biology and Ecology, 150, 223±247. Ferraro SP, Swartz RC, Cole FA et al. (1991) Temporal changes in the benthos along a pollution gradient: discriminating the ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

M A R I N E E U T R O P H I C A T I O N A N D B E N T H O S : A R E V I E W 827 effects of natural phenomena sewage-industrial wastewater effects. Estuarine, Coastal and Shelf Science, 33, 383±407. Fisher TR, Gustafson AB, Sellner K et al. (1999) Spatial and temporal variation of resource limitation in Chespeake Bay. Marine Biology, 133, 763±778. Flindt MR, Pardal JA, Lillebo AI et al. (1999) Nutrient cycling and plant dynamics in estuaries: a brief review. Acta Oecologica, 20 (4), 237±248. FreÂchette M, Butman CA, Geyer WR (1989) The importance of boundary-layerflows in supplying phytoplankton to the benthic suspension-feeder, Mytilus edulis L. Limnology and Oceanography, 34, 19±36. Gee JM, Warwick RM (1994) Metazoan community structure on relation to the fractal dimensions of marine algae. Marine Ecology Progress Series, 103, 141±150. Gibbs PE, Bryan GW (1996) TBT-induced imposex in neogastropod snails: masculinization to mass extinction. In: Case Study of an Environmental Contaminant: Tributyltin (ed. De Mora SJ), pp. 212±236, Cambridge University Press, Cambridge. Gili JM, Coma R (1998) Benthic suspension feeders: their paramount role in littoral marine food webs. Trends in Ecology and Evolution, 13 (8), 316±321. Graf G (1992) Benthic-pelagic coupling: a benthic view. Oceanography and Marine Biology Annual Review, 30, 149±190. Graf G, Bengtsson W, Diesner U et al. (1982) Benthic response to sedimentation of a spring phytoplankton bloom: process and budget. Marine Biology, 67, 201±208. Grall J, GleÂmarec M (1997) Using biotic indices to estimate macrobenthic community perturbations in the bay of Brest. Estuarine, Coastal and shelf Science, 44 (Supplement A), 43±53. Graneli E, Sundbaeck K (1985) The response of planktonic and microbenthic algal assemblages to nutrient enrichment in shallow coastal waters, southwest Sweden. Journal of Experimental Marine Biology and Ecology, 85, 253±268. Grassle JF, Grassle JP, Brown-Leger LS et al. (1985) Subtidal macrobenthos of Narragansett Bay. Field and mesocosm studies of the effects of eutrophication and organic input on benthic populations. Marine Biology of Polar Regions and Effects of Stress on Marine Organisms (eds Gray JS, Christiansen ME), pp. 421±434. Wiley, New York. Gray JS (1982) Effects of pollutants on marine ecosystems. Netherland Journal of Sea Research, 16, 424±443. Gray JS (1992) Eutrophication in the sea. In: Proceedings of the 25th European Marine Biology Symposium (eds Colombo G, Ferrari I, Ceccherelli VU, Rossi R), pp. 3±15. Olsen & Olsen, Fredendsbog. Gray JS, Christie H (1983) Predicting long-term changes in marine benthic communities. Marine Ecology Progress Series, 13, 87±94. Gray JS, Clarke KR, Warwick RM et al. (1990) Detection of initial effects of pollution on marine benthos: an example from the Ekofisk and Edfisk oilfields, North Sea. Marine Ecology Progress Series, 66, 285±299. Grenz C, Cloern JE, Hager SW et al. (2000) Dynamics of nutrient cycling and related benthic nutrient and oxygen fluxes during a spring phytoplankton bloom in South San Francisco Bay (USA). Marine Ecology Progress Series, 197, 67±80.

ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

Grenz C, Hermin MN, Baudinet D et al. (1990) In situ biochemical and bacterial variation of sediments enriched with mussel biodeposits. Hydrobiologia, 207, 153±160. Gunnarson JS, SkoÈld M (2000) Accumulation of polychlorinated biphenyls by the infaunal brittle stars Amphiura filiformis and A. chiajei: effects of eutrophication and selective feeding. Marine Ecology Progress Series, 186, 173±185. Hall SJ (1994) Physical disturbance and marine benthic communities: life in unconsolidated sediments. Oceanography and Marine Biology: an Annual Review, 32, 179±239. Hallegraeff GM (1993) A review of harmful algal blooms and their apparent global increase. Phycologia, 32, 79±99. Hall-Spencer JM, Moore PG (2000) Scallop dredging has profound, long-term impacts on maerl habitats. ICES Journal of Marine Science, 57, 1407±1415. Harding GC, Hargrave BT, Vass WP et al. (1987) Vertical flux of particulate matter by sedimentation and zooplankton movements in St George's Bay, the southern Gulf of St. Lawrence. Biological Oceanography, 4, 323±357. Harvell CD, Kim K, Burkholder JM et al. (1999) Emerging marine diseases ± climate links and anthropogenic factors. Science 285, 1505±1510. Heck KL, Pennock JR, Valentine JF et al. (2000) Effects of nutrient enrichment and small predator density on seagrass ecosystems: an experimental assessment. Limnology and Oceanography, 45 (5), 1041±1057. Hecky RE, Kilham P (1988) Nutrient limitation of phytoplankton in freshwater and marine environments: a review of recent evidences on the effects of enrichment. Limnology and Oceanography, 33 (4), 796±822. Heip C (1995) Eutrophication and zoobenthos dynamic. Ophelia, 41, 113±136. Heip CHR, Goosen NK, Herman PMJ et al. (1995) Production and consumption of biological particles in temperate tidal estuaries. Oceanography and Marine Biology Annual Reviews, 33, 1±149. Herman PMJ, Scholten H (1990) Can suspension feeders stabilise estuarine ecosystems? In: Proceedings of the of the 24th European Marine Biology Symposium (eds Barnes M, Gibson RN), pp. 104±116. Aberdeen University Press, Aberdeen. Hily C (1983) Macrozoobenthic recolonisationa after dredging in a sandy mud area of the Bay of Brest enriched by organic matter. Oceanologica Acta, 15, 113±120. Hily C (1987) Spatio-temporal variability of Chaetozone setosa (Malmgren) population on an organic gradient in the bay of Brest, France. Journal of Experimental Marine Biology and Ecology, 112, 201±216. 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. Holmquist JG (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. Huet M, Paulet YM, GleÂmarec M (1996) Trybutylitin (TBT) pollution in the coastal waters of west Brittany as indicated by imposex in Nucella lapilus. Marine Environmental Research, 41 (2), 157±176.

828 J A C Q U E S G R A L L & L A U R E N T C H A U V A U D Hughes JE, Deegan LA, Peterson BJ et al. (2000) Nitrogen flow through the food web in the oligohaline zone of a New England estuary. Ecology, 81, 433±452. Hull SC (1987) Macroalgal mats and species abundance: a field experiment. Estuarine, Coastal and Shelf Science, 25, 519±532. Irigoien X, Castel J (1997) Light limitation and distribution of chlorophyll pigments in a highly turbid estuary: The Gironde (SW France). Estuarine, Coastal and Shelf Science, 44, 507±517. de Jonge VN (1990) Response of the Dutch Wadden Sea ecosystem to phosphorus discharges from the River Rhine. In: 18th Symposium of the Estuarine and Brackish-Water Sciences Assoc: North SeaÐEstuaries Interacions (eds Mclusky DS, de Jonge V N, Pomfret J), Hydrobiologia, Suppl. PP. 49±62. Jorgensen BB, Richardson K (1996) Eutrophication in Coastal Marine Ecosystems Coastal and Estuarine Studies 52. American Geophysical Union, Washington, DC, 267 PP. Kaiser MJ, Edwards DB, Armstrong PJ et al. (1998) Changes in megafaunal benthic communities in different habitats after trawling disturbance. ICES Journal of Marine Science, 55, 353±361. Kautsky H (1991) Influence of eutrophication on the distribution of phytobenthic plant and animal communities. International Revue ges Hydrobiology, 76 (3), 423±432. 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. Kemp WM, Boyton WR (1984) Influence of biological and physical processes on dissolved oxygen dynamics in an estuarine system: implications for measurement of community metabolism. Estuarine Coastal Marine Science, 11, 407±431. Kim Y, Powell E (1998) Influence of climate change on interannual variation in population attributes of Gulf of Mexico oyster. Journal of Shellfish Research, 8, 71±82. Kim Y, Powell EN, Wade TL et al. (1999) Inflence of climate change on interannual variation in contaminant body burden in Gulf of Mexico oysters. Marine Environmental Research, 48, 459±488. Kimmerer WJ, Gartside E, Orsi JJ (1994) Predation by an introduced clam as the probable cause of substantial declines in zooplankton in San Francisco Bay. Marine Ecology Progress Series, 113, 81±93. Kromkamp J, Peene J, Van Rijswijk P et al. (1995) Nutrients, light and primary production by phytoplankton and microphytobenthos in the eutrophic, turbid Westerschelde Estuary (The Netherlands). Hydrobiologia, 311, 9±19. Lapointe BE (1999) Simultaneaous top-down and bottom up forces control macroalgal blooms on coral reefs. Limnology and Oceanography, 44 (6), 1586±1592. Lee KS, Dunton KH (2000) Effects of nitrogen enrichment on biomass allocation, growth and leaf morphology of the seagrass Thalassia testudinum. Marine Ecology Progress Series, 196, 39±48. Lillebo AI, Pardal MA, Marques JC (1999) Population structure, dynamics and production of Hydrobia ulvae (Pennant) (Mollusca: Prosobranchia) along an eutrophication gradient in the Mondego estuary (Portugal). Acta Oecologica, 20, 289±304. Lucas LV, Koseff JR, Cloern JE et al. (1999) Processes governing phytoplankton blooms in estuaries. I. The local productionloss balance. Marine Ecology Progress Series, 187, 1±15.

MacDonald BA, Thompson RJ (1985) Influence of temperature and food availability on the ecological energetics of the giant scallop Placopecten magellanicus. I. Growth rates of shell and somatic tissue. Marine Ecology Progress Series, 25, 279±294. Magni P, Montani S, Takada C et al. (2000) Temporal scaling and relevance of bivalve nutrient excretion on a tidal flat of the Seto Inland Sea, Japan. Marine Ecology Progress Series, 198, 139±155. Majeed S (1987) Organic matter and biotic indices on the beaches of North Brittany. Marine Pollution Bulletin, 18 (9), 490±495. McClelland JW, Valiela I (1998) Changes of food web structure under the influence of increased anthropogenic nitrogen inputs to estuaries. Marine Ecology Progress Series, 168, 259±271. Mucha AP, Costa MH (1999) Macrobenthic community structure in two portuguese estuaries: relationship with organic enrichment and nutrient gradients. Acta Oecologica, 20 (4), 363±376. Nestlerode JA, Diaz RJ (1998) Effects of periodic hypoxia on predation of a tethered polychaete, Glycera americana: implication for trophic dynamics. Marine Ecology Progress Series, 172, 185±195. Newell RIE (1988) Ecological changes in Chesapeake Bay: are they the result of overharvesting the American oyster, Crassostrea virginica? In: Understanding the Estuary: Advances in Chesapeake Bay Research (eds Lynch MP, Krome EC), pp. 536±546. Chesapeake Research Consortium, Salomon's, Maryland. Nichols FH, Thompson JK, Schernel LE (1990) Remarkable invasion of San Francisco Bay California (USA) by the Asian clam Potamocorbula amurensis II. Displacement of a former community. Marine Ecology Progress Series, 66, 95±101. Nixon SW (1986) Nutrients dynamics and the productivity of marine coastal waters. In: Marine Environment and Pollution (eds Halwagy R, Clayton D, Behbehani M), pp. 97±115. The Alden Press, Oxford. Nixon SW (1992) Quantifying the relationship between nitrogen input and the productivity of marine ecosystems. In: Proceedings of Advanced Marine Technology Conference (eds Takahashi M, Nakata K, Parson TR), Vol. 5, pp. 57±83. Tokyo, Japan. Nixon SW (1995) Coastal marine eutrophication: a definition, social causes, and future concerns. Ophelia, 41, 199±219. Norkko A, Bonsdorff E (1996) Rapid zoobenthic community responses to accumulations of drifting algae. Marine Ecology Progress Series, 131, 143±157. Norkko J, 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. Odum EP (1970) Fundamentals of Ecology, 3rd edn. Saunders, Philadelphia, 574pp. Officer CB, Smayda TJ, Mann R (1982) Benthic filter feeding: a natural eutrophication control. Marine Ecology Progress Series, 9, 203±210. Olesen M, Lundsgaard C (1995) Seasonal sedimentation of autochthonous material from the euphotic zone of coastal ecosystem. Estuarine, Coastal and Shelf Science, 41, 475±490. ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

M A R I N E E U T R O P H I C A T I O N A N D B E N T H O S : A R E V I E W 829 Olsgard F (1993) Do toxic algal blooms affect subtidal softbottom communities? Marine Ecology Progress Series, 102, 269±286. Oviatt CA, Keller A, Sampou P et al. (1986) Patterns of productivity during eutrophication: a mesocosm experiment. Marine Ecology Progress Series, 28, 69±80. Paerl HW, Pinckney JL, Fear JM et al. (1998) Ecosystem responses to internal and watershed organic matter loading: consequences for hypoxia in the eutrophying Neuse River Estuary, North Carolina, USA. Marine Ecology Progress Series, 166, 17±25. Pardal MA, Marques JC, Metelo I et al. (2000) Impact of eutrophication on the life cycle, population dynamics and production of Ampithoe valida (Amphipoda) along an estuarine spatial gradient (Mondego estuary, Portugal). Marine Ecology Progress Series, 196, 207±219. Paterson DM, Crawford RM (1986) The structure of benthic diatom assemblages: a preliminary account of the use and evaluation of low-temperature scanning electron microscopy. Journal of Experimental Marine Biology and Ecology, 95, 279±289. Pearson TH, Josefson AB, Rosenberg R (1985) Petersen's stations revisited. Is the Kattegat becoming eutrophic? Journal of Experimental Marine Biology and Ecology, 92, 157±206. Pearson TH, Rosenberg R (1978) Macrobenthic successions in relation to organic enrichment and pollution of the marine environment. Oceanography and Marine Biology: an Annual Review, 16, 229±311. Petersen CH, Black R (1987) Resource depletion by active suspension feeders on tidal flat: influence of local density and tidal elevation. Limnology and Oceanography, 32, 143±166. Peterson CH, Irlandi EA, Black R (1994) The crash in suspensionfeeding bivalve populations (Katelysia spp.) in Princess Royal Harbour: an unexpected consequence of eutrophication. Journal of Experimental Marine Biology and Ecology, 176, 39±52. Philippart CJM, CadeÂe GC (2000) Was total primary production in the western Wadden Sea stimulated by nitrogen loading? Helgolander Marine Research, 54, 55±62. Pile AJ, Patterson MR, Witman JD (1996) In situ grazing on plankton , 10 mm by the boreal sponge Mycale lingua. Marine Ecology Progress Series, 141, 95±102. Piriou JY, Menesguen A (1992) Environmental factors controlling the Ulva sp. blooms in Brittany (France). In: Proceedings of the 25th European Marine Biology Symposium (eds Colombo G, Ferrari I, Ceccherelli VU, Rossi R), pp. 111±116. Olsen & Olsen, Fredendsbog. Poggiale JC, Dauvin JC (2001) Long-term dynamics of three benthic Ampelisca (Crustacea-Amphipoda) populations from the Bay of Morlaix (western English Channel) related to their disappearance after the `Amoco Cadiz' oil spill. Marine Ecology Progress Series, 214, 201±209. Radach G, Berg J, Hagmeier E (1990) Long term changes of the annual cycle of meteorological hydrographic, nutrient and phytoplankton time seriesat Helgoland and at LV ELBE 1 in the German Bight. Continental Shelf Research, 10, 305±328. Raffaelli D, Raven J, Poole L (1998) Ecological impact of green macroalgal blooms. Oceanography and Marine Biology: an Annual Review, 36, 97±126. ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830

Reise K, Herre E, Sturm M (1989) Historical changes in the benthos of the Wadden Sea around the island of Sylt in the North Sea. HelgolaÈnder Meeresunters, 43, 417±433. Ribes M, Coma R, Gili JM (1998) Heterotrophic feeding by symbiotic gorgonian corals. Limnology and Oceanography, 43(6), 1170±1179. Riisgard HU, Manriquez P (1997) Filter feeding in fifteen marine ectoprocts (Bryozoa): particle capture and water pumping. Marine Ecology Progress Series, 154, 223±239. Risgaard-Petersen N, Ditlev L, Ottosen M (2000) Nitrogen cycling in two temperate Zostera marina beds: seasonal variation. Marine Ecology Progress Series, 198, 93±107. Rizzo WM (1990) Nutrient exchanges between the water column and a subtidal benthic microalgal community. Estuaries, 13, 219±226. Rosenberg R, Loo LO (1988) Marine Eutrophication induced oxygen deficiency: effects on soft bottom fauna, western Sweden. Ophelia, 29, 213±225. Rosenberg R, MoÈller P (1979) Salinity stratified benthic communities and long term monitoring along the west coast of Sweden. Journal of Experimental Marine Biology and Ecology, 23, 175±203. Ryther JH, Dunstan WM (1971) Nitogen, Phosphorus and eutrophication in the coastal marine environment. Science, 171, 1008±10013. Schramm W (1999) Factors influencind seaweeds response to eutrophication: some results from EU-project EUMAC. Journal of Applied Phycology, 11, 69±78. Smaal AC, Prinz TC (1993) The uptake of organic matter and the release of inorganic nutrients by bivalve suspension feeder beds. In: Bivalve Filter Feeders in Estuarine and Coastal Ecosystem Processes (ed. Dame RF), Springer-Verlag, Berlin, pp. 271±298. Smaal AC, Verhagen JHG, Coosen J et al. (1986) Interaction between seston quantity and quality and benthic suspension feeders in the oosterschelde, the Netherlands. Ophelia, 26, 385±399. Smayda TJ (1990) Novel and nuisance of phytoplankton blooms in the sea: evidence for a global epimedia. In: Toxic Marine Phytoplankton (ed. Graneli E), Elsevier Science Publishing Co, New York, pp. 29±40. Smayda TJ (1992) Global epidemic of noxious phytoplankton blooms and food chain consequences in large ecosystems. In: Food Chains, Yields, Models, and Management of Large Marine Ecosystems (eds Sherman K, Alexander LM, Gold BD), pp. 275±307. Westview press, Boulder CO. Smetacek VS (1984) The supply of foods to the benthos. In: Flows of Energy and Materials in Marine Ecosystems: Theory and Pratice (ed. Fasham MJ), pp. 517±548. Plenum Press, New York. Smetacek VS (1985) Role of sinking in diatom life-history cycles: ecological, evolutionary and geological significance. Marine Biology, 84, 239±251. Smith SV (1984) Phosphorus versus Nitrogen limitation in the marine environment. Limnology and Oceanography, 29 (6), 1149±1160. Somerfield PJ, Rees HL, Warwick RM (1995) Interrelationships in community structure between shallow-water marine meiofauna and macrofauna in relation to dredging disposal. Marine Ecology Progress Series, 127, 103±112.

830 J A C Q U E S G R A L L & L A U R E N T C H A U V A U D Southgate T, Wilson K, Cross TF et al. (1984) Recolonization of a rocky shore in S.W. Ireland following a toxic bloom of the dinoflagellate Gyrodinium aureolum. Journal of the Marine Biological Association of the United Kingdom, 64, 485±492. Stachowicz JJ, Whitlatch RB, Osman RW (1999) Species diversity and invasion resistance in a marine ecosystem. Science, 286, 1577±1579. Sundbaeck K, Joensson B (1988) Microphytobenthic productivity and biomass in sublittoral sediments of a stratified bay, southeastern Kattegat. Journal of Experimental Marine Biology and Ecology, 122, 63±81. Taylor D, Nixon S, Granger S et al. (1995) Nutrient limitation and the eutrophication of coastal lagoons. Marine Ecology Progress Series, 127, 235±244. Thompson JK, Nichols FH (1988) Food availability controls seasonal cycle of growth in Macoma balthica (L.) in San Francisco Bay, California. Journal of Experimental Marine Biology and Ecology, 116, 43±61. Thompson JK (1999) The effect of infaunal bivalve grazing on phytoplankton bloom development in south San Francisco Bay. PhD thesis, Stanford University, 419pp. Trush SF (1986) The sublittoral macrobenthic community structure of an Irish sea lough: effect of decomposing accumulations of seaweed. Journal of Experimental Marine Biology and Ecology, 96, 199±222. Tunberg BG, Nelson WG (1998) Do climatic oscillations influence cyclical patterns of soft bottom macrobenthic communities on the Swedish west coast. Marine Ecology Progress Series, 170, 85±94.

Ulanowicz RE, Tuttle JH (1992) The trophic consequences of oyster stock rehabilitation in Chesapeake Bay. Estuaries, 15, 257±265. Valiela I, McClelland J, Hauxwell J et al. (1997) Macroalgal blooms in shallow estuaries: control and ecophysiological and ecosystem consequences. Limnology and Oceanography, 42 (5, 2), 1105±1118. Van der Veer HW, Bergman MJN, Beukema JJ (1985) Dredging activities in the Dutch Wadden Sea: effects on macrobenthic infauna. Netherlands Journal of Sea Research, 19, 183±190. Wallentinus I (1984) Comparisons of nutrient uptake rates for Baltic macroalgae with different thallus morphologies. Marine Biology, 80, 215±225. Ware DM (1995) A century and a half of change in the climate of the NE Pacific. Fisheries Oceanography, 4, 267±277. Wassmann P (1991) Dynamics of primary production and sedimentation in shallow fjords and poll of western Norway. Oceanography and Marine Biology: an Annual Review, 29, 87±154. Weigelt M (1991) Short and long term changes in the benthic community of the deeper part of Kiel Bay (Western Baltic) due to oxygen depletion and eutrophication. Meeresforschung, 33, 197±244. Weston DP (1990) Quantitative examination of macrobenthic community changes along an organic enrichment gradient. Marine Ecology Progress Series, 61, 233±244. Widbom B, Elmgren R (1988) Response of benthic meiofauna to nutrient enrichment of experimental marine ecosystems. Marine Ecology Progress Series, 42, 257±268. Wildish D, Kristmanson D (1997) Benthic Suspension Feeders and Flow. Cambridge University Press, Cambridge, 409 pp.

ß 2002 Blackwell Science Ltd, Global Change Biology, 8, 813±830