shellfish aquaculture and the environment (shumway/shellfish aquaculture and the environment) ||...

Chapter 7

Bivalve s hellfi sh a quaculture and e utrophication JoAnn M. Burkholder and Sandra E. Shumway

Summary

Increased nutrient supplies and associated pol-lutants from land - based human activities have pervasively degraded coastal ecosystems worldwide, destroying habitats, causing fi nfi sh and shellfi sh disease and death, and promoting harmful algal blooms. While these effects of cultural eutrophication (anthropogenic nutri-ent overenrichment) from land - based human activities are well known, recent controversy has focused on another source of nutrient enrichment, aquaculture, as potentially a major contributor to eutrophication. Here we evaluate the signifi cance of bivalve shellfi sh aquaculture in the eutrophication of coastal waters based on the available evidence and,

conversely, the impacts of land - based nutrient pollution and associated pollutants on bivalve aquaculture.

Of the 62 ecosystems reviewed here, ∼ 7% or four ecosystems have sustained system - level adverse impacts from large, intensive bivalve culture operations. The other 93% have sus-tained either negligible or only localized sig-nifi cant adverse effects contributing to eutrophication from bivalve shellfi sh aquacul-ture. Thus, the great majority of ecosystems with bivalve aquaculture studied to date have been described as sustaining minimal or only localized signifi cant eutrophication effects from shellfi sh farming. Instead, the utility of bivalve aquaculture in effectively reducing phytoplankton and the nutrients available for

Shellfi sh Aquaculture and the Environment, First Edition. Edited by Sandra E. Shumway.© 2011 John Wiley & Sons, Inc. Published 2011 by John Wiley & Sons, Inc.

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156 Shellfi sh Aquaculture and the Environment

contribute little to eutrophication except in some poorly fl ushed areas with high shellfi sh density, and aquaculturists should strive to maintain cultures below ecological carrying capacity to prevent such ecosystem - level adverse effects. In contrast to the generally minimal effects of bivalve aquaculture on eutrophication, major, pervasive nutrient pol-lution from many urban and agricultural sources is seriously affecting shellfi sh popula-tions and shellfi sh aquaculture in many coastal waters of the world, and these impacts are expected to increase with rapidly expanding coastal development. Considering that shell-fi sh aquaculture is vital to meet the seafood demands of the rapidly increasing global human population, there is a pressing need for resource managers and policymakers to increase protection of shellfi sh aquacul-ture operations from land - based nutrient pollution.

Introduction

Cultural eutrophication, the process of over-enriching of surface waters with excessive nutrients from human activities, is among the most serious recognized threats to present - day coastal ecosystems (National Research Council [NRC] 2000 ; GEOHAB, Global Ecology and Oceanography of Harmful Algal Blooms Programme 2006 ). Increased nutrient supplies from land - based human activities have perva-sively degraded coastal ecosystems to the extent that more than two - thirds of coastal rivers and bays in the United States are now moderately to severely degraded from cultural eutrophication, exacerbated by poor fl ushing and signifi cant human population growth in coastal areas (Bricker et al. 1999 , NRC 2000 ). The excessive nutrients can stimulate blooms of noxious and toxic algae, and increase water column turbidity from the algal overgrowth together with pollutants such as suspended sediments that accompany nutrient loading. The reduced light causes benefi cial seagrass

blooms is being harnessed by some nations for economic benefi t to offset nutrient overenrich-ment from land - based sources.

The effects of bivalve culture on the sur-rounding environment are site - specifi c and especially depend on the hydrography (fl ush-ing and water exchange) and shellfi sh density. Exceptions to minimal or localized effects have been documented at the ecosystem scale, mostly in poorly fl ushed lagoons with high - density shellfi sh culture. These exceptions underscore the need to consider the ecosys-tem ’ s carrying capacity, rather than only the carrying capacity for maximal shellfi sh pro-duction, in bivalve aquaculture over large areas within a given system. Such consider-ations increasingly are assisted by models such as the Farm Aquaculture Resource Management ( FARM ) model and the shellfi sh aquaculture waste model DEPOMOD.

In contrast to the localized effects generally reported for bivalve aquaculture, land - based sources of eutrophication have overwhelmed many coastal ecosystems worldwide, and coastal population growth and associated nutrient pollution are continuing to increase rapidly. The acute, obvious effects of urban and agricultural nutrient pollution, often accompanied by loadings of suspended sedi-ments, microbial pathogens, and toxic sub-stances, are fi sh kills and high - biomass algal blooms. Much more serious chronic impacts, however, include long - term shifts in nutrient supplies, increased “ dead zones ” of low - oxygen bottom water, loss of critical habitat such as seagrass meadows, stimulation of harmful algal species that are low in food quality, reduction of shellfi sh recruitment and grazing, and increased shellfi sh physiological stress, disease, and death. Increasing tempera-tures from present warming trends in climate change can stress shellfi sh, and would be expected to interact with pollution to weaken shellfi sh hosts further and facilitate pathogen attack.

Overall, relative to land - based pollution sources, bivalve aquaculture has been found to

Bivalve shellfi sh aquaculture and eutrophication 157

of balance, these multiple stressors interact to cause the disease and death of higher trophic level organisms such as wild fi nfi sh and shell-fi sh (Wiegner et al. 2003 ).

While the effects of eutrophication from land - based human activities are well known, recent controversy has focused on another source of nutrient overenrichment, aquacul-ture, as potentially a major contributor to cul-tural eutrophication (Goldberg and Triplett 1997 ; Kaiser et al. 1998 ; Newell 2004 ; Richard 2004 ) (Fig. 7.1 ). Finfi sh culture requires direct inputs of nutrient - rich feed, whereas shellfi sh culture relies mostly on naturally occurring phytoplankton and suspended matter, with no

meadows to die because of depressed photo-synthesis (Burkholder et al. 2007 and refer-ences therein). In addition, while the excess algae photosynthesize and increase the dis-solved oxygen (DO) during the day, at night their respiration can deplete most or all of the available oxygen so that fi sh suffocate to death (Breitburg 2002 ). As the algal blooms senesce and die, their decom position contributes to this oxygen “ sag. ” Harmful microbial pathogens — viruses, bacteria, fungi, and protozoans — also frequently are added along with nutrient overenrichment (Burkholder et al. 1997 , and references therein; Paul and Meyer 2001 ). As the ecosystem is driven out

Figure 7.1 Conceptual diagram of land - based nutrient sources from watershed runoff and atmospheric pollution (nitrogen [N] as inorganic forms [N i ≡ nitrate, ammonium], particulate N, dissolved organic N [DON]; phosphorus [P]; and carbon, especially considering dissolved organic carbon [DOC]), and bivalve shellfi sh aquaculture (bivalves) interactions with nutrient supplies in coastal ecosystems as related to (1) removal of seston (suspended particulate material) during fi lter feeding, (2) biodeposition of feces and pseudofeces, (3) excretion of nutrients (especially ammonia, which is ionized to ammonium; also phosphate and various organic nutrient forms), (4) removal of N, P, and organic carbon in bivalve harvest, and (5) resuspension of nutrients, detritus, and sediments into the water column during some harvesting activi-ties. (Modifi ed from Cranford et al. 2006 .) Note that 1 ° = primary; and that this diagram depicts oxygenated (aerobic) sediments beneath the shellfi sh cultures, although localized impacts often include anoxic sediments beneath farms in poorly fl ushed areas.

Land-basedatmospheric

Land-basedrunoff

Sunlight

Detritus

Phytoplankton

Sediment

Seston

Tidal

exchange

N, P, C

Benthic 1∞ Producers

Harvest

Nutrients

Grazers,

bacteria,

fungi

Biodeposits

Mix

ing

Bivalves

5

41

23

NO3–

NO3–

NO2–

PO4–3

NO2–

NH4+

NH4+

Organic N, P,C

DOC

Nitrification

DenitrificationN2

N2

Burial

N, P

Aerobicsediment

Anaerobicsediment

N,

P, O

2

Exc

hange


Kautsky 1989 ). Others describe extreme adverse impacts on benefi cial naturally occur-ring macrofauna and plankton (Mattsson and Lind é n 1983 , da Costa and Nalesso 2006 ). It has been suggested by some authors that, because of high regeneration of nutrients by shellfi sh, shellfi sh cultures may increase eutro-phication (Baudinet et al. 1990 ) by reducing nutrient limitation and stimulating algal growth rates (Prins et al. 1998 ). Decreased water movement and current velocity caused by the structural features of shellfi sh cultures (e.g., Nugues et al. 1996 ) would be expected to exacerbate these effects in the localized area. On the other hand, high densities of fi lter - feeding shellfi sh can depress phytoplank-ton biomass while promoting higher turnover rates (Doering et al. 1986 ; Sterner 1986 ; Doering 1989 , Asmus and Asmus 1991 ), and it has been argued that this control on phyto-plankton biomass can stabilize the ecosystem (Herman and Scholten 1990 ) as long as the algal assemblage does not escape control by shifts to species that are ineffi ciently fi ltered (Prins and Smaal 1994 ).

This chapter addresses two questions: How signifi cant is bivalve shellfi sh aquaculture in the eutrophication (nutrient pollution, oxygen defi cits) of coastal waters, based on present evidence? Conversely, what are the impacts of land - based nutrient pollution and association pollutants on bivalve aquaculture?

Most c ommonly r eported: l ocalized c hanges a ssociated with s hellfi sh a quaculture

General e ffects

In moderation, nutrient enrichment — regard-less of the source — promotes benefi cial incr-eases in phytoplankton and benthic algal production and, in turn, higher production of zooplankton, macroinvertebrates, fi nfi sh, and shellfi sh that use the primary (photosynthetic)

supplementary food added. Adverse effects of shellfi sh farming on the water column and benthic environments under and near subtidal mussel farms have been described as compara-tively much lower than those around salmon farms (Mirto et al. 2000 ; La Rosa et al. 2002 ; Yokoyama 2002 ; Crawford et al. 2003 ).

How does shellfi sh aquaculture affect nutri-ent enrichment of coastal areas? The general perception is that shellfi sh aquaculture is benign or benefi cial because it relies on ambient primary production, can improve water clarity and reduce nutrients and phytoplankton con-centrations through shellfi sh fi lter feeding, and does not require addition of fi sh or other food (Folke and Kautsky 1989 ; Crawford et al. 2003 ; Shumway et al. 2003 ). Shellfi sh fi lter feeding can depress phytoplankton biomass and alter phytoplankton assemblage structure, and the shellfi sh also can access carbon from the microbial loop through consumption of heterotrophic and mixotrophic bacteriovores (Lucas et al. 1987 ; Dupuy et al. 2000 ). In addi-tion, large macroinvertebrates and benthic fi shes sometimes respond positively to shellfi sh cultures (D ’ Amours et al. 2008a ; see also Chapter 9 in this book).

Nevertheless, the effects of shellfi sh culture on nutrient cycling and food web dynamics have received mixed reports. Increased nutri-ent supplies from shellfi sh biodeposits can promote phytoplankton and benthic algal growth, and the increased food supply can enhance shellfi sh growth (Weiss et al. 2002 ). In turn, the removal of phytoplankton by fi lter - feeding shellfi sh can effect a strong “ top - down ” control of eutrophication symptoms ( sensu Bricker et al. 2007 ), and the shellfi sh can also infl uence water column biogeochem-istry (Souchu et al. 2001 ). Some authors have reported that high densities of molluscs benefi -cially control eutrophication despite their addition of organic - rich biodeposits as feces and pseudofeces to the bottom sediments, and that mussel culture represents basically a self - regulated aquaculture system (Folke and


production directly or indirectly for food (Reitan et al. 1999 ; Paterson et al. 2003 ; Zeldis et al. 2008 ; Burkholder and Glibert 2011). But when added in excess, nutrient pollution can cause algal overgrowth. Nighttime respiration of the excess algal growth can cause oxygen depletion in bottom waters and sometimes throughout the water column. Decomposition of this excess production by aerobic bacteria and fungi can also lead to oxygen depletion. As more chronic, long - term effects, nutrient overenrichment promotes major shifts in the structure of plant and animal communities, often resulting in high biomass of a few toler-ant species and loss of overall biodiversity. Where aerobic surface sediments overlay deeper anaerobic sediments, microbially medi-ated, coupled nitrifi cation - denitrifi cation can convert organic and inorganic N from animal wastes, detritus, and other sources to nitrogen gas (N 2 ), which can effectively reduce the N available for most primary producers — until the microbial consortium depletes the sedi-ment oxygen content. At that point, the coupled nitrifi cation - denitrifi cation is inhib-ited, more phosphorus can be released to the water column, and toxic hydrogen sulfi de can begin to accumulate (see Newell 2004 , and references therein).

Shellfi sh beds take up chlorophyll a , seston, and particulate matter (particulate organic carbon, particulate organic nitrogen, and par-ticulate organic phosphorus — POC, PON, and POP, respectively), and tend to release ammo-nium, orthophosphate, and silicate (Dame et al. 1991 ; Prins and Smaal 1994 ). Filtration and biodeposition of shellfi sh is considered benefi cial to water quality by controlling phy-toplankton densities and sequestering nutri-ents that are removed from the system when shellfi sh are harvested, buried in the sediments, or lost through denitrifi cation (Kaspar et al. 1985 ; Newell et al. 2002 ). Bivalve shellfi sh enhance benthic/pelagic coupling through fi lter - feeding of phytoplankton, deposition of feces and pseudofeces to the sediments, and

increase of nutrient remineralization rates (Hatcher et al. 1994 ; Prins and Smaal 1994 ; Dame 1996 ). Thus, large densities of bivalves cultured in poorly fl ushed coastal waters can alter the pelagic - benthic energy fl uxes by depleting phytoplankton, zooplankton, and seston in the water column through fi lter feeding; by increasing sedimentation rates from biodeposition of feces and pseudofeces; and by decreasing oxygen, thereby changing sediment characteristics and benthic com-munity composition (Callier et al. 2008 ) (Table 7.1 ).

Nutrients such as nitrogen and phosphorus are excreted by shellfi sh and buried in the sedi-ments; a portion of this nutrient supply is also regenerated from the biodeposits and recycled back to the water column where it can support phytoplankton production (see Newell 2004 ). For example, bivalve molluscs have been esti-mated to digest and absorb about 50% of the particulate N that they fi lter from the water column (Newell and Jordan 1983 ); much of the absorbed fraction is used for tissue growth, but some is excreted, mostly as ammonium (Bayne and Hawkins 1992 ). Excretion rates of nutrients to the water column by shellfi sh can be substantial (e.g., Table 7.2 ), and shellfi sh have been reported to play a major role in benthic nutrient regeneration in coastal eco-systems through rapid and effi cient recycling of inorganic N and P to primary producers (Magni et al. 2000 ).

Algal production, including growth of certain harmful species, can be stimulated by excreta from some bivalve species. Intensive bivalve cultivation can alter the N:P nutrient stoichiometry and change the major N species to reduced forms, especially ammonia as well as certain organic forms, and these N forms are preferred by various harmful algae (e.g. Berg et al. 1997; Arzul et al. 2001 ; Glibert et al. 2005).

If the biodeposits from the cultured bivalves settle into aerobic sediments overlying anaerobic sediments, coupled nitrifi cation -

Table 7.1 Localized effects related to eutrophication that have been reported from bivalve shellfi sh culture, with examples of references.

Reduction or depletion of nanophytoplankton, zooplankton, and/or seston

Escaravage et al. (1989) ; Navarro et al. (1991) ; Perez Camacho et al. (1991) ; Newell and Shumway (1993) ; Dankers and Zuidema (1995) ; Boyd and Heasman (1998) , Heasman et al. (1998) ; Meeuwig et al. ( 1998 ); Pitcher and Calder (1998) ; Ogilvie et al. (2000) ; Pilditch et al. (2001) ; Souchu et al. (2001) ; Cranford et al. (2003) ; Dowd (2003) ; Condon (2005) ; Strohmeier et al. (2005) ; Banas et al. (2007)

Low risk of reduced food resources for fi lter feeders (Crawford et al. 2003 )

Effects not found (Fr é chette et al. 1991 ; Mojica and Nelson 1993 ; Murdoch and Oliver 1995 ; Danovaro et al. 2004 )

Increased water clarity that has promoted growth of benefi cial seagrasses

Deslous - Paoli et al. (1998)

Nutrient replenishment and/or enhanced phytoplankton productivity

Kaspar et al. (1985) ; Doering et al. (1986) ; Barranguet et al. (1994) ; Ball et al. (1997) Songsangjinda et al. ( 2000 )

Increased abundance of cyanobacteria under shellfi sh cultures

Mirto et al. (2000)

Higher POM under shellfi sh cultures than in a control site; and/or higher C : N ratios under the rafts indicated accumulation of refractory POM mostly from feces and decomposing mussels

Mojica and Nelson (1993) , Nugues et al. (1996) , Chivilev and Ivanov (1997) , Mirto et al. (2000) , Stenton - Dozey et al. (2001) , Bendell - Young (2006) , Metzger et al. (2007) , Lu and Grant (2008)

Higher total N, organic N, dissolved organic C, chlorophyll a , and/or phaeopigment concentrations in surfi cial sediments from bivalve biodeposition

Ito and Imai (1955) ; Grenz et al. (1991) ; Nugues et al. (1996) ; Mirto et al. (2000) ; Condon (2005) ; Giles et al. (2006) ; Munroe and McKinley (2007)

Not found (Mojica and Nelson 1993 )

Increased sedimentation from biodeposition of shellfi sh feces and pseudofeces, increased organic carbon content of sediments, and/or altered sediment geochemistry

Ito and Imai (1955) ; Mattsson and Lind é n (1983) ; Rosenberg and Loo (1983) ; Escaravage et al. (1989) ; Baudinet et al. (1990) ; Grenz et al. (1991) ; Perez - Camacho et al. (1991) ; Hatcher et al. (1994) ; Grant et al. (1995) ; Nugues et al. (1996) ; Spencer et al. (1996, 1998) ; Songsangjinda et al. (2000) ; Chamberlain et al. (2001) ; Jie et al. (2001) ; Christensen et al. (2003) ; Crawford et al. (2003) ; Danovaro et al. (2004) ; Hartstein and Rowden (2004) ; Callier et al. (2006) ; Giles et al. ( 2006, 2009 ); Mallet et al. (2006) ; Callier et al. (2007, 2008) ; Weise et al. (2009)

Increased deposition of fecal matter and increased oxygen depletion from substantial biomass of fouling organisms such as ascidians on culture gear and shellfi sh stock

Stenton - Dozey et al. (1999)

Not found (no effect on DO) (Mojica and Nelson 1993 )

Altered redox values and/or higher benthic respiration

Baudinet et al. ( 1990 — respiration; but exceeded oxygen production only in 1 month over the annual study); Stenton - Dozey et al. (2001) ; Crawford et al. (2003)

Mallet et al. (2006) — no signifi cant differences between low - density culture and reference sites

160

Increased bottom - water turbidity Crawford et al. (2003)

Increased abundance of resident benthic infauna Kaiser et al. (1996) ; Spencer et al. 1996 )

Depressed abundance or biomass, lower species richness, and/or mortality of benthic macrofauna

Tenore et al. (1982) ; Mattsson and Lind é n (1983) ; Jaramillo et al. (1992) ; Barranguet et al. (1994) ; Grant et al. ( 1995 — although macrofaunal biomass sometimes was higher under the mussel cultures); Beadman et al. (2004)

Not found (Mojica and Nelson 1993 ; Yokoyama 2002 ; Danovaro et al. 2004 ; Whiteley and Bendell - Young 2007 )

Reduced macrofaunal biomass, alteration of trophic structure, and/or competition with indigenous species

Sauriau et al. (1989) ; Stenton - Dozey et al. (2001) ; Bendell - Young (2006)

Modifi cation of current patterns and circulation Ottmann and Sornin (1985) ; Nugues et al. (1996) ; Boyd and Heasman (1998) ; Crawford et al. (2003)

Changes in benthic macrofaunal community composition and/or diversity; often, a decrease in dominance of native bivalve molluscs and sea urchins offset by an increase in opportunistic polychaetes

Mattsson and Lind é n (1983) ; Kaspar et al. (1985) ; Castel et al. (1989) ; Baudinet et al. (1990) ; Grant et al. (1995) ; Spencer et al. (1996, 1998) ; Mirto et al. (2000) ; Chamberlain et al. (2001) ; Beadman et al. (2004) ; Miron et al. (2005) ; Bendell - Young (2006) ; da Costa and Nalesso (2006)

Not found (Kaiser et al. 1996 ; Danovaro et al. 2004 )

Higher abundance of benthic predatory shellfi sh such as crabs, demersal fi shes, or starfi sh which were attracted to cultured shellfi sh that fell to the bottom sediments beneath cultures, and/or to increased abundance of deposit - feeding prey organisms

Iglesias (1981) ; Romero et al. (1982) ; Rosenberg and Loo (1983) ; L ó pez - Jamar et al. (1984) ; Gonz á lez - Gurriar á n (1986) ; Freire et al. (1990) ; Grant et al. (1995) ; Inglis and Gust (2003) ; Smith and Shackley (2004) ; D ’ Amours et al. (2008b)

Higher average diversity in benthic macrofauna under a mussel site than a reference site; higher abundance of individuals at the reference site; overall, no negative effect

da Costa and Nalesso (2006)

Higher diversity of macrofauna per unit area of oyster cage culture areas than in reference areas with or without aquatic vegetation

Dealteris et al. (2004)

Higher bacterial abundance under cultures, and/or higher activities of bacterial exoenzymes under cultures; higher density and biomass of microbial assemblages under cultures

Grenz et al. (1990) ; Danovaro et al. (2004) ; La Rosa et al. (2002)

Extensive bacterial mats under cultures Dahlb ä ck and Gunnarsson (1981)

Elevated anoxia, anaerobic metabolism (seasonal), and/or higher oxygen consumption

Barranguet et al. (1994) ; Grant et al. (1995) ; Chivilev and Ivanov (1997) ; Chamberlain et al. (2001) ; Stenton - Dozey et al. (2001) ; Giles et al. ( 2006 — higher oxygen consumption in summer)

Elevated sediment sulfi de concentrations Ito and Imai (1955) ; Mattsson and Lind é n (1983) ; Mariojouls and Sornin (1987) ; Barranguet et al. (1994) ; Stenton - Dozey et al. (2001) ; Crawford et al. (2003)

Table 7.1 (Continued)

161


These effects are minimized in sites with more rapid fl ushing and water exchange, so that hydrography is a key factor infl uencing the environmental effects of shellfi sh aquacul-ture (Boyd and Heasman 1998 ; Crawford et al. 2003 ; Hartstein and Rowden 2004 ). For example, comparison of two mussel ( Mytilus edulis ) farms in sites with high versus low current velocity along the coast of southwest-ern Ireland revealed signifi cant alterations in benthic community structure at the site with low current velocity, but not at the higher - velocity site (Chamberlain et al. 2001 ). Similarly, open - sea mussel cultures had minimal detrimental effect on benthic fauna along the western Adriatic coast (Fabi et al. 2009 ). Smaal and Zurburg (1997) found no signifi cant release of nutrients from oysters

denitrifi cation also removes nitrogen from the sediments as N 2 gas. Thus, if the surface sedi-ments contain some oxygen, biodeposits from shellfi sh aquaculture promote net ecosystem losses of nitrogen and phosphorus not only by sediment burial but also by microbial nitrifi cation - denitrifi cation. This important process is inhibited, however, if copious biode-posits from high bivalve densities cause anoxia of the surface sediments (Newell 2004 ). Filtration of particulate matter by bivalves also reduces turbidity so that more light is available for benthic microalgae which take up ammo-nium, nitrate, and phosphorus, effectively reducing the substrates needed for nitrifi cation - denitrifi cation but also helping to control regeneration of sediment nutrients to the water column (Newell 2004 ).

Accelerated sulfur cycle (higher sulfate reduction and sulfur oxidation rates) under intensive shellfi sh cultures

Asami et al. (2005)

Localized elevated rates of ammonium release (seasonal)

Kaspar et al. (1985) ; Baudinet et al. (1990) ; Barranguet et al. (1994) ; Hatcher et al. (1994) ; Grant et al. (1995) ; Stenton - Dozey et al. (2001) ; Giles et al. (2006) ; Nizzoli et al. (2007)

In well - fl ushed areas, signifi cantly higher nitrate fl uxes under cultures

Giles et al. (2006)

Higher rates of N remineralization in surfi cial sediments under cultures

Grenz et al. (1990) ; Grenz et al. (1991) ; Hatcher et al. (1994) ; Gilbert et al. (1997) — dissimilatory ammonium production (98% of the nitrate in the farming area was reduced to NH 4 + and 2% to N 2 O; thus, most of the N i remained available in the ecosystem)

Higher denitrifi cation capacity in mussel farm sediments; lower nitrifi cation (oxidative process) in mussel farm sediments

Kaspar et al. (1985) ; Hatcher et al. (1994) ; Gilbert et al. (1997)

Higher phosphate and/or silicate fl uxes under cultures

Baudinet et al. ( 1990 — P, Si); Magni et al. ( 2000 — P); Souchu et al. ( 2001 — P)

Note: not found by Kaspar et al. ( 1985 — P) or by Hatcher et al. ( 1994 — P)

Mixed effects, apparently depending on local current patterns: at one large - scale mussel farm, but not at a second farm, high sedimentation and organic enrichment of the benthos, and reduced diversity of benthic infauna

Chamberlain et al. (2001)


Table 7.2 Comparison of nutrient excretion rates by different species of clams, mussels, oysters, and scallops based on some examples from the published literature.

Species, study area, units of nutrient excretion Approach NH 4 + NO 3 − + NO 2 − PO 4 − 3

Temperature ( ° C) or period Source

Clams

Donax serra

Algoa Bay, South Africa ( μ g NH 4 + ind. − 1 h − 1 ; means)

Lab./fi eld 0.31 – 5.20 nd nd 15 – 19 Cockcroft (1990)

Maitland River, South Africa ( μ mol NH 4 N g DW − 1 h − 1 )

Field 0.35 – 8.10 nd nd na Prosch and McLachlan (1984)

Donax sordidus

Sundays River, South Africa ( μ mol g DW − 1 h − 1 )

Lab./fi eld 2.9 nd nd 15 – 19 Cockcroft (1990)

Macoma balthica

Wadden Sea, Denmark ( μ mol g WW − 1 h − 1 )

Lab. 0.1 neg nd 13 – 15 Henriksen et al. (1983)

Mercenaria mercenaria

Delaware Bay, USA ( μ mol g DW − 1 h − 1 )

Lab. 0.9 – 1.5 neg nd 20 Srna and Baggaley (1976)

Cultured seed clams (from Mook Sea Farms, Inc., Damariscotta, ME) ( μ g NH 4 N g clam − 1 day − 1 )

Lab. 20.0 – 89.4 nd nd 20 Pfeiffer et al. (1999)

Tapes [ Ruditapes ] philippinarum

Hatchery, Ireland ( μ mol g DW − 1 h − 1 )

Lab. 0.16 – 1 nd nd 18.8 Xie and Burnell (1995)

Marennes - Ol é ron, France ( μ mol g DW − 1 h − 1 )

Lab. 0.5 – 13 nd nd 5 – 25 Goulletquer et al. (1989)

International Shellfi sh Enterprises, Moss Landing, USA (as Tapes japonica ; μ g NH 4 + N g live wt. − 1 day − 1 )

Lab. 27.84 – 72.24 nd nd 12, 14, 16, 18

Mann and Glomb (1978)

Seto Inland Sea, Japan ( μ mol g DW − 1 h − 1 )

Lab. 3.8 – 10.6 0 – 12.3 0.7 – 3.9 19.6 – 21.6 Magni et al. (2000)

Seto Inland Sea, Japan (mmol m − 2 day − 1 )

Field ext. 1.2 – 14 0.6 – 6.8 0.03 – 3.6

Annual Magni et al. (2000)

Virgin Islands, USA ( μ mol g DW − 1 h − 1 )

Lab. 1.9 – 4.9 nd nd 20.1 Langton et al. (1977)

163



Mussels

Choromytilus chorus

Queule Estuary, southern Chile ( μ g g DW − 1 h − 1 )

Lab. 3.49 – 16.22 nd nd 12 (winter) Navarro (1988)

Guekensia demissus

Great Sippewissett, MA, USA ( μ mol g − 1 DW h − 1 )

Lab. 2.5 nd nd Annual Jordan and Valiela (1982)

Near Duke University Marine Laboratory, Beaufort, NC, USA ( μ mol g − 1 day − 1 ; as Modiolus demissus )

Lab. 1.66 – 4.46 nd nd ∼ 23 Lum and Hammen (1964)

Narragansett Bay, RI, USA ( μ mol g DW − 1 h − 1 )

Lab. 0.26 nd nd 21 Nixon et al. (1976)

Mytilus edulis

Eastern Scheldt, the Netherlands (AFDW)

Field 0 – 13.9 neg 0.35 – 1.7

Jun – Oct Dame et al. (1991)

Eastern Wadden Sea, Germany (mmol m − 2 h − 1 )

Field 0.32 – 5.5 neg 0.85 Apr – Jun Asmus et al. (1990)

Linher River, UK ( μ mol g DW − 1 h − 1 )

Lab. 4.9 – 34.6 nd nd 11 – 21 Bayne and Scullard (1977)

Lynher River near Plymouth, UK ( μ g N g DW − 1 h − 1 )

Lab. ∼ 6 – 18 nd nd 8, 12, 15 Livingstone et al. (1979)

Lynher Estuary, southwestern UK (means; μ g N g DW − 1 h − 1 )

Lab. ∼ 8 – 39 nd nd 5, 10, 15, 20 (up to 25)

Widdows (1978)

Narragansett Bay, RI, USA ( μ mol g DW − 1 h − 1 )

Lab. 3.1 nd nd 15 Nixon et al. (1976)

Sound, Denmark ( μ mol g DW − 1 h − 1 )

Field 0.14 – 3.1 nd 0.10 – 0.53

0.7 – 18 Schl ü ter and Josefsen (1994)

Swansea Bay, UK ( μ g N g DW − 1 h − 1 )

Field 22.7 – 29.3 nd nd ∼ Annual Bayne et al. (1979)

Western Scheldt, the Netherlands ( μ mol g − 1 DW h − 1 )

Field 1.1 nd nd 12 Smaal et al. (1997)

Western Wadden Sea, the Netherlands (AFDW)

Field 0 – 13.9 neg 0.35 – 1.7

Jun - Oct Dame et al. (1991)


164



Western Wadden Sea, the Netherlands (mmol m − 2 h − 1 )

Field 1 – 3.7 neg 0.05 – 0.43

Apr – Sep Prins and Smaal (1994)

Mytilus galloprovincialis

R í a de Arousa, Galicia, northwest Spain ( μ mol g − 1 day − 1 ; intertidal and raft culture habitats)

Lab. 1.66 – 4.46 nd nd July Lum and Hammen (1964)

R í a de Arousa, Galicia, northwest Spain ( μ g NH 4 + N g DW − 1 h − 1 ; means, intertidal, and raft culture habitats, days 1 and 15)

Lab. 4.92 – 8.17 nd nd 14 – 15 (collected in July)

Labarta et al. (1997)

Musculista senhousia

Seto Inland Sea, Japan ( μ mol g DW − 1 h − 1 )

Lab. 9.3 – 16.9 0 – 1.9 1.3 – 5.5 19.6 – 21.6 Magni et al. (2000)

Seto Inland Sea, Japan (mmol m − 2 day − 1 )

Field ext. 0.23 – 24 0.03 – 2.7 0.06 – 6.5

Annual Magni et al. (2000)

Oysters

Crassostrea virginica

Tidal creeks, North Inlet Estuary, SC, USA (mg N g oyster − 1 h − 1 )

Field 0.39 nd nd July – Aug Dame and Libes ( 1993 )

Charlestown Pond, RI, USA ( μ mol g tissue − 1 day − 1 )

Lab. 1.56 – 2.19 nd nd na Hammen et al. (1966)

Chesapeake Bay, USA (mg N g oyster C − 1 day − 1 )

Modeled 1.43 nd nd Summer average

Cerco and Noel (2007)

Delaware Bay, USA ( μ mol g DW − 1 day − 1 )

Lab. 0.28 nd nd 20 Snra and Baggaley (1976)

Near Duke University Marine Laboratory, NC ( μ mol g tissue − 1 day − 1 )

Lab. 0.298 – 0.978 nd nd 22.0 – 25.5 Hammen (1968)

North Inlet Estuary, SC, USA ( μ mol m − 2 h − 1 )

Field 2825 – 15,304

0 – 0.2 nd 28 – 30 Dame et al. (1985)

Western Wadden Sea, the Netherlands (g N or P m − 2 h − 1 )

Field 0.04 – 0.11 nd 0.05 – 0.08

June Dame and Dankers (1988)

Crassostrea gigas

Bay of Pempoul (Bay of Morlaix), North Brittany, France ( μ mol g total WW − 1 h − 1 )

Field 0.28 – 6.7 nd nd ∼ 7.5 – 17 Boucher and Boucher - Rodoni (1988)


165




Sanggou Bay, north China ( μ mol h − 1 g DW − 1 ; NH 4 + N, PO 4 − 3 P)

Field ext. 0.51 – 5.40 nd 0.11 – 0.64

Jan., Jul. Mao et al. (2006)

Sea Salter Shellfi sh, Ltd., Whitstable, UK (juveniles; μ g NH 3 N g live wt. − 1 day)

Lab. 9 – 38.8 nd nd 12, 15, 18, 21

Mann (1979)

Ostrea edulis

International Shellfi sh Enterprises, Moss Landing, CA, USA (juveniles; μ g NH 3 N g live wt. − 1 day)

Lab. 11.6 – 21.2 nd nd 12, 15, 18, 21

Mann (1979)

Scallops

Argopecten irradians concentricus

Anclote Estuary, Tarpon Springs, FL, USA ( μ g NH 3 N g DW − 1 h − 1 ; means)

Field 72 – 140 nd nd 21.5 – 31.7 (May – Nov)

Barber and Blake (1985)

Homosassa, FL, USA ( μ g N mg AFDW − 1 h − 1 , as means; larvae and juveniles)

Lab. 0.125 – 0.384 nd nd 25 Lu et al. (1999)

Argopecten purpuratus

Bay of Hueihue, Chlo é , Chile ( μ g NH 4 + N g DW − 1 h − 1 ; means)

Lab. 19.7 – 41.9 nd nd 12 (annual mean)

Navarro and Gonzalez (1998)

Chlamys farreri

Xujia Maidao, Qingda, China ( μ g g DW − 1 h − 1 )

Lab. 178 – 147 nd nd 17, 23 Yang et al. (1999)

Modifi ed from Magni et al. (2000) .


AFDW, ash - free dry weight; DW, dry weight; fi eld ext., extrapolation of laboratory experiments to a fi eld community situation; lab., laboratory experiments; live wt., live weight; na, not available; neg, negligible excretion found; nd, not determined; WW, wet weight.

( Crassostrea gigas ) and mussels ( Mytilus edulis ) cultured on intertidal tables in Marennes - Ol é ron Bay along the coast of southwestern France, and reasoned that most of the biodeposits had been fl ushed away so that mineralization occurred elsewhere.

Localized effects of biodeposits were reported in sheltered sites of Marlborough Sounds, New Zealand, within 30 – 50 m from aquacul-ture operations, but at times wave action resuspended and dispersed the biodeposits over wide areas so that there was little overall


it was a more naturally enriched environment (Bourget and Messier 1982 ). In contrast, at HAM, clear localized effects included an increase in the percent organic matter of the sediment, and a decrease in benthic macrofau-nal diversity and abundance under the mussel lines. Other features of the study design, such as the position sampled within the shellfi sh farm, could also have a major infl uence on data interpretations; for example, sites within areas of older mussels can differ substantially from sites with cultures of younger animals (Callier et al. 2007 ).

Considering sediment characteristics, Sundb ä ck et al. (2000) reported that rates of nitrifi cation - denitrifi cation were about 10 - fold higher on an annual basis in fi ne - grained sedi-ments with abundant bioturbator fauna than in sediments with higher porosity and lower bioturbator biomass (see Chapter 10 in this book). Anoxia in even the surface sediments from overenrichment by bivalve deposits in shellfi sh culture areas can be a cumulative, chronic effect — that is, the longer the shellfi sh are cultivated in a given location, the more frequent and sustained the anoxic sediment conditions (e.g., Ito and Imai 1955 ). As a signifi cant consequence, the high hydrogen sulfi de concentrations in anoxic sediments can kill nitrifying bacteria, and even if the surface sediments can be reoxygenated, this microbial consortium must be restored before nitrifi cation can resume (Sloth et al. 1995 ). Some researchers have reported no sig-nifi cant localized impacts from suspended shellfi sh cultures on the underlying benthic environment. For example, Shaw (1998) found negligible effects of suspended mussel cultures in Prince Edward Island, Canada, based on a benthic sediment survey of organic matter and oxygen levels in 20 estuaries with culture and reference sites. Most published studies, in contrast, have described signifi cant but localized effects from suspended bivalve aquaculture on the underlying area and a limited area surrounding the farms (Tables 7.1 and 7.3 ).

effect on the sediments beneath the farms (Hartstein and Stevens 2005 ).

Thus, the effects of shellfi sh aquaculture on the surrounding environment are site - specifi c (Chamberlain et al. 2001 ). Accordingly, the detectable farm footprint (term from Giles et al. 2009 ) is mostly localized and can be minimal, slight, or severe, depending on an array of factors such as background enrich-ment, sediment composition and porosity, hydrographic features, depth, bottom topogra-phy, water exchange, abundance of bioturba-tor fauna, and culture density (Pietros and Rice 2003 ; Newell 2004 ; Gren et al. 2009 ). The degree and spatial extent of environmen-tal impacts is related to dispersal of the biode-posits (Chamberlain et al. 2001 ; Newell 2004 ), so it has been suggested that benthic effects of shellfi sh farms in high - energy environments should extend over larger areas (Giles et al. 2009 ). The substantial dilution in high - energy environments would “ lose the signal, ” however, so that detectable effects would likely remain localized.

The choice of sampling stations and param-eters of focus strongly infl uences interpreta-tions about how shellfi sh aquaculture affects coastal environments, as shown in an evalua-tion of two mussel farms versus control sites in Great - Entry ( GE ; area 2.5 km 2 , annual harvest 180 tonnes) and Havre - aux - Maisons ( HAM ; area 1.25 km 2 , annual harvest 160 tonnes) lagoons in the Magdalen Islands, Quebec, Canada (Callier et al. 2008 ). These two farms were the same age, and the lagoons where each was sited had similar average depth, current velocity, tidal range, tempera-ture, and salinity. In both farms, mussels were cultivated on suspended longlines over a 2 - year growout cycle. Yet contrasting patterns were found: GE, which had previously been described as a more naturally enriched envi-ronment (Bourget and Messier 1982 ), had low benthic species diversity, abundance and biomass, and sediment characteristics. The mussel operation there was assessed as having little apparent localized effect, possibly because

Table 7.3 Examples of locations and characteristics of bivalve culture systems for which effects have been reported.

Location Culture, wild stocks (information given) Reference(s)

Localized effects

Africa (1)

Small Bay within Saldanha Bay, South Africa (Benguela system) — current velocity averaged 7.5 cm s − 1 between rafts and 1.25 cm s − 1 within rafts; depth 12 – 15 m; water column strongly stratifi ed in summer

Mussels ( Mytilus galloprovincialis ) in intensive suspended culture ongoing for 10 years; culture area 80 ha; annual marketable yield 1880 – 2720 tonnes (2000 – 3000 tons) wet weight with shell; 320 ropes per raft

Stenton - Dozey et al. (1999, 2001) ; also Boyd and Heasman (1998) ; Heasman et al. (1998) ; Pitcher and Calder (1998)

Asia – Orient (7)

Lagoon of Takapoto Atoll, French Polynesia — within a nearly enclosed reef rim; mean depth 25 m; water residence time ∼ 4 years

“ Extensive ” pearl oyster ( Pinctada margaritifera ) culture for 20 years; ∼ 2 million cultured animals on down lines suspended on subsurface longlines

Niquil et al. (2001)

Mangoku - ura Inlet, Miyagi Prefecture, northern Japan — experimental suspended oyster farms — strong winds often suspended the bottom sediments

Harinohama farm (depth 4 m) and Sawada farm (depth 2.6 m)

Ito and Imai (1955)

Matsushima Bay, northern coast of Katsurashima Island, Japan — mean depth 4 m (at mean tide)

Suspended oyster cultures — 40 rafts maintained for 6 years, 150 rafts maintained for 3 years, and 20 racks maintained for 2 years

Ito and Imai (1955)

Nikolskaya Inlet, Kandalaksha Bay, White Sea — semi - enclosed and relatively shallow (mean depth 60 m); water residence time 5 – 6 years (Cobelo - Garc í a et al. 2006 )

Two suspended mussel farms ( Mytilus edulis ), each 15,000 m 2 , maintained for ∼ 10 years; depth 26 – 28 m or 13 – 16 m

Chivilev and Ivanov (1997)

Seto Inland Sea, southwestern Japan — sheltered waters that had sustained substantial land - based pollution (nutrients, other)

Clam ( Tapes philippinarum ) and mussel ( Musculista senhousia ) cultures

Magni et al. (2000)

Xuejiadao intertidal area in Jiaozhou Bay, eastern China — tidal range ∼ 1.4 m

Manila clam ( Tapes philippinarum ) farming transects (472 animals m − 2 ; ∼ 197 g wet weight m − 2 )

Jie et al. (2001)

Yamada Bay, Sanriku coast, northeastern Japan — semi - enclosed; area 30 km 2

Oysters ( ∼ 4,535 tonnes year − 1 ) and scallops — intensive aquaculture

Asami et al. (2005)

Australia – New Zealand (10)

Beatrix Bay (sheltered area; area 24 km 2 ; water current 6 – 12 cm s − 1 , depth ∼ 19 m at mid - tide, tidal range up to 4 m)

Longline cultures of greenshell mussels ( Perna canaliculus ); 45 farms cover ∼ 8.4% of the total bay area, 50 – 200 m from shore; total annual harvest ∼ 3630 tonnes (4000 tons)

Christensen et al. (2003)

Catherine Cove (CC; sheltered area; depth 25 – 42 m, area 1.8 ha)

Longline mussel cultures ( Perna canaliculus ) CC, EB — 11 – 12 longlines (length 3.5 – 5 km); cultures maintained for 15 years; BP — seven longlines; area 1.95 ha; cultures maintained for ∼ 3 years

Murdoch and Oliver (1995) , Ogilvie et al. (2000)

Elaine Bay (EB; moderately exposed; depth ∼ 30 m)

Blowhole Point (BP; exposed; depth 8 – 14 m, area 1.15 ha)

168


Elie Bay, Laverique Bay (sheltered; depth 15 – 25 m)

Longline mussel cultures ( Perna canaliculus ); two farms in Elie Bay had been in production 1 – 10 years; two farms in Laverique Bay had been in production 10 years

Inglis and Gust (2003)

Kenepuru Sound, a side - arm of Pelorus Sound — mean depth ∼ 11 m, 3 – 4 m tidal fl uctuation, maximum current speed ∼ 4 km h − 1

Longline mussel cultures ( Perna canaliculus ) in the center of a row of 5 similar farms that had been maintained during the past fi ve years

Kaspar et al. (1985)

Port Esperance ( PE ), St. Helens ( SH ), and Eaglehawk Bay ( EB ), Tasmania, Australia

Crawford et al. (2003)

PE — depth 8 – 12 m PE — mostly oyster suspended cultures ( Crassostrea gigas ), also mussels ( Mytilus planulatis ), maintained since 1984; total culture area 5.6 ha; average annual harvest ∼ 108 tonnes

SH — depth 7.5 – 9 m, average fl ow 3.8 cm s − 1 at 2 – 6 m above the seabed, and ∼ 18 cm s − 1 in the upper water column

SH — oyster and mussel cultures maintained since the mid - 1990s, expanded from 3 to 6 ha in 1999; average annual harvest ∼ 73 tonnes

EB — depth mostly 8 – 10 m (in some areas, 4 m), average fl ow 3.4 cm s − 1 at 4 – 8 m above the seabed

EB — mostly oysters with some mussels, maintained since the late 1970s; total culture area 13.8 ha; average annual harvest ∼ 220 tonnes

Western Firth of Thames, New Zealand — high - energy environment; current speeds from 3.1 to 32.8 cm s − 1 during fl ood tide and from 3.7 to 31.7 cm s − 1 during ebb tide

Mussel culture ( Perna canaliculus ), area 45 ha; maintained since 1994; 145 longlines in the upper 6 m of the water column; each of three blocks of longlines ∼ 1000 m long × 108 – 186 m wide; production 1520 tonnes (1680 tons) per harvest at 15 - month intervals

Giles et al. (2006)

Europe (17)

Arcachon Bay, southwestern France — total bay area 155 km 2 , mostly intertidal and channels; tidal range from 2.0 to 3.9 m

Oyster parks ( Crassostrea gigas ) covered 10 km 2 of the 40 - km 2 subtidal area

Castel et al. (1989) ; Escaravage et al. (1989)

Bangor Pier in the Menai Strait, north Wales

Mussels ( Mytilus edulis ) cultivated in the specifi c study area since 2000, by laying directly onto the substratum ( “ seabed ” ); large - scale experiment at four densities (2, 3, 5, 7.5 kg m − 2 ) on 400 - m 2 plots

Beadman et al. (2004)

Bay of Aiguillon and the Marennes - Oleron basin, western France

Oyster tables with bouchots (poles 4 – 6 m in height, sunk halfway into the sediments) in a tidal zone

Ottmann and Sornin (1985)


169


Canale, Bay of Mont - Saint - Michel, English Channel, northern France

Oyster tables with bouchots in a tidal zone

Ottmann and Sornin (1985)

Carlingford Lough — small embayment along the east coast of Ireland — shallow (mostly 2 – 10 m), with extensive areas exposed at low tide; maximum depth 35 m, maximum current speed 0.35 – 0.87 m s − 1 , tidal range 3.7 – 4.0 m

Intertidal oyster and clam cultures ( Crassostrea gigas — 400 tonnes; Tapes semidecussata — had not yet been harvested); and wild mussels ( Mytilus edulis — 1000 tonnes dredged in 1992)

Ball et al. (1997)

Carteau Bay in the Gulf of Fos, Mediterranean Sea — sheltered; mean depth 4.5 m, surface area ∼ 150 km 2

Mussels ( Mytilus galloprovincialis ) — intensive culture for 5 years (70 tables, each 15 m × 50 m, and each supporting 1000 – 1500 mussel ropes 3 – 4 m in length); cultivation area covered 0.4% of total cove area

Grenz (1989) ; Baudinet et al. (1990) ; Grenz et al. (1990, 1991) ; Barranguet et al. (1994) ; Barranguet (1997)

Cattolica - Rimini, Middle Adriatic Sea — ∼ 2.41 – 3.22 km from the coast (average depth 11 m)

Large longline mussel culture (species not given); farm area 2 km 2

Danovaro et al. (2004)

Coast of southwestern Ireland — low - energy environments (depth 12 – 15 m, mean current velocity 0.4 – 1.25 cm s − 1 )

Two longline mussel farms ( Mytilus edulis ) that had been in production for 8 – 14 years; each produced ∼ 90 – 135 tonnes (100 – 150 tons) of mussels year − 1

Chamberlain et al. (2001)

Gaeta Gulf, Tyrrhenian Sea, western Mediterranean Sea — microtidal ( ∼ 30 cm), sheltered site

Mussel cultures (annual production ∼ 360 tonnes [400 tons])

La Rosa et al. (2002)

Gaeta Gulf, Tyrrhenian Sea — sheltered area, microtidal ( ∼ 30 cm), turbid

Longline mussel cultures; annual biomass production ∼ 360 tonnes (400 tons)

Mirto et al. (2000)

Northwestern coast of Sweden — Island of Tj ä rn ö , northwestern coast of Sweden (depth 8 – 13 m, tide ∼ 0.3 m, generally weak currents averaging ∼ 3 cm s − 1 )

Longline mussel cultures ( Mytilus edulis ); total area 2800 m 2 ; yielded 100 tonnes in 18 months

Dahlb ä ck and Gunnarsson (1981)

Skagerrak archipelago northwestern coast of Sweden

Longline mussel culture ( Mytilus edulis ), 4500 m 2 (lines 180 m in length)

Rosenberg and Loo (1983)

Northern archipelago of the Swedish west coast — currents weak ( ∼ 3 cm s − 1 )

Longline mussel cultures Mattsson and Lind é n (1983)

Lysefjorden, southern Norway Mussel farm ( Mytilus edulis ), 200 m × 15 m

Strohmeier et al. (2005)

R í a de Arosa (Arousa) within R í as Bajas, Galicia, northwestern Spain — the cultivation area was sheltered by a natural sand bank

Mussels ( Mytilus edulis ) — more than 90,700 tonnes (100,000 tons) total wet weight year − 1 over a 230 - km 2 area (20 m × 20 m rafts; more than 2000 rafts in more than 30 areas, each with an average of 600 hanging ropes)

Iglesias (1981) ; Marino et al. (1982) ; Romero et al. (1982) ; Tenore et al. (1982) ; L ó pez - Jamar et al. (1984) ; Gonz á lez - Gurriar á n (1986) ; P é rez Camacho et al. (1991)


170


Sheltered sites in the inner estuary of R í a de Arosa

Mussel rafts ( Mytilus galloprovincialis ) — modeling study (2000 rafts)

Navarro et al. (1991)

R í a de Muros e Noia, Galicia, northwestern Spain — mean depth 25 m, area 90 km 2

Longline mussel cultures ( ∼ 70 rafts in two areas)

Gonz á lez - Gurriar á n (1986)

River Exe Estuary, Devon, UK — sheltered intertidal area; current velocity ∼ 30 cm s − 1 , current velocity outside the culture area ∼ 39 and ∼ 54 cm s − 1 in ebb tide and fl ood tide, respectively

Oysters ( Crassostrea gigas ) in intensive but relatively small - scale trestle cultivation; ∼ 110 – 120 oysters per bag (individual animal weight ∼ 25 – 40 g); depth less than 1 m

Nugues et al. (1996)

Southend - on - Sea, Whitstable, Kent, southeast England — shallow shelving mudfl at

Benthic cultures of Manila clams ( Tapes philippinarum ), harvested by suction dredging; ∼ 3500 clams m − 2

Kaiser et al. (1996)

Western inner Swansea Bay, Wales, UK — shallow, high - tidal - energy environment (maximum depth 20 m)

Mussel cultures ( Mytilus edulis ) since 1998; maximum abundance ∼ 600 m − 2

Smith and Shackley (2004)

Western Wadden Sea, the Netherlands — tidal range 0.9 – 1.9 m; 50% of the total area was exposed at low tide; water mass turned over every 5 – 15 days

Mussels ( Mytilus edulis ) cultures on subtidal fl ats since 1949; total culture area 70 km 2 ( = 5% of the total area); annual production ∼ 33 – 136 × 10 6 kg fresh weight

Dame et al. (1991) ; Dankers and Zuidema ( 1995 , and references therein)

North America (21)

Baynes Sound, British Columbia, Canada — on the eastern side of Vancouver Island, separated from the Strait of Georgia by Denman Island

Intensive culture of manila clams ( Tapes [ Venerupis ] philippinarum ) and oysters ( Crassostrea gigas ) along 90% of the intertidal beaches; clams were seeded on the beaches at a density of ∼ 400 animals m − 2

Jamieson et al. (2001) ; Bendell - Young (2006) ; Munroe and McKinley (2007)

Great - Entry Lagoon (GE) (Havre - aux - Maisons Lagoon [HAM]), Magdalen Islands, southern Quebec, Canada — average tidal range ∼ 0.6 m; typically weak currents (5 – 18 cm s − 1 ); strong winds effected a well - mixed water column; depth in culture areas ∼ 6 m

Suspended mussel farms ( Mytilus edulis ) in operation since the 1980s; at GE, cultures cover 2.5 km 2 (400 91 - m longlines) and yields ∼ 180 tonnes of mussels year − 1 ; at HAM, cultures cover 1.25 km 2 (200 76 - m longlines)

Callier et al. (2006, 2007, 2008)

Great - Entry Lagoon (GE) (House - Harbour Lagoon [HH]) in the Magdalen Islands; and Cascapedia Bay (CAS), southern Quebec, Canada (generally more exposed and energetic); mean tidal range 1.9 m; typical current velocity ∼ 10 cm s − 1 (note: HAM, above, is synonymous with HH)

Suspended mussel farms, as above for GE and HH; at the CAS site, the 1.4 - km 2 mussel culture area was ∼ 3.5 km offshore (depth 20 m) and consisted of 150 backlines (each 142 m long and supporting 1100 m of mussel line seeded at a density of 820 mussels m − 2 )

Weise et al. (2009)


171


Indian River Lagoon, FL — shallow, nontidal, wind - mixed lagoon; mean depth 1.5 m

Hard clam ( Mercenaria mercenaria ) aquaculture (benthic)

Mojica and Nelson (1993)

Malpeque Bay, Cascumpec Bay, and New London Bay, northern Prince Edward Island, Canada — sheltered areas, depth 6 – 15 m; mean height of high tide 0.9 m, mean height of low tide 0.2 m

Longline mussel cultures ( Mytilus edulis ); four farms, each consisting of a 100 - m horizontal surface rope from which mussel socks were suspended at depths from 3 – 10 m

D ’ Amours et al. (2008a)

Point Judith Pond, RI, USA — depth 2.4 – 3.0 m

Oyster cultures ( Crassostrea virginica ) — culture area 1 ha; more than 600 cages, each 0.6 m × 0.6 m and each holding 12 mesh bags of shellfi sh

Dealteris et al. (2004)

Sites along the coasts of Prince Edward Island, Canada

Longline mussel cultures ( Mytilus edulis )

Meeuwig et al. (1998)

Brudenell River: mean depth 14.5 m, water residence time 44 days

2.03 × 10 7 g dry weight standing stock

Broughton ’ s Creek: mean depth 18.0 m, water residence time 332 days


Cardigan Bay: mean depth 17.6 m, water residence time 255 days


Murray River: mean depth 10.0 m, water residence time 356 days


Rustico Bay: mean depth 7.1 m, water residence time 83 days


St. Peters Bay: mean depth 10.3 m, water residence time 61 days


St. - Simon Bay, New Brunswick, Canada — a shallow open bay with excellent water exchange; mean depth 0.6 m at low tide

Eastern oyster cultures ( Crassostrea virginica ; maximum density equivalent to 4000 oyster bags ha − 1 ; fi nal oyster biomass 8 kg m − 2 ); sampling area was a 35 - ha oyster lease used for bottom culture (1982 – 1997), then for fl oating bag and oyster table culture

Mallet et al. (2006) ; Lu and Grant (2008)

Tracadie Bay, northern Prince Edward Island, Canada — sheltered areas; mean tidal range ∼ 1 m

Modeling effort in an area of intensive longline mussel culture ( Mytilus edulis )

Dowd (2003)


172


Tracadie Bay ( TB ) and Winter Bay ( WB ), northern shore of Prince Edward Island, Canada, facing the Gulf of St. Lawrence — shallow, enclosed tidal lagoon; mean depth 3 m, maximum depth 6 m

Longline mussel cultures ( Mytilus edulis ) — ∼ 10 leases in WB; more than 40 plots in TB; overall, spat and market - sized mussel mussel production (year, 2000) was more than 210 tonnes (WB) and 4600 tonnes (TB)

Miron et al. (2005)

Upper South Cove, Lunenburg, Nova Scotia, Canada (Corkum ’ s Island Mussel Farm) — mean depth 1 – 2 m, maximum depth 10 m; protected area, with maximum current speeds 15 cm s − 1 at 1 m above the seabed

Longline mussel culture ( Mytilus edulis , also Mytilus trossulus ) in the upper 3 m of water column; culture area ∼ 4000 m 2 ; ∼ 400 mussels m − 2

Hatcher et al. (1994) ; Grant et al. (1995)

Whitehaven Harbour, northeastern Nova Scotia, Canada — mean tidal range 1.4 m; experimental site 0.5 km southeast of Munroe Point (mean depth 19 m)

Suspended cultures of sea scallops ( Placopecten magellanicus ) — cultivation area 80 m × 50 m; 10 longlines rigged with vertically hanging ropes spaced 1 m apart; average scallop density 12.2 ind. m − 3

Pilditch et al. (2001)

Willapa Bay, WA — shallow, coastal - plain, upwelling - infl uenced estuary; more than 50% of the surface area and volume were in the intertidal zone; tidal range 3.5 m; half of the bay volume entered and left with every tide, and 30% of the intertidal volume was replaced with new water on every tide; yet in the landward reach of the estuary, the average water age was 3 – 5 weeks

Intensive culture of Pacifi c oysters ( Crassostrea gigas )

Banas et al. (2007)

South America (2)

Coqueiro ’ s Beach, Anchieta, southeastern Brazil — average depth 3 m; tidal currents up to 25 cm s − 1 in winter and 35 cm s − 1 in summer

Longline mussel cultures ( Perna perna ), maintained since 1998 (100 lines; estimated annual production 24 tonnes)

da Costa and Nalesso (2006)

Queule River Estuary, southern Chile — central portion, depth ∼ 4 m

Mussel cultures ( Choromytilus chorus, Mytilus chilensis , up to 250 – 300 adults m − 2 )

Jaramillo et al. (1992)

Ecosystem - scale effects (4)

Cherrystone Inlet, eastern Chesapeake Bay, USA — small, shallow embayment (6 km 2 , mean depth 1 m; volume 1.54 × 10 7 m 3 at high tide; average volume [time - averaged tidal prism] 5.8 × 10 6 m 3 )

Hard clam ( Mercenaria mercenaria ) benthic cultures for ∼ 25 years (4.5 × 10 7 adults, at or near carrying capacity); annual harvest 2 × 10 7 clams (average shell length 60 mm)

Kuo et al. (1998); Luckenbach and Wang ( 2004a,b ); Condon (2005)


173



Marennes - Ol é ron Bay, western Atlantic coast of France — shallow macrotidal bay with high turbidity from suspended sediments; area ∼ 136 km 2 (Dame (1996) ; mean depth ∼ 5 m; water residence time 5 – 10 days; bivalve populations in the bay could fi lter the water volume in ∼ 2.7 days (Smaal and Prins 1993 )

Mudfl ats (700 – 800 m in width) with mussel culture structures (bouchots) and oyster racks ( Mytilus edulis , Crassostrea gigas ), operated since 1850; shellfi sh cultures covered ∼ 60% of the bay area ( ∼ 60 km 2 mussel cultures, 32 km 2 oyster cultures) and had added high levels of biodeposits; on at least two occasions, this system had been overstocked with shellfi sh and overexploited; standing stock of Crassostrea gigas was ∼ 100,000 tonnes fresh weight; annual production of Crassostrea gigas was ∼ 30,000 tonnes fresh weight

Sornin et al. (1983) ; Sornin et al. (1986) ; Sauriau et al. (1989) ; H é ral (1993) ; Raillard and M é nesguen (1994) ; Smaal and Zurburg (1997) ; Bacher et al. (1998) ; Dame and Prins (1998) ; Gouleau et al. (2000)

Sacca di Goro Lagoon, Northern Adriatic Sea, Italy — had two 900 - m openings; mean depth 1.5 m area 26 km 2 ; water residence time up to 25 days; freshwater fl ows were highly managed with dredging

Intensive suspended clam ( Tapes philippinarum ) and mussel ( Mytilus galloprovincialis ) culture; as of 2004, approximately one - third ( ∼ 8 km 2 ) of the total lagoon area was licensed for clam culture and 0.4 km 2 was licensed for mussel culture; annual harvest ∼ 13,600 tonnes (15,000 tons)of clams and 900 tonnes (1000 tons)of mussels

Mazouni et al. (1996) ; Viaroli et al. (1996) ; Mazouni et al. (1998) ; Bartoli et al. (2001) ; Mazouni et al. (2001) ; Viaroli et al. (2003) ; Mazouni (2004) ; Nizzoli et al. (2005, 2006a, 2006b, 2007)

Thau Lagoon, France — had two small openings; mean depth 4.5 m; had excessive nutrient inputs from the watershed, very low water renewal (water residence time ∼ 3 – 5 months; limited rainfall, warm temperatures; wind - mixed and microtidal with bottom - water anoxia in summer)

Intensive suspended oyster ( Crassostrea gigas — 80%) and mussel ( Mytilus galloprovincialis — 20%) culture covered ∼ 20% ( ∼ 15 km 2 ) of the total lagoon area (40 oysters m − 2 ; total standing stock 22,670 tonnes [25,000 tons]; annual harvest ∼ 13,600 tonnes [15,000 tons])

Deslous - Paoli et al. (1993, 1998) ; Mazouni et al. (1996, 1998, 2001) ; De Casabianca et al. (1997) ; Gilbert et al. (1997) ; Souchu et al. (2001) ; Mazouni (2004) ; Mesnage et al. (2007) ; Metzger et al. (2007)

Note that the number in parentheses after each region indicates different ecosystems studied. Note that some studies that are described in the text are not included here (e.g., Prins and Smaal 1994 ) because specifi cs about the aquaculture were not provided.


Typical fi ndings are illustrated by Baudinet et al. ’ s (1990) or Grant et al. ’ s (1995) compari-sons of an area beneath a suspended mussel culture ( Mytilus galloprovincialis ) versus a ref-erence site. Working in a well - mixed area of the Gulf of Fos, France, Baudinet et al. (1990) examined upward fl uxes of nitrate, nitrite, ammonia, silicate, and phosphate as well as oxygen at the sediment – water interface. Both

the culture area and the reference site had organic - rich sediments, especially the culture area because of high deposition of fecal matter. Transformation of biodeposited organic matter seasonally increased phosphate, silicate, and ammonia fl uxes, whereas there were neg-ligible or minor differences in nitrate/nitrite fl uxes and the oxygen production/consumption equilibrium. The ratio of Si : P fl uxes (as


luscs Ilyanassa spp. and Nuca tenusucata , which apparently were attracted to mussels that were lost from the culture, and to enriched organic matter in the biodeposits. The authors ’ overall assessment was that there were minimal adverse impacts on the surrounding area from bivalve aquaculture.

Severe localized negative effects of bivalve shellfi sh farming are infrequently reported. For example, Dahlb ä ck and Gunnarsson ( 1981 ) reported severe localized effects from intensive blue mussel farming in a poorly fl ushed area along the northern Swedish coast (water depth 8 – 13 m, tide ca. 0.3 m, weak currents at ∼ 3 cm s − 1 ). The sedimentation rate (3 g C m − 2 day − 1 ) under the culture area was about threefold higher than at a nearby reference site. There was accumulation of sedi-ment rich in organic matter and sulfi de: at 15 ° C, 30.5 mmol sulfate m − 2 day − 1 in the upper 10 cm of sediment under mussel cultures, and much higher sulfi de seasonally. As a second example, Stenton - Dozey et al. (2001) mea-sured signifi cantly higher particulate organic matter ( POM ) and sediment anoxia under mussel ( Mytilus galloprovincialis ) rafts than in a control reference site in a South African bay (fi ne to medium sand sediments, average current velocity 1.25 cm s − 1 within rafts and 7.5 cm s − 1 between rafts). The refractory POM from mussel feces, decomposing dead mussels, and biofouling organisms caused elevated C : N ratios. Total reducible sulfi des also increased threefold downcore in sediments under mussel rafts. Highest and most variable rates of ammonium effl ux occurred under the mussel cultures (825 ± 500 mmol NH 4 + m − 2 h − 1 ) as well. Other localized impacts can include depressed biomass and altered community structure of benthic macrofauna (Table 7.1 ).

Localized effects can persist for some time after farming is discontinued, especially in poorly fl ushed areas. Martin et al. (1991) reported that biodeposits accumulated in sandy sediments beneath oyster tables, but were no longer detected 2 months after the

Si(OH) 4 : PO 4 − 3 ) ranged from 6 to 15, much lower than the ratio of the concentrations of these nutrients in the water column (18 – 70), suggesting that the fl ux values mainly refl ected decomposition of recently sedimented phyto-plankton. Benthic respiration and consump-tion of oxygen exceeded oxygen production under the mussel cultures only in 1 of 5 months when surveys were conducted. Reference site macrofauna (adult invertebrates larger than 0.5 mm) were poorly diversifi ed (Shannon Index between 2 and 3) and consisted mostly of polychaete detritivores (40% Cirratulidae, 30% Capitellidae — mainly Mediomastus sp.), with maximal densities at 50,000 m − 2 . The redox break line (aerobic/anaerobic boundary) was fairly deep and occurred at ca. 10 mm. Under the mussel cultures, there was yet lower diversity of benthic macrofauna (Shannon Index between 1.5 and 2). Nearly all of the fauna were polychaetes that are considered to be indicative of high organic pollution — for example, 10,000 Capitella capitata m − 2 , and 60,000 Ophryotrocha sp. m − 2 , small species which exist in the surfi cial sediments and consume biodeposits. The redox break line was much shallower, only ca. 2 mm from the sediment surface. The authors ’ overall interpretation from their data was that biodeposits from high - density mussel cultures cause localized rather than ecosystem - level effects.

Grant et al. (1995) compared the area under a suspended mussel culture ( Mytilus edulis, Mytilus trossulus ) versus a reference site of similar sediment texture in a small Nova Scotia cove (7 - m depth): Sedimentation rates of mussel feces and pseudofeces were higher under the mussel culture lines; there was a shift toward more anaerobic metabolism at the mussel site; and maximum rates of ammonium release at the mussel site were twofold higher than maximal rates at the reference site. Benthic macrofauna also differed in abun-dance and community structure at the two sites; the mussel site was dominated by mol-


Special mention is merited of a major con-troversy in Puget Sound of the northwestern United States concerning the environmental impacts of the Pacifi c geoduck ( Panopea abrupta ; Fig. 7.2 ). Geoducks are the largest known burrowing clams; they range from 0.5 to 1.5 kg at maturity, but can weigh up to 7.5 kg with a siphon length of more than 1 m (Hilborn et al. 2004 ). Market size in Washington State is just under 1 kg (Hilborn et al. 2004 ). Geoduck aquaculturists use poly-vinyl chloride (PVC) pipes (length ca. 360 mm, diameter ca. 152 mm) pushed into intertidal sediments as predator exclusion devices or “ nursery tubes ” to protect the seed stock, which are set in high densities (ca. 20,000 to 43,500 PVC pipes per acre, or ∼ 8100 to ∼ 17,600 pipes per hectare; W. Dewey, Taylor Shellfi sh Farms, Inc., Shelton, WA, pers. comm.). Considering their size and the typical density of culture, geoducks could be signifi -cant contributors to nutrient supplies in

cultures were removed. On the other hand, Mattsson and Lind é n (1983) found that the benthic fauna at a site along the Swedish west coast was dominated in number by the clam Nucula nitidosa , and in biomass by the heart urchin Echinocardium cordatum and brittle stars Ophiura spp. prior to initiation of mussel farming. After 6 – 15 months of mussel aqua-culture, these species were replaced by oppor-tunistic polychaetes ( Capitella capitata , Scolelepis fuliginosa , Microphthalmus sczel-kowii ), localized within a zone under and 5 – 20 m around the cultures. After the mussels were harvested, only limited recovery was observed in the localized affected area after 1.5 years. In Saldanha Bay, South Africa, Stenton - Dozey et al. (1999, 2001) documented accu-mulation of refractory organic matter under mussel ( Mytilus galloprovincialis ) rafts and adverse effects on benthic macrofauna, with only marginal recovery 4 years after the cul-tures were removed.

Figure 7.2 (A, B) Geoduck farming operations (nogeoduckfarm.com/_wsn/page2.html and www.protectourshoreline.org ); (C) geoduck ( www.pac.dfo - mpo.gc.ca/ … /geopath/intro - eng.htm ).

(A) (B)

(C)


P) [is] continuously supplied from outside, molluscs meet their food requirements in situ ” (Bartoli et al. 2001 ). Although there are local-ized increases in nutrient supplies in shellfi sh aquaculture sites, the total amount of nutrients regenerated from bivalve biodeposits is com-parable with the nutrients that would be released from other means of phytoplankton decomposition. Thus, whereas fi nfi sh aquacul-ture is driven by the addition of large quanti-ties of nutrients within fi sh food, in shellfi sh aquaculture sites, the maximum phytoplank-ton biomass supported by nutrients that are regenerated from shellfi sh biodeposits cannot exceed the level that would be sus-tained by ambient nutrients (see Newell et al. 2005 ).

Localized e ffects on n utrient d ynamics

The most commonly reported infl uence of bivalve shellfi sh aquaculture on nutrient supplies is high rates of ammonium fl ux (direct excretion plus regeneration from biodeposits in the sediments). Maximal rates of 5 mmol N m − 2 h − 1 have been reported, espe-cially during warmer seasons (see review in Newell 2004 ). Sources of the nitrogen released include not only N - rich phytoplankton but also bacteria and heterotrophic fl agellates (Asmus and Asmus 1991 ; Kreeger and Newell 2001 ). Considering these high levels of ammo-nium regeneration, a common interpretation has been that bivalve populations rapidly recycle nutrients and thus enhance phyto-plankton production.

For example, Magni et al. (2000) assessed the effects of organic loads from biodeposits of a mussel farm ( Tapes philippinarum ) in the Tyrrhenian Sea, a poorly fl ushed area along the western Mediterranean. Their analysis indicated that direct excretion of N and P by the mussels accounted for up to 90% of benthic nutrient fl uxes in this system over an

farm areas, as evidenced by the abundant mac-roalgae that commonly thrive in the localized areas (Fig. 7.2 ). However, the environmental effects of geoduck aquaculture remain to be determined — a white paper commissioned by the state of Washington to determine environ-mental impacts of geoduck aquaculture reported that there was no peer - reviewed research on eutrophication effects as of late 2007 (Straus et al. 2008 ), and information is still not available.

The available research about the effects of shellfi sh aquaculture on eutrophication mostly has focused on oysters or mussels that are suspended in the water column or attached to supports. Infl uences of cultured shellfi sh such as clams, which are grown directly within the bottom sediments, are logistically more diffi -cult to assess, but similarly have been reported to include displacement of the natural benthic macrofauna and loss or replacement of their associated functions (e.g., sediment mixing or oxygen transport into anoxic sediments), increased oxygen consumption, excretion of nitrogenous organic wastes, enhanced sedi-mentation of suspended particulate matter, and concentration of biodeposits in the farm areas that promotes sulfi de accumulation in the surfi cial sediments (Sorokin et al. 1999 ; Bartoli et al. 2001 ). Clam harvesting tech-niques such as dredging alter sediment strati-fi cation and can increase nutrient fl uxes to the overlying water (Kaiser et al. 1998 ; Bartoli et al. 2001 ; see also Chapter 11 of this book).

Despite major emphasis on assessing the role of suspended mussel farming in eutrophi-cation, the approach of most studies has been described as consisting mostly of “ intuitive considerations about the relationships between mollusc culture and eutrophication processes: the collection and removal from the system of organisms that feed on microalgae and par-ticulate matter should result in a net reduc-tion of nutrient loads. … This consideration is based on the assumption that, contrary to fi sh farming systems in which food (and so, N and


1994 ), and removal of N from the ecosystem by denitrifi cation under suitable conditions, can be enhanced by bivalve biodeposition (Newell et al. 2002 ; Newell et al. 2005 ).

Beyond a commonly reported increase in ammonia fl ux and organic carbon concentra-tions beneath suspended bivalve shellfi sh aquaculture sites, and estimates of effects on primary producers in the culture area, few attempts have been made to evaluate localized effects of bivalve aquaculture on nutrient pools and transformations. Among the most detailed studies is that of Kaspar et al. (1985) , who attempted to assess the nitrogen cycle seasonally over an annual cycle at a mussel ( Perna canaliculus ) farm versus a nearby refer-ence site in Kenepuru Sound, New Zealand (Table 7.4 ). Nitrate + nitrite pools were similar at the two sites, but sediment ammonium was

annual cycle. In other work, a small suspended mussel culture operation ( Perna canaliculus , 45 ha) in a well - fl ushed area of the Firth of Thames, New Zealand, was estimated to con-tribute substantially to the inorganic N sup-plies needed by primary producers in the localized area (Giles et al. 2006 ) (Table 7.1 ). At a reference site, sediment dissolved inor-ganic N ( DIN ) release was calculated to supply ∼ 74% of the N requirements of the primary producers, indicating that benthic nutrient regeneration was important in supporting primary production for this system (Zeldis 2005 ; Giles et al. 2006 ). Under the mussel farm, in contrast, DIN release would have met an estimated 94% of the N requirements for primary producers. In addition to nutrient recycling, burial of N and P with accumulating sediments (Kaspar et al. 1985 ; Hatcher et al.

Table 7.4 Comparison of nitrogen dynamics at a mussel culture site versus a nearby reference site in Kenepuru Sound, New Zealand, based on seasonal measurements over an annual cycle (from Kaspar et al. 1985 ).

Parameter Reference site Mussel culture site

Organic N (top 12 cm of sediment) 7.4 – 10.8 mol m − 2 6.1 – 8.9 mol m − 2

Nitrate + nitrite pools - - - - - - Similar at both sites - - - - - -

Sediment ammonium (surface; 12 cm) 86 nm cm − 3 ; 112 nm cm − 3 418 nm cm − 3 ; 149 nm cm − 3

Molar C : N ratio of sediment organic matter 7.9 – 10.0 6.2 – 7.2

Molar N : P ratio of sediment organic matter 3.3 – 6.1 4.3 – 7.2

Total N mineralization rate (top 12 cm of sediment) 8.5 – 25.0 mmol m − 2 day − 1 21.7 – 37.1 mmol m − 2 day − 1

Sediment denitrifi cation 0.1 – 0.9 mmol m − 2 day − 1 0.7 – 6.1 mmol m − 2 day − 1

Percentage of reduced nitrate that was denitrifi ed (sediment)

93% 76%

Ammonium excretion by mussels — 4.7% (summer) and 7.4% (winter) of the combined N mineralization by mussels + sediment

Chlorophyll a (surface sediments) ∼ 75 mg m − 2 ∼ 75 mg m − 2

Phaeophytin levels (surface sediments) 52 mg m − 2 137 mg m − 2

Phytoplankton productivity (comparable except in winter [shown] — 14 C - CO 2 fi xation)

232.5 mmol m − 2 day − 1 1183 mmol m − 2 day − 1


fates often characterize lagoons and embay-ments where most shellfi sh aquaculture occurs. Concentrations are relatively easy to measure, but mass water trans port, needed to esti-mate nutrient loadings (Burkholder et al. 2004, 2006 ), is more diffi cult to assess accurately.

In addition, an array of biogeochemical processes (assimilation, sedimentation, adsorp-tion, denitrifi cation, chemical reduction, bioturbation) — each a challenge to assess at the ecosystem level — infl uence nutrient trans-port and fate. While the net removal of nutri-ents from the system with shellfi sh harvest is relatively easy to quantify, the biogeochemical processes involved in nutrient assimilation, regeneration, and mobilization rates associ-ated with the metabolic activities of the shell-fi sh are more diffi cult to assess (Bartoli et al. 2001 ). Several attempts to evaluate the role of shellfi sh in eutrophication at the ecosystem scale are reviewed here. Thus far, as would be expected, the data indicate that the effects of bivalve shellfi sh aquaculture on coastal ecosys-tems depend on biogeochemical processes as well as the characteristics of the shellfi sh species being cultured (e.g., the species, size, age, health, density of animals per unit area, nutrient assimilation effi ciency — Bayne et al. 1976 ), the areal extent of the cultures in com-parison with the total system area, the water residence time and mixing characteristics, the balance between effective nutrient removal (via particle fi ltering) from the water column versus nutrient additions from sediment biode-posits, and the amount and types of other potential nutrient sources. Rather than attempting to rigorously assess each of the many variables — which would be cost - prohibitive for most scientists — the relatively few ecosystem - scale assessments have used a more simplistic or “ fi rst cut ” approach.

It is generally accepted that natural popula-tions of bivalve fi lter - feeders can be important in nitrogen cycling in coastal ecosystems, but

up to fi vefold higher in the mussel farm site; in addition, total denitrifi cation was 21% higher, and total loss of N through mussel harvest and denitrifi cation was 68% higher.

In Upper South Cove, Nova Scotia, Hatcher et al. (1994) examined the effects of enhanced sedimentation under suspended mussel cul-tures ( Mytilus edulis , Mytilus trossulus ) on the respiration and nutrient fl uxes of the benthic communities and benthic/pelagic coupling over an annual cycle. The reference site sedi-ments were a net sink for total dissolved N, whereas the sediments under the mussel lines were a source. A signifi cant relationship between water - column chlorophyll and sedi-mentation rates was enhanced under the mussel lines, and ammonium release was sig-nifi cantly higher under the mussel cultures throughout the year. However, most of the POM trapped in sediment traps at both the mussel farm site and the reference site did not contribute signifi cantly to C or N fl ux or long - term burial in the underlying sediments. The authors suggested that most of the POM likely had been transported from both sites as tidally driven horizontal fl ux.

Interpretations from an e cosystem a pproach

Logically, the potential for ecosystem - scale effects from bivalve aquaculture on nutrient supplies, phytoplankton blooms, oxygen defi -cits, and other factors associated with eutro-phication would be expected to increase in quiet, poorly fl ushed lagoons and embayments as the relative proportion of total area covered by the cultures increased. Although an ecosys-tem approach is needed to evaluate the effects of shellfi sh aquaculture on the surrounding ecosystem, few such studies exist in the pub-lished literature, likely because of the massive effort and cost involved in adequate assess-ment. Complex nutrient loading patterns and


system area, and/or the culture is sited in mod-erately fl ushed to well - fl ushed conditions. For example, Baudinet et al. (1990) used localized fl ux data collected at a reference site versus beneath suspended mussel cultures ( Mytilus galloprovincialis ; n = 3 hemispheric bell jars, 17 L in volume, per site) to estimate the effect of mussel farming on Carteau Cove, a small embayment in southern France along the Mediterranean Sea. The bivalve farm covered only about 0.4% of the total 13 km 2 area of the cove, and consisted of 70 rope hanging structures that had been in place for 3 years; each 15 × 50 m table supported 1000 – 1500 mussel ropes that were 3 – 4 m in length. The area was somewhat sheltered by a natural sandbank. The fl ux data were collected on fi ve dates during spring, summer, and/or fall of 2 years. There were signifi cant localized effects of the mussel farms on nutrient fl uxes and benthic infauna, but on an areal basis, the infl uence of the mussel cultivation area on all of the cove ranged from 1.4% increase (in Si(OH) 4 ) to 2.3% increase (in NH 4 + ) (Table 7.5 ).

As another example, Niquil et al. (2001) estimated the effects of farmed pearl oysters ( Pinctada margaritifera ) and associated bivalves (pen shells Pinctada maculata ) on the lagoon of Takapoto Atoll (Takapoto, French Polynesia), which has a relatively long water residence time of 4 years. Pearl oyster farming was described as extensively developed in the lagoon for 20 years, including stock of ∼ 2 million P. margaritifera and ∼ 10 million P. maculata on the culture systems, but the

their effects on P cycling and on N and P stoi-chiometry appear to be more variable (Nizzoli et al. 2006a, 2006b ). Some authors have not found a signifi cant effect of bivalve shellfi sh on P recycling (Kaspar et al. 1985 ; Dame et al. 1991 ; Hatcher et al. 1994 ; Condon 2005 ); others have reported a preferential recycling of N over P in high - density oyster reefs (Sornin et al. 1986 ) and mussel beds (Prins and Smaal 1994 ), or a balanced release of N and P (based on measurements of single Tapes animals; Magni et al. 2000 ). A decrease in the N : P ratio of seston with increasing mussel bio-masss was documented by Prins et al. (1995) but, in general, sediment oxygenation and iron, sulfur, and calcium cycling are consid-ered to control P net effl ux from sediments more strongly than bivalve shellfi sh (Krom and Berner 1981 ; Hatcher et al. 1994 ; Nizzoli et al. 2006a ). Effects of natural and cultured populations on nitrogen cycling are much clearer, especially regarding ammonium regeneration.

Negligible e ffects of m any b ivalve f arms

Although regeneration of nutrients from the sediments to the water column is commonly reported to be stimulated by shellfi sh aquacul-ture (Doering et al. 1987 ; Hatcher et al. 1994 ; Chapelle et al. 2000 ), ecosystem - scale effects have been evaluated as negligible for many bivalve culture operations because the culture area is small in comparison with the total eco-

Table 7.5 Impact of mussel farming area (0.0525 km 2 ) fl uxes on Carteau Cove (total area, 13 km 2 ).

NH 4 + PO 4 − 3 Si(OH) 4

Nutrient input by sediments in the site outside the mussel farm (mol h − 1 ) 612.3 169.0 2012 Nutrient input by sediments under the mussel farm (mol h − 1 ) 14.0 3.0 28.7 Infl uence of mussel cultivation zone on Carteau Cove 2.3% 1.8% 1.4%

From Baudinet et al. (1990) . Average input values to the water column were estimated from fl uxes of three surveys.


Oosterschelde EcoSystem ; SMOES ). Excretion by the mussels was estimated to contribute 31 – 85% of the total phosphate fl ux from the mussel bed and, on average, only about two - thirds of the particulate organic P taken in by the mussels was recycled as phosphate. Estimated ammonium regeneration by the mussels was about 50% of the median value of the model result for total N mineralization. Ammonium excretion by the mussels accounted for 17 – 94% of the total ammonium fl ux from the shellfi sh bed (Fig. 7.3 ), supporting the authors ’ hypothesis that the mussel population played a major role in N recycling in the central part of the Oosterschelde ecosystem.

The case of shallow but fairly well - fl ushed systems with intensive bivalve cultures at

animals on the culture systems were only a small percentage of the total bivalve shellfi sh populations present. While Niquil et al. (2001) did not estimate the effects of the shellfi sh on the culture systems on nutrient supplies, their model indicated that those shellfi sh consumed only 0.24% of the planktonic gross primary production in the Takapoto Atoll lagoon.

Signifi cant e ffects of i ntensive b ivalve c ulture on n utrient i nputs, e specially in p oorly fl ushed a reas

In contrast to the above fi ndings are data for systems with bivalve cultures that are at or near exploitation carrying capacity, especially in poorly fl ushed systems. Many studies from such areas have focused on naturally occurring shellfi sh beds or individual cultured animals and used the data to make inferences about the importance of bivalve shellfi sh beds to the total system. For example, in the Seto Inland Sea, Japan, Magni et al. (2000) extrapolated data from individual incubations of Manila (short - necked) clams ( Tapes philippinarum ) and the Japan (Asian) mussel ( Musculista sen-housia ) to estimate that benthic nutrient fl uxes were 10 - fold higher from shellfi sh areas than diffusive fl uxes modeled from sediment pro-fi les. In this poorly fl ushed system, the authors also suggested that cultured shellfi sh caused elevated rates of ammonium production and oxygen consumption at the ecosystem scale.

As another example, over a 2 - year period, Prins and Smaal (1994) used a benthic ecosys-tem tunnel device to examine fl uxes of particu-late materials and dissolved substances between bivalve beds and the water column in the Oosterschelde Estuary along the coast of the Netherlands. The authors compared rates of N and P regeneration from mussels, pre-dominantly consisting of blue mussel cultures (5.3 g C m − 2 ), with estimates of the total nitro-gen (TN) mineralization (pelagic + benthic) as calculated from a model ( Simulation Model

Figure 7.3 The amount of N mineralized by mussel beds in the central part of the Oosterschelde Estuary, compared with estimates of TN mineralization (benthic + pelagic) from the Oosterschelde ecosystem model SMOES. The bar represents the 10 – 90% range of model estimates, and the line shows the median (Prins and Smaal 1994 ).

Mussels

Total

5

4

3

2

1

0

N m

iner

aliz

atio

n in

ton

day−

1

Apr June Sept Apr June Sept1988 1989


versus sinks for the N pool in Cherrystone Inlet during 2003. Clam harvest that year ( ∼ 20 × 10 6 clams; shell height ∼ 60 mm) was estimated to remove 18,000 kg N. Twice that amount ( ∼ 36,000 kg N) was removed to the atmosphere by denitrifi cation of biodeposits, whereas 900 kg N day − 1 were released to the water column by clam excretion (Luckenbach and Wang 2004 ). In contrast, during the higher - harvest year 2004 (57.3 × 10 6 clams, shell height 43 mm), Condon (2005) estimated that ∼ 11,600 – 13,000 kg N were removed through denitrifi cation of biodeposits. A pos-sible reason given for this discrepancy was that different literature values were used to calcu-late clam biodeposition rates. The total esti-mated annual removal of N by harvesting was 2360 kg N year − 1 in 2003, versus 5450 kg N year − 1 in 2004. Estimates of N recycling by clam excretion differed between the two studies (12 – 30 kg N day − 1 estimated by Condon 2005 , versus 900 kg N day − 1 given above), in part because the 2005 study, but not the earlier work, included ammonifi cation of clam biode-

or near exploitation carrying capacity is exemplifi ed by the northern quahog ( Mercenaria mercenaria ) benthic cultures in Cherrystone Inlet, on the Delmarva Peninsula near the mouth of Chesapeake Bay (6 km 2 , mean depth 1 m, salinity 14 – 23) (Condon 2005 ). For the past ∼ 25 years, juvenile clams (shell height 10 – 15 mm) have been planted in rows ( ∼ 4 m × 18 m) at densities of 550 – 1650 clams m − 2 , covered by polyethylene netting to reduce predation over a ∼ 18 – 30 - month growout period (Luckenbach and Wang 2004b ) (Fig. 7.4 ). Estimates indicated that the bivalve cultures largely control phytoplankton dynamics in Cherrystone Inlet; the clams fi l-tered ∼ 10.1 – 81.9% of the tidal creek volume per day, depending on the time of year (Condon 2005 , and references therein). Work by Condon (2005) also supported the premise that the cultured clams also signifi cantly infl u-ence nitrogen cycling in this ecosystem.

Available literature values for clam metabo-lism were used to develop a model (described below) to assess the role of clams as sources

Figure 7.4 Hard clam ( Mercenaria mercenaria ) cultures in Cherrystone Inlet, tributary to Chesapeake Bay along the eastern shore of Virginia, USA. Each rectangular net ( ∼ 4 m × 18 m) covers ∼ 50,000 clams (Luckenbach and Wang 2004b ).


et al. 2007 ). De Casabianca et al. (1997) described the lagoon as having sustained “ shellfi sh farming - dominant eutrophication, ” with abundant macroalgae Ulva rigida and Gracilaria bursa - pastoris in or near the bivalve culture areas. Mazouni et al. (1996, 1998) reported a signifi cant infl uence of oyster culture on water - column DO and dissolved N concen-trations, less so on water - column phosphate. Oxygen uptake in the culture area ranged from 0 μ mol m − 2 h − 1 (January) to 11,823 ± 377 μ mol m − 2 h − 1 (July). Ammonium and nitrate + nitrite were released from the culture areas during the summer season at 2905 ± 327 μ mol m − 2 h − 1 and 891 ± 88 μ mol m − 2 h − 1 , respectively. In that season, the nitrate + nitrite fl ux represented about 20% of the total DIN production. During winter, ammonium fl ux was negligible, whereas nitrate + nitrite was released at 177 ± 97 μ mol m − 2 h − 1 . Phosphate release was low except in 2 months, May and November ( ∼ 1700 to ∼ 2700 μ mol m − 2 h − 1 ). Mazouni et al. (1998) suggested that in this lagoon, oyster cultures (i.e., oysters and their epibiota) pro-duced 2 × 10 7 mol N year − 1 and “ play a central role in N renewal in the water column. ” Souchu et al. (2001) reported that grazing by the cul-tured oysters controlled phytoplankton biomass during all seasons except summer. Thus, for most of the year, the ammonium regenerated from the shellfi sh cultures was not used by phytoplankton for new production but, rather, was oxidized to nitrate by nitrify-ing pelagic bacteria. The authors suggested that at least a portion of this nitrate likely dif-fused into the sediments where it was denitri-fi ed to N 2 gas and effectively removed from the ecosystem.

Culture of the mussel Mytilus galloprovin-cialis and the Manila clam Tapes philippina-rum dominates the Sacca di Goro. The lagoon is 26 km 2 in area with mean depth of 1.5 m, and it receives freshwater infl ows from the Po di Volano and the Po di Goro deltaic branches (Nizzoli et al. 2005 ). As of 2004, about one - third of the lagoon area ( ∼ 8 km 2 )

posits in estimating N release by clams. Nitrifi cation - denitrifi cation and ammonifi ca-tion of clam biodeposits accounted for more of the total clam - induced N fl ux than clam excretion and harvest, especially during summer, and total N recycling by clams exceeded N removal by clams in both 2003 and 2004 (Condon 2005 ).

The modeling effort also indicated that the cultured clam population was large enough to dominate carbon and nitrogen processes in Cherrystone Inlet, and large enough to have signifi cant effects on the phytoplankton assem-blage. Clams were estimated to consume more than 50% of the gross primary production in spring and fall months when temperatures were optimal for clam growth. The large DIN fl uxes from the clam cultures were suspected to support not only phytoplankton and benthic microalgal growth but also the development of thick macroalgal mats that covered the clam nets throughout much of the growing season.

In poorly fl ushed conditions, perhaps the best studied ecosystems with intensive bivalve culture operations are the Thau Lagoon along the Mediterranean coast of southern France, and the Sacca di Goro Lagoon along the northern coast of Italy. The fi ndings from these systems are described in detail here to provide examples of “ worst - case ” effects of high - density bivalve cultures in poorly fl ushed conditions:

In the Thau Lagoon, areas of intensive Japanese oyster culture ( Crassostrea gigas ) were compared with reference sites without shellfi sh culture. The Thau Lagoon is a shallow microtidal, wind - mixed system that sustains bottom - water anoxia during calm summer conditions (Deslous - Paoli et al. 1993 ). It is eutrophic from excessive land - based nutrient inputs as well as shellfi sh culture (Mesnage et al. 2007 ). Culture operations were described to cover ∼ 20% of the lagoon with 40 oysters m − 2 area (total standing stock 22,675 tonnes, annual harvest ∼ 13,600 tonnes) (Deslous - Paoli et al. 1993, 1998 ; Souchu et al. 2001 ; Mesnage


was licensed for clam culture and 0.4 km 2 was licensed for mussel culture. The mussel farms consisted of fi ve parallel 1 - km lines of trellis separated by ∼ 50 m of water. Mussels were cultivated on ropes (1 – 1.5 m in length) sus-pended from the trellis (mean density, 5 ropes m − 2 ), and at maturity the biomass on each rope ranged from 9 to 12 kg fresh weight. Maximum annual production was ∼ 1360 – 1815 tonnes (1500 – 2000 tons), but part of the culturing was moved to areas outside of the lagoon because of mussel death from high summer temperatures coupled with episodic anoxia. Thus, mussel production within the lagoon had declined to less than 900 tonnes year − 1 (Nizzoli et al. 2005 ).

Study of oxygen, nitrogen, and phosphorus fl uxes from the mussel farm area versus a ref-erence site revealed that the mussel farm con-tributed “ intense biodeposition ” of organic matter that stimulated sediment oxygen demand and inorganic N and P regeneration rates (Nizzoli et al. 2005 ). The overall benthic fl uxes measured ( − 11.4 ± 6.5 mmol O 2 m − 2 h − 1 ; 1.59 ± 0.47 mmol NH 4 + m − 2 h − 1 and 94 ± 42 μ mol PO 4 − 3 m − 2 h − 1 ) were among the highest ever recorded from a bivalve shellfi sh aquacul-ture site (Nizzoli et al. 2005 ). The mussel rope community was described as “ an enormous sink ” for oxygen and POM, and a large source of dissolved inorganic N and P to the water column: The authors estimated that 1 m 2 of mussel farm had an oxygen demand of 46.8 mmol m − 2 h − 1 , and released inorganic N and P at 8.5 mmol m − 2 h − 1 and 0.3 mmol m − 2 h − 1 , respectively. Thus, the mussel ropes accounted for 70% to more than 90% of the overall oxygen and nutrient fl uxes. Nizzoli et al. (2005) suggested that the net effect of the mussel farm on phytoplankton might be to increase phytoplankton turnover and overall production, rather than to limit phytoplank-ton biomass through fi lter feeding.

Manila clam farms were reported to cover about one - third of the area of the Sacca di Goro Lagoon, at densities up to 2500 clams m − 2

(Bartoli et al. 2001 ; Nizzoli et al. 2007 ). Bartoli et al. (2001) described an example of more adverse ecosystem - scale effects from shellfi sh aquaculture: The clam farms were estimated to have stimulated whole - lagoon dark oxygen consumption and ammonium recycling by a factor of 1.8 and 6.5, respectively. The authors assessed the effects of dense clam cultures (up to ∼ 2300 adults m − 2 ) that covered about one - third of the area of the lagoon. Shortly follow-ing initiation of these cultures and their harvest by sediment dredging, extensive macroalgal overgrowth (the chlorophyte Ulva rigida , which is common in nutrient overenriched areas) developed along with anoxic events and mass - death of the cultured shellfi sh. The anoxic events generally occurred during summer when the massive Ulva growth (up to 800 g dry weight m − 2 ) suddenly died and decomposed, promoting high fl uxes of sulfi de from the sediments (Viaroli et al. 1996 ).

Bartoli et al. (2001) compared a reference site with few clams versus a culture area for benthic nutrient fl uxes, oxygen and carbon dioxide concentrations, and extrapolated the data to estimate ecosystem - scale effects. The data were based on fi ve sediment cores (height 40 cm, inner diameter 20 cm) collected from both the control site and the culture site during one summer season. Densities of other macro-fauna were low in the cores from both sites. Oxygen, CO 2 , NH 4 + , reactive silica, and PO 4 − 3 fl uxes were signifi cantly higher in the culture areas, attributed to clam metabolism and to the reducing conditions in the surfi cial sedi-ments (Table 7.6 ). Phytoplankton removal or chlorophyll a fl ux to the sediments was highly variable, with the highest net fl ux (630 μ g chlo-rophyll a m − 2 h − 1 ) in the core with the highest clam density.

These data were used to make interpreta-tions about the infl uence of clam aquaculture in the Sacca di Goro at the ecosystem scale. When the data were extrapolated to the lagoonal system, average whole - lagoon dark sediment oxygen demand and CO 2 production


sediments to the overlying water: During mid - May to mid - September (water temperature > 20 ° C, dark period 10 h day − 1 ), an estimated 457 tonnes of NH 4 + and 50.8 tonnes of PO 4 − 3 were regenerated to the water column in the clam culture areas. By comparison, the amount of N and P removed by clam harvest was small. The authors ’ overall interpretation was that the cultured clams had signifi cantly affected this lagoonal ecosystem through increased sediment and water column anoxia and high nutrient fl uxes to the water column, and that the premise of clam aquaculture func-tioning as a control for eutrophication is unre-alistic in the Sacca di Goro because of the high densities of clams and extensive culture area.

In later work in the same lagoon (Nizzoli et al. 2005, 2006a, 2006b, 2007 ), Nizzoli and coworkers developed N and P budgets for a control area (1600 m − 2 ) with low density of

were stimulated by a factor of 1.8 and 3.3, respectively, and nutrient release was 6.5 - fold higher for NH 4 + N and 4.6 - fold higher for PO 4 − 3 (Bartoli et al. 2001 ). The highly biode-gradable clam feces and pseudofeces could potentially promote rapid nutrient recycling — up to 4000 μ mol NH 4 + m − 2 h − 1 and 150 μ mol PO 4 − 3 m − 2 h − 1 — that would stimulate macroal-gal growth. It was estimated that at maximal production (13,600 tonnes [15,000 tons], although 18,000 – 22,700 tonnes was suggested to be more realistic) 41.7 tonnes of N and 9 tonnes of P were removed with clam harvest. Annual nutrient loads from land - based (fresh-water) inputs were estimated at 1179 tonnes of N and 36 tonnes of P. Thus, clam harvest was estimated to have removed about 4% of the N and 25% of the P in external, land - based inputs. On the other hand, the clams stimu-lated inorganic N and P regeneration from the

Table 7.6 Average environmental conditions in a reference site with few short - necked clams ( Tapes [ Ruditapes ] philippinarum ) versus a dense culture site.

Parameter Reference site Clam culture area

Dark oxygen consumption * − 2.67 ± 0.58 mmol O 2 m − 2 h − 1 − 12.49 ± 4.54 mmol O 2 m − 2 h − 1 (4.7 × higher)

CO 2 production rates * 1.22 ± 0.81 mmol CO 2 m − 2 h − 1 10.42 ± 5.61 mmol CO 2 m − 2 h − 1 (8.5 × higher)

O 2 fl ux * Estimated oxygen demand 2.78 mmol O 2 m − 2 h − 1

Estimated oxygen demand 4.72 mmol O 2 m − 2 h − 1 (1.7 × higher)

NH 4 + fl ux from sediments * 0.15 ± 0.22 mmol NH 4 + m − 2 h − 1 2.65 ± 0.22 mmol NH 4 + m − 2 h − 1 (17.7 × higher; 4.13 mmol NH 4 + m − 2 h − 1 in the core with highest clam density)

Oxidized N (NO 3 − , NO 2 − ), urea fl uxes

Highly variable; not statistically different between sites

PO 4 − 3 fl ux from sediments * Negligible 0.15 ± 0.02 mmol PO 4 − 3 m − 2 h − 1

Silicate fl ux from sediments * 0.04 ± 0.08 mmol SiO 2 m − 2 h − 1 0.37 ± 0.14 mmol SiO 2 m − 2 h − 1 (9 × higher)

Potential O 2 consumption * from resuspension of sediments (0 – 5 cm in depth), simulating harvest

27 ± 9 mmol O 2 m − 2 h − 1 56 ± 12 mmol O 2 m − 2 h − 1 (2 × higher)

Compiled from Bartoli et al. (2001) . An asterisk ( * ) indicates that the difference between the two sites was statistically signifi cant ( P < 0.001); all of these parameters except PO 4 − 3 were signifi cantly correlated with clam biomass.


ticulate P uptake was fi ve - (low - density culture) to ninefold (high - density culture) higher than at the control site, indicating that a major frac-tion of the suspended particulate matter was retained by the bivalves.

Over the entire farming cycle, the total dis-solved phosphorus ( TDP ) internal loading at the low - and high - density culture sites was two - to fourfold higher than at the control site. Whereas the sediments in the control site were a net sink for dissolved N, substantial N was released, mostly as ammonium, at the culture sites.

The contribution of the bivalves to N and P loads at the lagoon level through fi ltration, assimilation, regeneration, and burial path-ways was estimated assuming a clam market

Manila clams (30 ind. m − 2 ), a low - density culture area ( ∼ 400 m 2 ; 300 young individuals seeded m − 2 ), and a high - density culture area ( ∼ 110 m 2 ; ∼ 800 young individuals seeded m − 2 ). External freshwater nutrient loads were esti-mated for comparison with excretion and fi l-tration activity of the clams, deposition of particulate matter, and nutrient recycling including light and dark fl uxes at 1, 3, 5, and 7 months after April seeding. Relative to the control area, in the culture areas there was a signifi cant increase in the downward fl uxes of particulate nutrients coincident with an enhancement of dissolved nutrient forms (ammonium, soluble reactive phosphate) that were released (effl uxed) to the water column (Table 7.7 ). Estimated particulate N and par-

Table 7.7 Estimated infl uence of short - necked clam cultures ( Tapes philippinarum ) on nutrient cycling and removal in the Sacca di Goro Lagoon (compiled from Nizzoli et al. 2006a ).

Parameter (entire farming cycle) Control Low density High density

Entire farming cycle (mol m − 2 )

Particulate N uptake by clams from the water 1.7 9.1 16.3

Total dissolved N (TDN) fl ux (mostly as ammonium) − 0.3 1.6 6.9

Particulate P uptake 0.1 0.6 1.0

TDP effl ux 0.2 0.5 0.8

End of farming cycle

Harvested N as mollusc fl esh ∼ 0 0.4 1.8

Harvested P as mollusc fl esh ∼ 0 0.02 0.04

Fraction of biodeposited N exported as commercial product — 1.2% 6%

Fraction of biodeposited P exported as commercial product — 0.75% 3%

Fraction of biodeposited N recycled as dissolved inorganic or organic N

— ∼ 7.5% 30%

Fraction of biodeposited P recycled as dissolved inorganic or organic P

— ∼ 2% 3%

Overall (annual cycle — 5440 tonnes of clams produced in the lagoon over the study duration)

Removal of nutrients by Tapes philippinarum : 0.23 g N, 0.03 g P per individual → 124 tonnes PN, 17 tonnes PP

Recycling of nutrients by Tapes philippinarum : 0.15 g N, 0.02 g P per individual → 83 tonnes TDN, 11 tonnes TDP

Removal of nutrients by harvesting Tapes philippinarum : 15 tonnes N, 0.8 tonnes P


or only localized signifi cant adverse effects contributing to eutrophication from bivalve shellfi sh aquaculture. This analysis is based upon peer - reviewed, published data. Never-theless, merits mention that Pawlowski et al. (accepted ; see below) described some coastal waters along the Orient such as China as having sustained system - level adversed impacts from bivalve aquaculture, based on review of unpublished data from the Food and Agriculture Organization of the United Nations.

Modeling e fforts to a ssess r elationships between b ivalve a quaculture and e utrophication

Nutrient enrichment interactions with bivalve aquaculture only recently began to be assessed through modeling efforts, mostly within the past decade. Earlier production models focused on growth and development of bivalves under different environmental conditions (Bayne and Warwick 1998 ; Henderson et al. 2001 ). Models of the population dynamics of a given cultured bivalve species were developed that considered the number of individuals, growth, food availability, population renewal through seeding, marketable size, water residence time, certain ecophysiological traits, and other vari-ables (e.g., Bacher et al. 1998 ) to estimate the carrying capacity of an ecosystem for bivalve production, that is, the maximum biomass of cultured shellfi sh that a farm or waterway could sustain without a decrease in production (Dame and Prins 1998 ; Smaal et al. 2001 ; Duarte et al. 2003 ). Some of these models included spatial features of the embayment based on a hydrodynamic model, and also depicted the nitrogen or carbon cycling among phytoplankton, cultured oysters, and detritus. Some models were constructed with a two - dimensional, coupled physical - biogeochemical framework that considered more than one shellfi sh species as polycultures (Duarte et al.

size of 10 g wet weight and a crop of 5440 tonnes (6000 tons), roughly half of the total biomass produced over an annual cycle in the Sacca di Goro Lagoon from aquaculture of Tapes philippinarum (Table 7.7 ). Com-parison of the bivalve contribution to loadings delivered to the lagoon by freshwater sources indicated that the amount of particulate matter processed by the clams was within the same order of magnitude as the land - based loadings. The regenerated fraction was about 30% of the external TDN load, and about 90% of the external TDP load. Overall, Nizzoli et al. (2006a) estimated that in this poorly fl ushed lagoon, there were elevated rates of ammo-nium production, phosphorus enrichment, and oxygen consumption at the ecosystem scale.

This study indicated rapid coupling between sedimented bivalve biodeposits and benthic recycling. Nizzoli et al. (2006a) suggested that the mollusc cultures likely reduced the export of particulate matter from the lagoon to the open sea. The authors ’ overall interpretation was that in this lagoon, a signifi cant fraction of particulate N and P external loads is retained and recycled as dissolved nutrients by clam aquaculture. They suggested that the retention of particulate matter in culture areas, and the alteration of the particulate - to - dissolved nutri-ent ratio, could negatively affect water and sediment quality and stimulate the growth of nuisance macroalgae in the lagoon ecosystem. The authors acknowledged that their compari-son did not consider nutrients supplied by phytoplankton growth, or nutrients contrib-uted to the lagoon by tidal currents from the sea. They also recognized that their extrapola-tion to the ecosystem level from the small experimental control and aquaculture plots should be considered with caution.

Of the 72 ecosystems reviewed here (Table 7.3 ), only ∼ 6% or four ecosystems have sus-tained system - level adverse impacts from large, intensive bivalve culture operations (Table 7.3 ). The other 94% have sustained negligible


the effects of hard clam cultures on carbon and nitrogen cycling in Cherrystone Inlet based on data for the clam population and water quality, using published values for clam feeding and respiration rates. The model was used to eval-uate the potential infl uence of clam cultures on the particulate carbon pool through feeding and respiratory demands, and on carbon and nitrogen cycling via feeding and biodeposition, excretion, microbial processing of wastes, and clam harvest. This model indicated that Cherrystone Inlet was at or near exploitation carrying capacity for clam aquaculture.

Numeric models are becoming increasingly popular as management tools to assist in the rapid expansion of shellfi sh aquaculture worldwide by refi ning site selection, defi ning site limitations, optimizing production, and designing and implementing monitoring pro-grams (Chamberlain et al. 2006 ; Giles et al. 2009 ). Some of these models, mostly adapted from fi nfi sh operations, are being used to assess the magnitude and spatial extent of environmental effects from shellfi sh aquacul-ture. It is important to note, however, that shellfi sh aquaculture operations generally are larger and more diffuse than fi nfi sh farms, characteristics that would result in different nutrient dispersal patterns (Hartstein and Stevens 2005 ). These operations can also attenuate fl ow over substantial areas (Plew et al. 2005 ). In addition, sinking speeds of fecal and pseudofecal material from shellfi sh culture would be expected to be slower than that of fecal material from fi nfi sh operations because the shellfi sh feces are derived from phytoplankton, whereas the fi nfi sh feces consist of the remains of relatively more dense fi sh food (Cromey et al. 2002 ).

More recently, various modeling efforts have aimed to link watershed nutrient loading and ecosystem carrying capacity for shellfi sh aquaculture (e.g., Fig. 7.5 ). Luckenbach and Wang (2004a,b) described work to link a watershed - based loading model with a physi-cal transport - based water quality model to simulate primary production and predict car-

2003 ). Other efforts assessed ecosystem effects of shellfi sh culture using a carbon - based food web model that examined benthic/pelagic cou-pling by forcing a shift from pelagic fi lter - feeders to benthic consumers (Leguerrier et al. 2004 ).

Eutrophication began to be more directly considered in carrying capacity models that were developed for shellfi sh culture at local scales or sites. For example, a mussel produc-tion model (MUSMOD © — Campbell and Newell 1998 ) was created to guide seeding of bottom culture lease sites in Maine, USA, to optimal carrying capacity, and it predicted mussel production based on physical and bio-logical variables. Some carrying capacity models have considered shellfi sh growth as a function of ecosystem characteristics related to nutrient enrichment, such as DO defi cits or primary production. EMMY (Ecophysiological Model of Mytilus edulis ), an early model by Scholten and Smaal (1999) , simulated growth and reproduction of individual mussels and examined the effects of eutrophication reduc-tion scenarios on mussel growth under con-trolled experiments over a range of nutrient loads to mesocosms. The model was designed for application as a management tool to esti-mate carrying capacity.

As other examples, Inglis et al. (2000) described a carrying capacity model for mussel growth and condition that consists of three integrated submodels including (1) a hydrody-namics model that simulates effects of tides, freshwater inputs, and weather on current fl ows, fl ushing rates, and water column struc-ture; (2) an “ ecosystem model ” to simulate phytoplankton abundance, which includes water stratifi cation, light penetration/intensity, nutrient supplies and recycling within both the water column and sediments, as well as mor-tality, sedimentation, and predation of phyto-plankton; and (3) a mussel energetics mode that considers fi ltration rates, the amount of food ingested and assimilated, and the propor-tions allocated to growth and reproduction. Condon (2005) developed a model to assess


(Deslous - Paoli et al. 1993 ), which collected 5 years of data to assess interactions among oyster culture, water column/sediment nitro-gen cycling, and land - based (watershed) versus climatic infl uences on the lagoonal ecosystem. It coupled hydrodynamics from a two - dimensional model with nutrient cycling inte-grated into a box model. Simulations indicated that nitrogen cycling and oxygen defi cits were driven by meteorological forcing during wet seasons, especially precipitation events which caused land - based nutrient inputs that stimu-lated new primary production. During the dry summer season, oyster excretion/sediment release and microzooplankton excretion/mineralization produced substantial ammo-nium that stimulated “ regenerated ” primary production, so that the ecosystem remained highly productive without land - based inputs. Thus, depending on the season, both land - based inputs and shellfi sh cultures were impor-tant in the nitrogen dynamics of this poorly fl ushed lagoon. The model suggested that biodeposition from the oyster cultures and sub-sequent sediment release was a major source of N for the lagoonal ecosystem, and was linked to oxygen reduction and localized hypoxia.

Giles et al. (2009) noted that previous numerical models of biodeposition from shell-fi sh farms have overlooked biodeposit decay and most erosion features, which could sub-stantially affect estimates. They used two par-ticle tracking models, one to simulate initial dispersal of fecal pellets and the other for initial dispersal and erosion, to estimate

rying capacity for intensive hard clam aqua-culture in the Chesapeake Bay area. Their water quality model realistically simulated primary production and various water quality parameters. They also developed and tested watershed loading models that predict surface and groundwater inputs to coastal waters. Planned efforts included coupling the water quality and watershed loading models, devel-oping clam physiology and population - level submodels and a sediment deposition/resuspension submodel, and then linking all of these components to estimate exploitation car-rying capacity for clam production in selected areas such as Cherrystone Inlet. The ultimate goal is to use the coupled models to predict how land use changes will affect water quality, primary production, and shellfi sh carrying capacity in this system and, with parameter modifi cations, in other coastal waters.

Marinov et al. (2007) developed a coupled watershed and three - dimensional biogeochem-ical model for the Sacca di Goro Lagoon. It considered clam productivity with versus without macroalgal blooms, tied to nutrient enrichment and infl uences of climatic variabil-ity as dry, average, and wet years. Chapelle et al. (2000) developed an ecosystem model for the Thau Lagoon, based on nitrogen cycling and DO concentrations, toward evaluating the effects of intensive oyster aquaculture versus watershed inputs on the lagoonal ecosystem. The watershed includes substantial agriculture, industries, and urbanized areas. The model used data from the OxyThau program

Figure 7.5 Integrated model for the Sacca di Gorro Lagoon and its watershed. (Redrawn from Marinov et al. 2007 .)

Meteorologicaldata

Meteorologicaldata

Po Rivernutrient data

scenarios WatershedModel

LagoonModel

Flowsnutrients

Aquaculture: area,initial conditions,

and seeding

densities

Open sea:

flows, nutrientsUlva (on/off)


not considered. Highest biodeposition rates ( > 15 g m − 2 day − 1 ) coincided with localized changes in benthic community structure, as indicated by the Infaunal Trophic Index (Maurer et al. 1999 ) and the Marine Biotic Index (Borja et al. 2000 ).

The Farm Aquaculture Resource Manag-ement (FARM) model (Fig. 7.6 ) was designed for prospective analyses of culture location and species selection; ecological and economic optimization of culture practices; and environ-mental assessment of farm - related eutrophica-tion effects, including mitigation (Ferreira et al. 2007 ; see also Chapter 1 in this book). FARM can be used to screen various water quality effects and to examine nutrient mass balance, so it provides a valuation methodol-ogy for integrated nutrient management. The modeling framework applies a combination of physical and biogeochemical models, bivalve growth models, and screening models to assess shellfi sh production and eutrophication. It originally was parameterized for fi ve species, alone or mixed, including the Pacifi c oyster ( Crassostrea gigas ), the blue mussel ( Mytilus edulis ), the Manila clam ( Tapes philippina-rum ), the cockle ( Cerastoderma edule ), and the Chinese scallop ( Chlamys farreri ). FARM adapts the Assessment of Estuarine Trophic Status ( ASSETS ) model (Bricker et al. 2008 ) for use at local scales, considering chlorophyll a and DO. Its eutrophication simulation includes a sustainability metric for carrying capacity, allowing users to test for thresholds of low DO and to assess potential conse-quences for water quality and stock mortality.

Ferreira et al. (2007) used the FARM model to assess quantitatively the role of shellfi sh culture in controlling nutrient emissions to coastal waters. They developed a mass balance for nutrients in a 6000 - m 2 bottom - culture oyster ( Crassostrea gigas ) operation, using data from various cultivated coastal ecosys-tems and realistic oyster densities. The simula-tion was for a 45 - day period, using seed

biodeposition from a suspended mussel farm of the greenshell mussel Perna canaliculus . In sheltered areas, represented by the initial dis-persal model, biodeposit decay mostly affected fecal pellet density on the seafl oor. In high - energy areas, represented by the erosion model, decay more strongly affected the spatial extent of the detectable farm footprint. The model ’ s predicted fl uxes underestimated measured rates by about 50%.

Two other models, DEPOMOD and FARM, are briefl y described here to illustrate the utility of modeling approaches in assessing interactions between shellfi sh aquaculture and eutrophication, and also the economic benefi t of bivalve farms in mitigating land - based eutrophication (below). First, the fi nfi sh aqua-culture waste model DEPOMOD (models the deposition and biological effects of waste solids from marine salmon cage farms; Cromey et al. 2002 ) was adapted for suspended mussel aquaculture by Weise et al. ( 2009 ), considering fi eld data for species - specifi c biodeposition rates and particle settling velocities, along with several fi nfi sh model parameters. Shellfi sh - DEPOMOD was tested at three Mytilus edulis farms in Quebec, Canada, that differed in hydrologic regime. Model predictions for sedi-mentation rates were compared with data for deposition rates from sediment traps. Localized effects of sedimentation rates were reported at ∼ two to fi ve fold higher than rates in corre-sponding control sites without mussel aqua-culture. The model accurately predicted accumulation of sediments within 30 m to more than 90 m for farms in shallow versus deeper sites, respectively, except for a site in House - Harbor Lagoon. The authors attrib-uted the disparity between model predictions and observed sedimentation rates to resuspen-sion and advection of nonfarm - derived materi-als and complex hydrodynamics. The model also correctly predicted patterns of waste dis-posal at one site while underestimating biodeposition, attributed to the fact that biodeposits from biofouling communities were


shellfi sh aquaculture was contributed by Pawlowski et al. (accepted ), who developed a model that estimates shellfi sh culture inputs worldwide. Most of the data used for the model simulations were obtained from the Food and Agricultural Organization of the United Nations (2008) . Pawlowski et al. acknowledged major uncertainties in their approach because of its global scale and lack of suffi cient information on some parameters. Noting the major increase in shellfi sh aquacul-ture that is both underway and anticipated, Pawlowski et al. assessed nutrient release from

densities of 25, 100, and 500 individuals m − 3 . The model predicted that the lowest shellfi sh density would reduce chlorophyll a by 15%, while DO remained at or above ∼ 6 mg/L. The moderate shellfi sh density was predicted to reduce chlorophyll a by 45%, but DO decreased to ∼ 4 mg/L or more. While the high shellfi sh density would have effected a 92% reduction in chlorophyll a , it also was pre-dicted to cause hypoxic conditions from a DO sag to ∼ 1.8 mg/L (Ferreira et al. 2007 ).

Among the most ambitious efforts to date to assess contributions of N and P from

Figure 7.6 Conceptual scheme of the FARM model. POM, particulate organic matter; MPP, marginal physical product; VMP, value of the MPP; APP, average physical production. (Modifi ed from Ferreira et al. 2007 ).

Farm length

Width

Current Current

Chl a Chl a

POM POM

Depth 2 31 n

Sections

Shellfish

Chl a, dissolved oxygen

Eutrophicationassessment

screening model

Physics and biogeochemistry

models

Shellfish populationdynamics model

Shellfish individualgrowth models

Shellfish growth

Optimization of farm activitiesM

orta

lity e.g

., food d

eple

tion

e.g., harvestable

biomass, APP

e.g., Chl a, POM

Shellfish productionscreening model

e.g

., M

PP, V

MP

e.g

., AS

SE

TS

gra

de

Nutrie

nt tra

din

g


7.7 ). Phosphorus, another major nutrient that can stimulate phytoplankton overgrowth, has shown a more modest increase as well (Glibert and Burkholder 2006 ).

Nutrient or e cological s toichiometry

The ratio of N to P, or the nutrient stoichiom-etry, has also been greatly altered by land - based, anthropogenic nutrient additions. Nutrient stoichiometry relates changes in the relative composition of N and P in cells and tissues of aquatic organisms versus the water column. Changes in the relative proportion of N and P have promoted major alterations in metabolism, species composition, and food web structure (Sterner et al. 2002; Elser et al. 2007 ). Overall, as Howarth (2008) wrote,

The past few decades have seen a massive increase in coastal eutrophication globally, leading to widespread hypoxia and anoxia, habitat degradation, alteration of food web structure, loss of biodiversity, and increased frequency, spatial extent, and duration of harmful algal blooms. … Agricultural sources are the largest source of nitrogen pollution to many of the planet ’ s coastal marine ecosys-tems. The rate of change in nitrogen use in agriculture is incredible, and over half of the synthetic nitrogen fertilizer that has ever been produced has been used in the past 15 years. Atmospheric deposition of nitrogen from fossil fuel combustion [urban source] also contributes . . . and is the largest single source of nitrogen pollution in some regions.

Estuaries and coastal waters are now the most nutrient overenriched ecosystems in the world (Wassmann 2005 ), attributed primarily to land - based nutrient sources, and coastal human population growth and nutrient loading from land - based pollution sources are projected to increase exponentially over the next two decades (Howarth et al. 2002 ) (Fig. 7.8 ). Regions of large - scale nutrient overen-richment from land - based sources include (among many examples) the Kattagat/Baltic

marine shellfi sh culture (bivalves, crustaceans, gastropods, collectively) as an important con-tributor to dissolved and particulate nutrients in coastal marine ecosystems, especially in eastern Asia. The authors projected dramatic increases in nutrient contributions from shell-fi sh cultures by 2050. In some areas, such as coastal waters of China, marine shellfi sh aqua-culture was evaluated as already contributing signifi cantly to total nutrient inputs: As of 2006, shellfi sh aquaculture was estimated to have contributed 18% and 30% of all marine aquaculture + river exports of N and P, respec-tively. Nevertheless, the overall contribution of shellfi sh aquaculture to nutrient inputs was projected to increase from ∼ 1% of total river exports in 2006 to, at a maximum, ∼ 6% by 2050. Thus, Pawlowski et al. ’ s model indi-cated that projected as well as recent contribu-tions of shellfi sh aquaculture to global N and P loading of coastal marine ecosystems are small in comparison with global river N and P exports.

Eutrophication of c oastal w aters from l and - b ased n utrients

In comparison with the generally localized effects of bivalve aquaculture on nutrient sup-plies, the following information depicts large - scale, extensive, ubiquitous impacts of land - based nutrient pollution on coastal eco-systems: By the turn of the twenty - fi rst century, about 60% of U.S. coastal rivers and bays already were moderately to severely degraded from land - based nutrient pollution (National Research Council 2000 ). Nitrogen is the primary nutrient that limits phytoplankton growth in many estuarine and coastal ecosys-tems and, thus, is a key nutrient in eutrophica-tion (Burkholder and Glibert 2011). Global consumption of nitrogen fertilizer has dramati-cally increased over the past 70 years, and much of this increase is in the form of urea fertilizer which is an organic N form that has been linked to increased growth of various harmful algal species (Glibert et al. 2006 ) (Fig.


noxious/toxic blooms of phytoplankton and macroalgae, oxygen defi cits, loss of benefi cial submersed aquatic vegetation, and other factors (Bricker et al. 2007, 2008 ). Conditions are predicted to worsen in nearly two - thirds of these estuaries within the next decade (Bricker et al. 2008 ).

Increased phytoplankton biomass from nutrient overenrichment may be benefi cial to bivalve shellfi sh aquaculture until noxious or toxic algal species begin to be directly or indi-rectly stimulated by excessive nutrient inputs (Burkholder 2001 ; Burkholder et al. 2008 ). Altered nutrient supplies and supply ratios from land - based sources have been directly related to the loss of benefi cial algal food species, and their replacement by undesirable algae that are toxic, not readily fi ltered, and/

Sea and the eastern North Sea in northern Europe; the northern Adriatic Sea and north-western Black Sea in southern Europe; the Seto Inland Sea, Yellow Sea, and East China Sea in the Orient; and Long Island Sound, Chesapeake Bay, the Albemarle - Pamlico Estuarine System, and the northern Gulf of Mexico in the United States (Boesch 2002 ). Many estuaries, coastal embayments, and coastal lagoons in Europe (Crouzet et al. 1999 ; Conley et al. 2000 ), Japan (Suzuki 2001 ), Australia (McComb 1995 ), and the United States (Bricker et al. 2008 ) have been adversely affected by land - based nutrient pollution. A recent assessment of nutrient - related impacts in U.S. estuaries indicated that nearly two - thirds of the assessed systems are moderately to seriously degraded by land - based eutrophication, considering

Figure 7.7 (A) The change in world consumption (million metric tons of N) of total synthetic nitrogen fertilizers (solid line) and urea consumption (solid bars) since 1960. (From Glibert et al. 2006 in Biogeochemistry , with permission.) (B) Global riverine (land - based) export of nitrogen through the 1990s, and estimated through 2030. (Redrawn from Howarth et al. 2002 .)

250

200

150

100

50

0

(A)

(B)

1960

1970

1980

1990

2000

2010

2020

7.0

6.5

6.0

5.5

5.0

4.5

4.0

3.5

3.0

2.51960 1970 1980 1990 2000 2010 2020 2030

Riv

erin

e ex

port

(T

g N

yea

r−1 )

Mill

ion

tonn

es N


the northwestern United States, a strong cor-relation has been established between harmful algal blooms — including species that thrive in nutrient overenriched waters (Burkholder et al. 2008 ; Heisler et al. 2008 ) — and human population density, rather than geoducks, in the Northwest (GEOHAB 2006 ). While the comparative role of geoduck culture in con-tributing to eutrophication is not yet known, the role of other anthropogenic inputs, col-lectively considered, is clear.

The nutrient overenrichment and associated pollutants from land - based eutrophication, such as suspended sediments, microbial patho-gens, and pesticides and other toxic substances, directly or indirectly have increased oxygen defi cits, reduced or eliminated habitat for wild and cultured shellfi sh, contaminated coastal waters with fecal bacteria and microbial pathogens, depressed recruitment and survival of shellfi sh larvae and juveniles, and increased physiological stress and disease (Bricker et al. 1999 ; Mallin et al. 2000 ; Breitburg 2002 ; Wiegner et al. 2003 ; Glasoe and Christy 2004 ; Bricker et al. 2008 , and references therein). Accordingly, as an example among many such studies, Scott et al. (1996) compared 60 sta-

or not as nutritious (Ryther 1954 ; Smayda 1989 ; Cloern 2001 ).

Scientists have now reached consensus that land - based anthropogenic nutrient enrichment is an important cause of many harmful algal blooms worldwide (Heisler et al. 2008 ). Beyond depressed food quality, the toxins from some algal species can bioaccumulate in shellfi sh and decrease fecundity, promote disease and death, and render shellfi sh unsafe for human consumption (see reviews in Shumway 1990 ; Burkholder 1998 ). Shellfi sh cultures in many regions of the world must be carefully monitored for algal toxins, and this problem is apparently increasing in some areas (Shumway 1990 ; Curtis et al. 2000 ). In some areas such as along the coast of Sweden, bio-toxin accumulation is now considered the largest impediment to further expansion of commercial shellfi sh operations (Lindahl et al. 2005 ). A positive relationship between nitro-gen loading and harmful algal blooms is evident from comparison of the global distri-bution of land - based N export and the docu-mented occurrences of several major harmful algal species (Glibert and Burkholder 2006 ). Back to consideration of geoduck culture in

Figure 7.8 Phosphorus and nitrogen inputs to various types of ecosystems, showing the highest nutrient enrichment for estuaries and coastal waters (Wassmann 2005 ).

10,000

1000

100

10

1

0.110 100 1000 10,000 100,000

P in

puts

(m

mol

m–2

yea

r–1)

Freshwaterwetlands and lakes

Agroecosystems

Estuaries andcoastal waters

Forests

N inputs (mmol m–2 year–1)


1997 ). Instead, bivalve shellfi sh culture gener-ally reduces phytoplankton and other turbid-ity, thus affording more light for seagrass growth (Newell 2004 ).

Another compelling example of the fact that land - based nutrients, in most coastal waters, represent the overwhelming cause of eutrophication was given by P á ez - Osuna et al. (1998) for shrimp aquaculture (white and blue shrimp — Penaeus vannarnei and Penaeus styl-irostris , respectively) in coastal waters of Mexico. Shrimp culture is generally considered to cause substantially more environmental degradation than bivalve culture (Naylor et al. 1998 ). About 2 kg of feed are needed to produce 1 kg of shrimp in semi - intensive and intensive production systems, as much as one - third of the feed is not consumed, and pond draining during shrimp harvest releases about 90% of all of the nutrients that are produced (see references in P á ez - Osuna et al. 1998 ). The authors linked the ∼ 250 shrimp farms along the northwestern coast of Mexico to signifi -cant localized impacts. Nevertheless, the inten-sive shrimp cultures were estimated to contribute only about 1.5% of the land - based anthropogenic N and about 0.9% of the land - based anthropogenic P to the coastal waters of Mexico.

Ecological and e conomic b enefi t of b ivalve a quaculture in c ombating e utrophication

Native shellfi sh additions commonly have been considered as a means of helping to reverse cultural eutrophication effects in shallow waters. As a recent example, Cerco and Noel (2007) added an oyster module to a predictive eutrophication model of Chesapeake Bay to assess the potential utility of native oyster restoration ( Crassostrea virginica ) on DO, phytoplankton biomass as chlorophyll a , light attenuation, and submersed aquatic vegetation. The model predicted that a

tions in two tidal creeks, one of which drained a highly urbanized watershed and the other, a relatively undisturbed watershed in the south-eastern United States. Nearly 70% of the sampling sites from the tidal creek in the urbanized watershed were closed to shellfi sh harvesting because of excess fecal coliform densities, and mortality rates of juvenile and adult eastern oysters ( Crassostrea virginica ) were much higher than in a tidal creek that drained the undisturbed watershed. Monitoring of shellfi sh meats indicated that more than 50% of stations in both tidal creeks exceeded the Interstate Shellfi sh Sanitation Conference Depuration Meat Standard.

The pervasive, major impacts of land - based eutrophication overwhelm the mostly local-ized effects of bivalve aquaculture. During the past several decades, for example, catastrophic losses of seagrass meadows have been docu-mented worldwide, especially in quiet, poorly fl ushed estuaries and coastal embayments and lagoons with reduced tidal fl ushing where land - based nutrient loads are both concen-trated and frequent (see review in Burkholder et al. 2007 ). Along with inland watershed inputs of nutrients transported to marine coasts by rivers and estuaries (Caraco 1995 ; Vitousek et al. 1997 ), rapidly increasing human population density on coastlands is more than double that in inland areas (Nicholls and Small 2002 , McGranahan et al. 2007 ). Seagrass decline in favor of macroalgae or phytoplankton is a typical response, and cul-tural eutrophication from land - based sources has been invoked as a major cause of seagrass disappearance worldwide (Burkholder et al. 2007 ). The loss of seagrass meadows has destroyed habitat for wild shellfi sh and many other benefi cial fauna. By contrast, minimal seagrass loss from bivalve shellfi sh culture gen-erally has been reported (e.g., Crawford et al. 2003 ), mostly in localized areas from physical disturbance during placement and harvest (e.g., Everett et al. 1995 ), and rarely from nutrient inputs (e.g., De Casabianca et al.


(Lindahl et al. 2005 ). Integration of shellfi sh culture with some forms of agriculture for overall reduction of nutrient inputs appears to be increasingly justifi ed considering nutrient discharge regulations and increasing effl uent treatment costs (Andrew and Frank 2004 ). Beyond qualitative or localized studies, few analyses are available on the economic value of bivalve aquaculture in reducing nutrient supplies to coastal waters, but the concept is promising and the knowledge base is begin-ning to rapidly expand, as the following exam-ples illustrate.

Hart (2003) applied a dynamic linear - quadratic model to test the effectiveness of two control measures on N pollution to coastal waters along western Sweden — one upstream as agricultural abatement, and the other down-stream as mussel aquaculture. With respect to mussel aquaculture, the model considered harvest and removal of the mussels as the main N reduction measure. In the north-western fjords, mussel aquaculture was about 2000 tonnes per year, but was estimated at potentially 15,000 tonnes (Haamer 1996 , Kollberg 1999 in Hart 2003 ), which would correspond to removal of 150 tonnes of N per year or about 20% of the infl ow to these waters from land - based sources (Haamer 1996 ). The overall interpretation was that mussel cultivation could be a cost - effective measure against N pollution along the west coast of Sweden, although Hart (2003) cau-tioned that various assumptions used to model the costs of mussel cultivation needed to be further tested.

Also focusing on the Swedish west coast, Lindahl et al. (2005) modeled the potential effects of blue mussel aquaculture on N cycling within the Gullmar Fjord. It was assumed that the outfl ow water from the mussel culture area had unchanged concentrations of phosphate and nitrate, but an 18% increase in ammo-nium and a 17% increase in detrital particles (from mussel intake of N). It was also assumed that when the concentration of plankton in the water column was greater than 4 μ g chloro-

10 - fold increase in the existing oyster biomass would reduce the summer surface chlorophyll a , system - wide, by about 1 mg m − 3 , increase the summer average deep - water DO by 0.25 g m − 3 , substantially increase benefi cial submersed aquatic vegetation, and remove 30,000 kg N day − 1 through enhanced denitrifi -cation. This latter amount of N removal was estimated to be more than the nitrogen added to the bay by direct atmospheric deposition, or about 10% of the total system loading.

Similarly, most shellfi sh aquaculture is thought to have an overall positive effect on water quality, primary production, and biodi-versity except, as mentioned, for intensive culture in localized, poorly fl ushed waters (Naylor et al. 2000 ; Gibbs 2004 ; McKindsey et al. 2006 ). Thus, shellfi sh aquaculture has been considered as bioremediation tool for polluted sites, not only for reducing nutrient loads and phytoplankton blooms but also for removing toxic contaminants and reducing concentrations of microbial pathogens (Rice 2001 ; Gifford et al. 2004 ). The grazing role of shellfi sh in removing phytoplankton from the water by fi lter feeding can offset phytoplank-ton stimulation by nutrient overenrichment from land - based sources. Shellfi sh culture has also been considered as a means to reduce the primary symptoms of eutrophication such as increased chlorophyll a , and associated oxygen defi cits that are caused by high phytoplankton respiration at night and bloom senescence, death, and decomposition (Newell 2004 ). Shellfi sh removal of excess phytoplankton and other particulate matter can also increase the light available for growth of seagrasses that provide benefi cial habitat (Burkholder et al. 2007 ). Harvest of the cultured bivalves addi-tionally removes nutrients from the system (Songsangjinda et al. 2000 ), although this effect may be minor depending on the system (e.g., Bartoli et al. 2001 ). The harvested animals could be used for seafood, fodder, and agricultural fertilizers, “ thus recycling nutri-ents from sea to land ” as a cost - effective method to improve coastal water quality


was estimated potentially to accomplish removal of 25 tonnes of N through harvest of 2540 tonnes (2800 tons) of mussel biomass with N content of ∼ 1%. About two - thirds of the harvested mussels could be used for human consumption, and the remaining small or damaged mussels could be used for Agro - Aqua recycling of nutrients (Fig. 7.9 ). In a fi eld study, 4.5 – 18.1 tonnes (5 – 20 tons) of mussel tissue per hectare were applied as organic fer-tilizer to grow barley. As a second example, mussel meat was fed to laying hens and resulted in higher egg yield and improved taste. Overall, this effort has led to a test of nutrient trading at the local scale wherein the sewage treatment plant is allowed to trade N cleaning with a mussel farm (Lindahl and Kollberg 2009 ). Depending on the outcome, shellfi sh farming

phyll a L − 1 , the mussels would not be able to digest all of the food and would reject some fi ltered plankton as pseudofeces that sank as detritus. The model output indicated that net transport of dissolved and particulate N at the mouth of the fjord was reduced by 20% through mussel farming. Lindahl et al. (2005) suggested that nutrient trading systems involv-ing mussel aquaculture should be introduced to improve coastal water quality, augmenting N reduction by the sewage treatment plant in the Lysekil community. As of 2004 (most recent available data), the plant was releasing more than 36 tonnes (nearly 40 tons) of N per year to the Gullmar Fjord, with plans to reduce that number by 25 tonnes (28 tons) in accor-dance with European Union regulation 91/271/EEG. Realistic expansion of mussel farming

Figure 7.9 The Agro - Aqua recycling system of nutrients from sea to land. (Redrawn from Lindahl et al. 2005 ; see also Chapter 8 in this book.)

ORGANIC FERTILIZER ORGANIC FODDER HUMAN CONSUMPTION

NUTRIENTS PHYTOPLANKTON

MUSSEL FARM

BIODEPOSITS


tures to savings of control costs for achieving the Baltic Sea Action Plan by Helcom (2007) would range from 5 – 60% of the market price of live mussels as seafood. The large range for the estimated value of mussel cultures in com-bating eutrophication underscored the need for more empirical research on mussel growth parameters, nutrient concentrations under dif-ferent salinity and current conditions, and locations of mussel operations. The options of selling mussels, which is infl uenced by toxin and pathogen content, would also be impor-tant in determining the marginal cleaning cost of nutrient by mussel culture.

Possible increase in nutrient regeneration from mussel fi ltration was not considered in Gren et al. ’ s (2009) model. The rationale given was that although mussel farming affects bio-geochemistry and the benthic ecosystem below the longlines through biodeposits, dropped mussels, and other detritus, the negative effects are known to be localized near the farms, “ and have to be judged in relation to the overall positive effects of using mussels to improve coastal ecosystem quality ” (Gren et al. 2009 , p. 8). Gren et al. emphasized that the focus of their study was to assess the potential for mussel farming as a cost - effective environmen-tal measure in the Baltic Sea, and recom-mended further work to examine whether the “ eventual negative effects on the Baltic ecosys-tem (from mussel aquaculture) can be kept local and be acceptable. ” The authors sup-ported Lindahl and Kollbergs ’ s (2009) pro-posed use of mussel farming as a compensation measure for agricultural nutrient emissions in a trade bidding system, and suggested that the utility of mussel farming in combination with nutrient emission trading could also be extended to an international scale.

For all such endeavors, it is important that bivalve stocking densities be suffi ciently constrained to maintain aerobic conditions in the surfi cial sediments overlying anaerobic sediments, so that coupled nitrifi cation - denitrifi cation can occur (Newell 2004 ). Areas

in some areas, together with nutrient emission trading, may be applied to other areas in various countries.

The potential value of mussel farming for alleviating the effects of land - based eutrophi-cation is being explored in the Baltic Sea as well, wherein the “ replacement value of nutri-ent cleaning ” by mussel cultures has been esti-mated using a nonlinear programming model that compares costs and impacts of the mussel farms with other abatement measures such as sewage treatment plants, changes in land use and fertilizer practices, and increased cleaning by households and industries not connected to municipal sewage treatment (Gren et al. 2009 ). The recently developed cost minimization model considered 20 abatement measures that affect agriculture, industry, transport, and households in 24 basins of the Baltic Sea. Under multiple scenarios, mussel aquaculture was evaluated to be a cost - effective method to alleviate eutrophication, even when mussels from some basins of the Baltic Sea were too small for seafood harvest (Gren et al. 2009 ).

The cost - effectiveness of mussel culture in nutrient removal from the water column depended on mussel growth, sales options, assumptions about mussel farming capacity, and the nutrient load targets (Gren et al. 2009 ). Estimated marginal cleaning costs of nutrients by mussel aquaculture, calculated as the difference in minimum costs for given nutrient reduction targets with versus without mussel farms as a cleaning option, ranged from 20 – 138 million euros per year. Mussel culture had a positive value for a large range of nutrient conditions but also varied greatly, from 0.1 to 1.1 billion euros per year. Evaluation of mussel culture as a cleaning device under the Helcom Baltic Sea Action Plan (Helcom 2007 ) indicated that inclusion of mussel aquaculture could decrease the total abatement cost by ca. 5%, corresponding to a value of 0.22 euro kg − 1 live mussel under favor-able cost and growth conditions. Moreover, the value from contributions of mussel cul-


2009 ). Gren et al. (2009) suggested that bivalve aquaculture could be so harnessed similarly as wetlands have been valuated as nutrient sinks, or forests as carbon sinks.

The potential utility of shellfi sh aquaculture combined with seaweed culture for N removal and improved DO were assessed by Miller and Wands (2009) in Long Island Sound. A mecha-nistic numerical model of eutrophication pro-cesses in the Sound, the System Wide Eutrophication Model ( SWEM ), was modifi ed to include empirical data for fi ltration of par-ticulate organic nutrients by bivalves and uptake of dissolved inorganic nutrients by cul-tured marine macroalgae. Model simulations indicated that bivalve mollusc cultures com-bined with seaweed cultures could increase minimum DO by as much as 2 mg/L, to at least 3.5 mg/L. The SWEM results additionally sug-gested that bivalve culture and macroalgal har-vesting would be more effective than additional reductions in land - based N loadings, beyond the reductions already mandated by an exist-ing total maximum daily load for N, to attain DO standards and provide improved habitat for benefi cial aquatic life.

There is also strong interest in the potential for cultured shellfi sh, in polyculture with

with moderate current fl ow would continually add oxygenated water to culture areas and help to keep the surface sediments aerobic, while additionally dispersing the biodeposits across a larger bottom area and effectively diluting their oxygen demand in decomposi-tion (Haven and Morales - Alamo 1966 ; Newell 1994). Nevertheless, bivalve aquaculture holds promise for effective application in mitigating the effects of land - based nutrient pollution with economic benefi t. For example, Ferreira et al. (2007) used the FARM model to assess the role of bottom culture of oysters in nutri-ent removal over about half a year, including an integration analysis of revenue sources. The model indicated that a ∼ 0.61 hectare (1.5 acre) oyster farm would achieve a net removal of 9.7 tonnes of N per year, equivalent to the amount of N contributed from untreated wastewater discharge of more than 3000 people or the treated sewage of about 18,000 people (Fig. 7.10 ). As yet there has been no attempt to estimate the value of bivalve aqua-culture (specifi cally, mussel farms) with a replacement cost method to combat eutrophi-cation within a broader context that considers alternative abatement measures, spatial scales, and different nutrient load targets (Gren et al.

Figure 7.10 Use of the FARM model to assess mass balance of N and the nutrient emissions trading potential of bivalve aquaculture, here, a 1.5 - acre oyster farm versus sewage treatment (Modifi ed from Ferreira et al. 2007 ) .

Net N removal from a ~0.61

hectare (1.5 acre) oyster farm

would correspond to the

amount of N from untreated

wastewater discharge from

more than 3000 people, or

treated sewage of about

18,000 people

Phytoplankton removal31,000 kg C year–1

Detritus removal84,540 kg C year–1

Population equivalents3237 PEQ year–1

INCOME PARAMETERSShellfish farming: 2300 k€ year–1

2000 k€ year–1

4300 k€ year–1

Sewage treatment:

Total income:

Density of 500 oysters m–3

180-day cultivation period

3.3 kg N year–1 PEQ

ASSETS

Chl aO2

Score11 μg L−1 Chl a

Net removal9.7 tonnes year–1

Net N removal (kg year–1)

PhytoplanktonDetritusExcretionFeces

−4822–13,151

37453545

Mass balance –10,683

40% of ingested Nreturned to ecosystem

⎫⎬⎭


in a deep, three - dimensional production system for extended periods, since high light likely would be required, and since light, tempera-ture, and other parameters that vary with depth would affect macroalgal growth. The authors suggested that if the concept can be developed to be commercially viable, an obvious application would be to assist the aquaculture industry along the southern coast of Norway, and that similar systems could be used in sheltered locations of other regions that have access to deeper waters.

Conclusions

This chapter addressed two questions: How signifi cant is bivalve shellfi sh aquaculture in the eutrophication (nutrient pollution, oxygen defi cits) of coastal waters, based on present evidence? Conversely, what are the impacts of land - based nutrient pollution and association pollutants on bivalve aquaculture? In response to the fi rst question, four, or ∼ 7%, of the 62 ecosystems described in the many publica-

fi nfi sh and with macroalgae, to alleviate the nutrient pollution from fi nfi sh aquaculture (Folke and Kautsky 1989 ; Shpigel et al. 1993 ; Buschmann et al. 1996 ; Parsons et al. 2002 ). For example, Bodvin et al. (1996) developed a theoretical model linking the production of salmon, blue mussels, and macroalgae (based on small - scale culture information for the kelp, Laminaria digitata , since data were not available for mass - culture of appropriate sea-weeds in systems other than “ two - dimensional ” shallow basins) (Fig. 7.11 ). The authors acknowledged that such theoretical models are a modest fi rst step, as some of the critical assumptions require further development. For example, the mussels were assumed to consume all incoming particles in the outfl ow from salmon aquaculture. Although mussels have been shown to use particulate wastes from fi sh or shrimp culture as a food resource (Hopkins et al. 1993 ; Kwei Lin et al. 1993 ; Shpigel et al. 1993 ), they cannot be expected to be 100% effi cient in removing such wastes. It is doubtful that seaweeds (whatever the species) could be maintained with the assumed uniform growth

Figure 7.11 Theoretical model linking production of salmon, blue mussels, and macroalgae. (Based on Bodvin et al. 1996 .)

Salmon cultures

300 tonnes in 12 closed production units, each 500 m3;total water flow, 60 m3 min–1 standard high-energy dry feed

Outflow Water

Outflow Water

15 tonnes of N (87% dissolved), 2.4 tonnes of P (29% dissolved)

Blue mussel cultures

112.5 tonnes of mussels (wet weight) needed to filter 60 m3 min−1

If all particles were filtered, mussels would retain 25% of the N

25−30% N released as feces; 40–50% released as dissolved N(total 13.9 tonnes of dissolved N; including 0.9 tonnes from mussels)

Seaweed cultures

Closed units; 1000 m3; assume 4% N in dry weight; dry weight 20% of wet weight;Estimated growth rate of 10% day−1

45 tonnes of seaweed needed to take up all dissolved Nfrom salmon and mussel production


stress shellfi sh (Philippart et al. 2003 ), and are expected to interact with nutrient over - enrichment and related pollution to weaken shellfi sh and make them more prone to disease.

In summary, relative to land - based pollution sources, bivalve aquaculture has been found to contribute little to eutrophication except in some poorly fl ushed areas with high shellfi sh density. Aquaculturists should strive to main-tain cultures below ecological carrying capacity to prevent such ecosystem - level adverse effects. Within the constraints of ecosystem carrying capacity, the benefi cial effects of bivalve shell-fi sh aquaculture in effectively reducing phyto-plankton and the water - column nutrients available for blooms are beginning to be har-nessed for economic benefi t to offset nutrient overenrichment from land - based sources in coastal zones. In contrast to the generally minimal effects of bivalve aquaculture on eutrophication, major, pervasive nutrient pol-lution from many urban and agricultural sources is seriously affecting shellfi sh popula-tions and shellfi sh aquaculture in many coastal waters of the world, and these impacts are expected to increase with rapidly expanding coastal development. Considering that shellfi sh aquaculture is vital to meet the seafood demands of the rapidly increasing global human popula-tion, there is a pressing need for resource man-agers and policymak ers to increase protection of shellfi sh aqua culture operations from land - based nutrient pollution.

Literature cited

Andrew , M.L. , and Frank , L. 2004 . Integrated aquaculture system for nutrient reduction in agricultural wastewater: potential and chal-lenges . Bulletin of Fisheries Research Agency (Japan) Suppl. 1 : 143 – 152 .

Arzul , G. , Seguel , M. , and Clement , A. 2001 . Effect of marine animal excretions on differential growth of phytoplankton species. International Council for the Exploration of the Sea (ICES) . Journal of Marine Science 58 : 386 – 390 .

tions reviewed here have sustained ecosystem - level eutrophication from bivalve shellfi sh aquaculture. These impacts mostly have occurred in poorly fl ushed systems with extremely high densities cultured shellfi sh that exceeded the ecological carrying capacity. The remaining 93%, or the great majority, of the ecosystems thus far have sustained negligible or only localized eutrophication effects from bivalve culture. The four exceptions under-score the need to consider ecosystem carrying capacity rather than the carrying capacity for maximal shellfi sh production to minimize adverse effects.

The response to the second question is also clear: Land - based sources of eutrophication have seriously degraded most estuaries and coastal waters throughout the world. Their major, pervasive infl uence overwhelms the mostly localized impacts that have been docu-mented from bivalve shellfi sh aquaculture. Bivalve aquaculture is projected to increase signifi cantly during the coming decades (Shumway et al. 2003 ; Shumway and Kraeuter 2004 , Food and Agricultural Organization of the United Nations 2006 ; Pawłowski et al. accepted). Coastal human population growth, already comprising more than half of the ∼ six billion people on the Earth, is increasing in many regions and projected to continue to rapidly increase. Land - based sources of eutro-phication are expected to continue to be the clear, dominant force driving eutrophication of most estuarine and coastal marine ecosystems worldwide. The acute, obvious effects of urban and land - based agricultural nutrient pollution and associated pollutants are fi sh kills and high - biomass algal blooms, but serious, more insidious chronic impacts include long - term shifts in nutrient supplies, large areas of hypoxic and anoxic bottom habitats, loss of benefi cial submersed aquatic vegetation, reduction in shellfi sh recruitment and grazing, and increased shellfi sh physiological stress, disease, and death. Increasing temperatures from warming trends in climate change can


Asami , H. , Aida , M. , and Watanabe , K. 2005 . Accelerated sulfur cycle in coastal marine sedi-ment beneath areas of intensive shellfi sh aqua-culture . Applied and Environmental Microbiology 71 : 2925 – 2933 .

Asmus , R.M. , and Asmus , H. 1991 . Mussel beds, limiting or promoting phytoplankton . Journal of Experimental Marine Biology and Ecology 148 : 215 – 232 .

Asmus , H. , Asmus , R.M. , and Reise , K. 1990 . Exchange processes in an intertidal mussel bed: a sylt - fl ume study in the Wadden Sea . Berichte der Biologische Anstalt Helgoland 6 : 1 – 79 .

Bacher , C. , Duarte , P. , Ferreira , J.G. , H é ral , M. , and Raillard , O. 1998 . Assessment and comparison of the Marennes - Ol é ron Bay (France) and Carlingford Lough (Ireland) carrying capacity with ecosystem models . Aquatic Ecology 31 : 379 – 394 .

Ball , B. , Raine , R. , and Douglas , D. 1997 . Phytoplankton and particulate matter in Carlingford Lough, Ireland. An assessment of food availability and the impact of bivalve culture . Estuaries 20 : 430 – 440 .

Banas , N.S. , Hickey , B.M. , Newton , J.A. , and Ruesink , J.L. 2007 . Tidal exchange, bivalve grazing, and patterns of primary production in Willapa Bay, Washington, USA . Marine Ecology Progress Series 341 : 123 – 139 .

Barber , B.J. , and Blake , N.J. 1985 . Substrate catab-olism related to reproduction in the bay scallop Argopecten irradians concentricus , as deter-mined by O/N and RQ physiological indices . Marine Biology 87 : 13 – 18 .

Barranguet , C. 1997 . The role of microphytoben-thic primary production in a Mediterranean mussel culture area . Estuarine, Coastal and Shelf Science 44 : 753 – 765 .

Barranguet , C. , Alliot , E. , and Plante - Cluny , M. - R. 1994 . Benthic microphytic activity at two Mediterranean shellfi sh cultivation sites with reference to benthic fl uxes . Oceanologica Acta 17 : 211 – 221 .

Bartoli , M. , Nizzoli , D. , Viaroli , P. , Turolla , E. , Castaldelli , G. , Fano , E.A. , and Rossi , R. 2001 . Impact of Ruditapes philippinarum farming on nutrient dynamics and benthic respiration in the Sacca di Goro . Hydrobiologia 455 : 203 – 212 .

Baudinet , D. , Alliot , E. , Berland , B. , Grenz , C. , Plante - Cuny , M.R. , Plante , R. , and Salen - Picard ,

C. 1990 . Incidence of a mussel culture on bio-geochemical fl uxes at the sediment water inter-face . Hydrobiologia 207 : 187 – 196 .

Bayne , B.L. , and Hawkins , A.J.S. 1992 . Ecological and physiological aspects of herbivory in benthic suspension - feeding molluscs . In: John , D.M. , Hawkins , S.J. , and Price , J.H. (eds.), Plant - Animal Interactions in the Marine Benthos . Systematics Association, Special Volume No. 46. Clarendon Press , Oxford, UK , pp. 265 – 288 .

Bayne , B.L. , and Scullard , C. 1977 . Rates of nitro-gen excretion by species of Mytilus (Bivalvia: Mollusca) . Journal of the Marine Biological Association of the United Kingdom 57 : 355 – 369 .

Bayne , B.L. , and Warwick , R.M. (eds.). 1998 . Feeding and growth of bivalves: observations and models . Journal of Experimental Marine Biology and Ecology 219 (Special Issue): 1 – 262 .

Bayne , B.L. , Thompson , R.J. , and Widdows , J. 1976 . Physiology: I . In: Bayne , B.L. (ed.), Marine Mussels: Their Ecology and Physiology . Cambridge University Press , Cambridge, UK , pp. 121 – 206 .

Bayne , B.L. , Moore , M.N. , Widdow , J. , Livingstone , D.R. , and Salkeld , P. 1979 . Measurement of the responses of individuals to environmental stress and pollution: studies with bivalve molluscs . Philosophical Transactions of the Royal Society of London. Series B, Biological Sciences 286 : 563 – 581 .

Beadman , H.A. , Kaiser , M.J. , Galanidi , M. , Shucksmith , R. , and Willows , R.I. 2004 . Changes in species richness with stocking density of marine bivalves . Journal of Applied Ecology 41 : 464 – 475 .

Bendell - Young , L.I. 2006 . Contrasting the commu-nity structure and select geochemical character-istics of three intertidal regions in relation to shellfi sh farming . Environmental Conservation 33 : 21 – 27 .

Bodvin , T. , Indergaard , M. , Norgaard , E. , Jensen , A. , and Skaar , A. 1996 . Clean technology in aquaculture — a production without waste prod-ucts? Hydrobiologia 326/327 : 83 – 86 .

Boesch , D.F. 2002 . Challenges and oppor-tunities for science in reducing nutrient over - enrichment of coastal ecosystems . Estuaries 25 : 886 – 900 .


Borja , A. , Franco , J. , and P é rez , V. 2000 . A marine biotic index to establish the ecological quality of soft - bottom benthos within European estuarine and coastal environments . Marine Pollution Bulletin 40 : 1100 – 1114 .

Boucher , G. , and Boucher - Rodoni , R. 1988 . In situ measurement of respiratory metabolism and nitrogen fl uxes at the interface of oyster beds . Marine Ecology Progress Series 44 : 229 – 238 .

Bourget , E. , and Messier , D. 1982 . Macrobenthic density, biomass, and fauna of intertidal and subtidal sand in a Magdalen Islands lagoon, Gulf of St. Lawrence . Canadian Journal of Zoology 61 : 2509 – 2518 .

Boyd , A.J. , and Heasman , K.G. 1998 . Shellfi sh mariculture in the Benguela system: water fl ow patterns within a mussel farm in Saldanha Bay, South Africa . Journal of Shellfi sh Research 17 : 25 – 32 .

Breitburg , D.L. 2002 . Effects of hypoxia, and the balance between hypoxia and enrichment, on coastal fi shes and fi sheries . Estuaries 25 : 767 – 781 .

Bricker , S.B. , Clement , C.G. , Pirhalla , D.E. , Orland , S.P. , and Farrow , D.G.G. 1999 . National Estuarine Eutrophication Assessment: A Summary of Conditions . National Oceanic and Atmospheric Association , Silver Spring, MD .

Bricker , S.B. , Longstaff , B. , Dennison , W. , Jones , A. , Boicourt , K. , Wicks , C. , and Woerner , J. 2007 . Effects of Nutrient Enrichment in the Nation ’ s Estuaries: A Decade of Change . NOAA Coastal Ocean Program Decision Analysis Series No. 26. National Centers for Coastal Ocean Service , Silver Spring, MD .

Bricker , S.B. , Longstaff , B. , Dennison , W. , Jones , A. , Boicourt , K. , Wicks , C. , and Woerner , J. 2008 . Effects of nutrient enrichment in the nation ’ s estuaries: a decade of change . Harmful Algae 8 : 21 – 32 .

Burkholder , J.M. 1998 . Implications of harmful marine microalgae and heterotrophic dinofl agel-lates in management of sustainable marine fi sh-eries . Ecological Applications 8 : S37 – S62 .

Burkholder , J.M. 2001 . Eutrophication and oligo-trophication . In: Levin , S. (ed.), Encyclopedia of Biodiversity , Vol. 2 . Academic Press , New York , pp. 649 – 670 .

Burkholder , J.M. , Mallin , M.A. , Glasgow , H.B. , Larsen , L.M. , McIver , M.R. , Shank , G.C. ,

Deamer - Melia , N. , Briley , D.S. , Springer , J. , Touchette , B.W. , and Hannon , E.K. 1997 . Impacts to a coastal river and estuary from rupture of a large swine waste holding lagoon . Journal of Environmental Quality 26 : 1451 – 1466 .

Burkholder , J. , Eggleston , D. , Glasgow , H. , Brownie , C. , Reed , R. , Melia , G. , Kinder , C. , Janowitz , G. , Corbett , R. , Posey , M. , Alphin , T. , Toms , D. , Deamer , N. , and Springer , J. 2004 . Comparative impacts of two major hurricane seasons on the Neuse River and western Pamlico Sound . Proceedings of the National Academy of Sciences of the United States of America 101 : 9291 – 9296 .

Burkholder , J.M. , Dickey , D.A. , Kinder , C. , Reed , R.E. , Mallin , M.A. , Melia , G. , McIver , M.R. , Cahoon , L.B. , Brownie , C. , Deamer , N. , Springer , J. , Glasgow , H. , Toms , D. , and Smith , J. 2006 . Comprehensive trend analysis of nutrients and related variables in a large eutrophic estuary: a decadal study of anthropogenic and climatic infl uences . Limnology and Oceanography 51 : 463 – 487 .

Burkholder , J.M. , Tomasko , D. , and Touchette , B.W. 2007 . Seagrasses and eutrophication . Journal of Experimental Marine Biology and Ecology 350 : 46 – 72 .

Burkholder , J.M. , Glibert , P.M. , and Skelton , H.M. 2008 . Mixotrophy, a major mode of nutrition for harmful algal species in eutrophic waters . Harmful Algae 8 : 77 – 93 .

Buschmann , A.H. , L ó pez , D.A. , and Medina , A. 1996 . A review of environmental effects and alternative production strategies of marine aqua-culture in Chile . Aquacultural Engineering 15 : 397 – 421 .

Callier , M.D. , Weise , A.M. , McKindsey , C.W. , and Desrosiers , G. 2006 . Sedimentation rates in a suspended mussel farm (Great - Entry Lagoon, Canada): biodeposit production and dispersion . Marine Ecology Progress Series 322 : 129 – 141 .

Callier , M.D. , McKindsey , C.W. , and Desrosiers , G. 2007 . Multi - scale spatial variations in benthic sediment geochemistry and macrofaunal com-munities under a suspended mussel culture . Marine Ecology Progress Series 348 : 103 – 115 .

Callier , M.D. , McKindsey , C.W. , and Desrosiers , G. 2008 . Evaluation of indicators used to detect mussel farm infl uence on the benthos: two case


studies in the Magdalen Islands, Eastern Canada . Aquaculture 278 : 77 – 88 .

Campbell , D.E. , and Newell , C.R. 1998 . MUSMOD © , a production model for bottom culture of the blue mussel, Mytilus edulis L . Journal of Experimental Marine Biology and Ecology 219 : 171 – 203 .

Caraco , N. 1995 . Infl uence of human populations on P transfers to aquatic systems: a regional scale study using large rivers . In: Tiessen , H. (ed.), Phosphorus in the Global Environment . SCOPE 54. John Wiley & Sons Ltd. , New York , pp. 235 – 247 .

Castel , J. , Labourge , J.P. , Escaravage , V. , Auby , I. , and Garcia , M.E. 1989 . Infl uence of seagrass and oyster parks on the abundance and biomass patterns of meio - and macrobenthos in tidal fl ats . Estuarine, Coastal and Shelf Science 28 : 71 – 85 .

Cerco , C.F. , and Noel , M.R. 2007 . Can oyster res-toration reverse cultural eutrophication in Chesapeake Bay? Estuaries and Coasts 30 : 331 – 343 .

Chamberlain , J. , Fernandes , T.F. , Read , P. , Nickell , T.D. , and Davies , I.M. 2001 . Impacts of biode-posits from suspended mussel ( Mytilus edulis L.) culture on the surrounding surfi cial sediments . International Council for the Exploration of the Sea 58 : 411 – 416 .

Chamberlain , J. , Weise , A.M. , Grant , J. , and Dowd , M. 2006 . Modeling the effects of biodeposition from shellfi sh farms on the near fi eld benthic environment . Modeling Approaches to Assess the Potential Effects of Shellfi sh Aquaculture on the Marine Environment . Division of Fisheries and Oceans — Canadian Science Advisory Secretariat Research Document 2006/032.

Chapelle , A. , Menesguen , A. , Paoli , J. , Souchu , P. , Mazouni , N. , Vaquer , A. , and Millet , B. 2000 . Modelling nitrogen, primary production and oxygen in a Mediterranean lagoon. Impact of oysters farming and inputs from the watershed . Ecological Modeling 127 : 161 – 181 .

Chivilev , S. , and Ivanov , M. 1997 . Response of the Arctic benthic community to excessive amounts of nontoxic organic matter . Marine Pollution Bulletin 35 : 280 – 286 .

Christensen , P.B. , Glud , R.N. , Dalsgaard , T. , and Gillespie , P. 2003 . Impacts of longline mussel farming on oxygen and nitrogen dynamics and

biological communities of coastal sediments . Aquaculture 218 : 567 – 588 .

Cloern , J.E. 1982 . Does the benthos control phyto-plankton biomass in South San Francisco Bay? Marine Ecology Progress Series 9 : 191 – 202 .

Cloern , J.E. 2001 . Our evolving conceptual model of the coastal eutrophication problem . Marine Ecology Progress Series 210 : 223 – 253 .

Cobelo - Garc í a , A. , Millward , G.E. , Prego , R. , and Lukashin , V. 2006 . Metal concentrations in Kandalaksha Bay, White Sea (Russia) following the spring snowmelt . Environmental Pollution 1433 : 89 – 99 .

Cockcroft , A.C. 1990 . Nitrogen excretion by the surf zone bivalves Donax serra and D. sordidus . Marine Ecology Progress Series 60 : 57 – 65 .

Condon , E.D. 2005 . Physiological ecology of the cultured hard clam . Mercenaria mercenaria. M.S. Thesis, College of William and Mary, Gloucester Point, VA, 209pp.

Conley , D.J. , Kaas , H. , M ø hlenberg , E. , Rasmussen , B. , and Wildolf , J. 2000 . Characteristics of Danish estuaries . Estuaries 23 : 848 – 861 .

Cranford , P. , Dowd , M. , Grant , J. , Hargrave , B. , and McGladdery , S. 2003 . Ecosystem level effects of marine bivalve aquaculture . In: Fisheries and Oceans Canada (ed.), In: A Scientifi c Review of the Potential Environmental Effects of Aquaculture in Aquatic Systems , Vol. 1 . Canadian Technical Report of Fisheries and Aquatic Science, Fisheries and Oceans Canada , Burlington, ON, Canada , pp. 51 – 95 .

Cranford , P. , Anderson , R. , Archambault , P. , Balch , T. , Bates , S. , Bugden , G. , Callier , M.D. , Carver , C. , Comeau , L. , Hargrave , B. , Harrison , G. , Horne , E. , Kepay , P.E. , Li , W.K.W. , Mallet , A. , Ouellette , M. , and Strain , P. 2006 . Indicators and Thresholds for Use in Assessing Shellfi sh Aquaculture Impacts on Fish Habitat . Canadian Science Advisory Secretariat Research Document 2006/034. Department of Fish and Oceans , Ottawa, ON, Canada .

Crawford , C.M. , Macleod , C.K.A. , and Mitchell , I.M. 2003 . Effects of shellfi sh farming on the benthic environment . Aquaculture 224 : 117 – 140 .

Cromey , C.J. , Nickell , T.D. , and Black , K.D. 2002 . DEPOMOD — modeling the deposition and bio-


logical effects of waste solids from marine cage farms . Aquaculture 214 : 211 – 239 .

Crouzet , P. , Leonard , J. , Nixon , S. , Rees , Y. , Parr , W. , Laffon , L. , B ø gestrand , J. , Kristensen , P. , Lallana , C. , Izzo , G. , Bokn , T. , and Bak , J. 1999 . Nutrients in European ecosystems . Environmental Assessment Report 4. European Environmental Agency, Copenhagen, Denmar.

Curtis , K.M. , Trainer , V.L. , and Shumway , S.E. 2000 . Paralytic shellfi sh toxins in geoduck clams ( Panope abrupta ): variability, anatomical distri-bution, and comparison of two toxin detection methods . Journal of Shellfi sh Research 19 : 313 – 319 .

D ’ Amours , O. , Archambault , P. , McKindsey , C.W. , and Robichaud , L. 2008a . The infl uence of bivalve aquaculture on ecosystem productivity . World Aquaculture Magazine , September : 26 – 29 .

D ’ Amours , O. , Archambault , P. , McKindsey , C. , and Johnson , L.E. 2008b . Local enhancement of epibenthic macrofauna by aquaculture activities . Marine Ecology Progress Series 371 : 73 – 84 .

da Costa , K.G. , and Nalesso , R.C. 2006 . Effects of mussel farming on macrobenthic community structure in Southeastern Brazil . Aquaculture 258 : 655 – 663 .

Dahlb ä ck , B. , and Gunnarsson , L.A.H. 1981 . Sedimentation and sulfate reduction under a mussel culture . Marine Biology 63 : 269 – 275 .

Dame , R.F. 1996 . Ecology of Marine Bivalves: An Ecosystem Approach . CRC Press , Boca Raton, FL .

Dame , R.F. , and Dankers , N. 1988 . Uptake and release of materials by a Wadden Sea mussel bed . Journal of Experimental Marine Biology and Ecology 118 : 207 – 216 .

Dame , R. , and Libes , S. 1993 . Oyster reefs and nutrient retention in tidal creeks . Journal of Experimental Marine Biology and Ecology 171 : 251 – 258 .

Dame , R.F. , and Prins , T.C. 1998 . Bivalve carrying capacity in coastal ecosystems . Aquatic Ecology 31 : 409 – 421 .

Dame , R.F. , Wolaver , T.G. , and Libes , S.M. 1985 . The summer uptake and release of nitrogen by an intertidal oyster reef . Netherlands Journal of Sea Research 19 : 265 – 268 .

Dame , R. , Dankers , N. , Prins , T. , Jongsma , H. , and Smaal , A. 1991 . The infl uence of mussel beds on

nutrients in the western Wadden Sea and eastern Scheldt . Estuaries 14 : 130 – 138 .

Dankers , N. , and Zuidema , D.R. 1995 . The role of the mussel ( Mytilus edulis L.) and mussel culture in the Dutch Wadden Sea . Estuaries 18 : 71 – 80 .

Danovaro , R. , Gambi , C. , Luna , G.M. , and Mirto , S. 2004 . Sustainable impact of mussel farming in the Adriatic Sea (Mediterranean Sea): evi-dence from biochemical, microbial and meiofau-nal indicators . Marine Pollution Bulletin 49 : 325 – 333 .

De Casabianca , M. - L. , Laugier , T. , and Collart , D. 1997 . Impact of shellfi sh farming eutrophication on benthic macrophyte communities in the Thau lagoon, France . Aquaculture International 5 : 301 – 314 .

Dealteris , J.T. , Kilpatrick , B.D. , and Rheault , R.R. 2004 . A comparative evaluation of the habitat value of shellfi sh aquaculture gear, submerged aquatic vegetation and a non - vegetated seabed . Journal of Shellfi sh Research 23 : 867 – 874 .

Deslous - Paoli , J. - M. , Mazouni , N. , Souchu , P. , Landrein , S. , Pichot , P. , and Juge , C. 1993 . Oyster farming impact on the environment of a Mediterranean lagoon (Thau). Preliminary results of the OXYTHAU program . NATO ASI Series 33 : 519 – 520 .

Deslous - Paoli , J. - M. , Souchu , P. , Mazouni , N. , Juge , C. , and Dagault , F. 1998 . Relations milieu - ressources: impact de la conchyliculture sur un environnement lagunaire m é diterran é en (Thau) . Oceanologica Acta 21 : 831 – 843 .

Doering , P.H. 1989 . On the contribution of the benthos to pelagic production . Journal of Marine Research 47 : 371 – 383 .

Doering , P.H. , Oviatt , C.A. , and Kelly , J.R. 1986 . The effects of the fi lter - feeding clam Mercenaria mercenaria on carbon cycling in experimental marine mesocosms . Journal of Marine Research 44 : 839 – 861 .

Doering , P.H. , Kelly , J.R. , Oviatt , C.A. , and Sowers , T. 1987 . Effect of the hard clam Mercenaria mercenaria on benthic fl uxes of inorganic nutri-ents and gases . Marine Biology 94 : 377 – 383 .

Dowd , M. 2003 . Seston dynamics in a tidal inlet with shellfi sh aquaculture: a model study using tracer equations . Estuarine, Coastal Shelf Science 57 : 523 – 537 .

Duarte , P. , Meneses , R. , Hawkins , A.J.S. , Zhu , M. , Fang , J. , and Grant , J. 2003 . Mathematical mod-


elling to assess the carrying capacity for multi - species culture within coastal waters . Ecological Modelling 168 : 109 – 143 .

Dupuy , C. , Pastoureaud , A. , Ryckaert , M. , Sauriau , P.G. , and Montanie , H. 2000 . Impact of the oyster Crassostrea gigas on a microbial commu-nity in Atlantic coastal ponds near La Rochelle . Aquatic Microbial Ecology 22 : 227 – 242 .

Escaravage , V. , Garcia , M.E. , and Castel , J. 1989 . The distribution of meiofauna and its contribu-tion to detritic pathways in tidal fl ats (Arcachon Bay, France) . Scientia Marina 53 : 551 – 559 .

Everett , R. , Ruiz , G. , and Carlton , J. 1995 . Effect of oyster mariculture on submerged aquatic veg-etation: an experimental test in a Pacifi c Northwest estuary . Marine Ecology Progress Series 125 : 205 – 217 .

Fabi , G. , Manoukian , S. , and Spagnolo , A. 2009 . Impact of an open - sea suspended mussel culture on macrobenthic community (Western Adriatic Sea) . Aquaculture 289 : 54 – 63 .

Ferreira , J.G. , Hawkins , A.J.S. , and Bricker , S.B. 2007 . Management of productivity, environ-mental effects and profi tability of shellfi sh aqua-culture — the Farm Aquaculture Resource Management (FARM) model . Aquaculture 264 : 160 – 174 .

Folke , C. , and Kautsky , N. 1989 . The role of eco-systems for sustainable development of aquacul-ture . Ambio 18 : 234 – 243 .

Food and Agricultural Organization of the United Nations (FAO) . 2006 . The State of the World Fisheries and Aquaculture . FAO Fisheries Department , Rome, Italy .

Food and Agricultural Organization of the United Nations (FAO) . 2008 . Fisheries and aquaculture information and statistics service: aquaculture production 1950 – 2006 . FishStat Plus; Rome, Food and Agriculture Organization of the United Nations ( www.fao.org/fi shery/statistics/software/fi shstat/en ).

Fr é chette , M. , Booth , D.A. , Myrand , B. , and Bermard , H. 1991 . Variability and transport of organic seston near a mussel aquaculture site . ICES Marine Science Symposium 192 : 24 – 32 .

Freire , J. , Fernandez , L. , and Gonzalez - Gurriara , E. 1990 . Infl uence of mussel raft culture on the diet of Liocarcinus arcuatus (Leach) (Brachyura: Portunidae) in the Ria de Arosa (Galicia, NW Spain) . Journal of Shellfi sh Research 9 : 45 – 57 .

GEOHAB, Global Ecology and Oceanography of Harmful Algal Blooms Programme . 2006 . In: Glibert , P. (ed.), Habs in Eutrophic Systems . IOC and SCOR , Paris and Baltimore , p. 74 .

Gibbs , M.T. 2004 . Interactions between bivalve shellfi sh farms and fi shery resources . Aquaculture 240 : 367 – 396 .

Gifford , S. , Dunstan , R.H. , O ’ Connor , W. , Roberts T. , and Toia , R. 2004 . Pearl aquaculture — prof-itable environmental remediation? The Science of the Total Environment 319 : 27 – 37 .

Gilbert , F. , Souchu , P. , Bianchi , M. , and Bonin , P. 1997 . Infl uence of shellfi sh farming activities on nitrifi cation, nitrate reduction to ammonium and denitrifi cation at the water - sediment inter-face of the Thau lagoon, France . Marine Ecology Progress Series 151 : 143 – 153 .

Giles , H. , Pilditch , C.A. , and Bell , D.G. 2006 . Sedimentation from mussel ( Perna canaliculus ) culture in the Firth of Thames, New Zealand: impacts on sediment oxygen and nutrient fl uxes . Aquaculture 261 : 125 – 140 .

Giles , H. , Broekhuizen , N. , Bryan , K.R. , and Pilditch , C.A. 2009 . Modelling the dispersal of biodeposits from mussel farms: the importance of simulating biodeposit erosion and decay . Aquaculture 291 : 168 – 178 .

Glasoe , S. , and Christy , A. 2004 . Literature Review and Analysis: Coastal Urbanization and Microbial Contamination of Shellfi sh Growing Areas . Puget Sound Action Team , Olympia, WA .

Glibert , P.M. , and Burkholder , J.M. 2006 . The complex relationships between increasing fertil-ization of the Earth, coastal eutrophication, and HAB proliferation . In: Gran é li , E. , and Turner , J. (eds.), The Ecology of Harmful Algae . Springer - Verlag , New York , pp. 341 – 354 .

Glibert , P.M. , Harrison , J. , Heil , C. , and Seitzinger , S. 2006 . Escalating worldwide use of urea — a global change contributing to coastal eutrophi-cation . Biogeochemistry 77 : 441 – 463 .

Goldberg , R. , and Triplett , T. 1997 . Murky Waters: Environmental Effects of Aquaculture in the United States . Environmental Defense Fund , Washington, DC .

Gonz á lez - Gurriar á n , E. 1986 . Seasonal changes of benthic megafauna in the Ria de Muros e Noia (Galicia, North - West Spain): II. Decapod crusta-ceans (Brachyura) . Marine Biology 92 : 201 – 210 .


Gouleau , D. , Jouanneau , J.M. , Weber , O. , and Sauriau , P.G. 2000 . Short - and long - term sedi-mentation on Montportail - Brouage intertidal mudfl at, Marennes - Oleron Bay (France) . Continental Shelf Research 20 : 1513 – 1530 .

Goulletquer , P. , Heral , M. , Deslous - Paoli , J.M. , Prou , J. , Garnier , J. , Razet , D. , and Boromthanarat , W. 1989 . Ecophysiologie et bilan é nerg é tique de la palourde japonaise d ’ é levage Ruditapes philip-pinarum . Journal of Experimental Marine Biology and Ecology 132 : 85 – 108 .

Grant , J. , Hatcher , A. , Scott , D.B. , Pocklington , P. , Schafer , C.T. , and Winters , G.V. 1995 . A multi-disciplinary approach to evaluating impacts of shellfi sh aquaculture on benthic communities . Estuaries 18 : 124 – 144 .

Gren , I. - M. , Lindahl , O. , and Lindqvist , M. 2009 . Values of mussel farming for combating eutro-phication . Ecological Engineering 35 : 935 – 945 .

Grenz , C. 1989 . Quantifi cation et de la Biodeposition en Zones de Production Conchylicole Intensive en Mediterranee . Ph.D. thesis, Universite d ’ Aix - Marseille II, 144pp.

Grenz , C. , Hermin , M. , Baudinet , D. , and Daumas , R. 1990 . In situ biochemical and bacterial varia-tion of sediments enriched with mussel biode-posits . Hydrobiologia 207 : 153 – 160 .

Grenz , C. , Plante - Cuny , M.R. , Plante , R. , Alliot , E. , Baudinet , D. , and Berland , B. 1991 . Measurements of benthic nutrient fl uxes in Mediterranean shellfi sh farms: a methodological approach . Oceanologica Acta 14 : 195 – 201 .

Haamer , J. 1996 . Improving water quality in a eutrophied fjord system with mussel farming . Ambio 25 : 356 – 362 .

Hammen , C.S. 1968 . Aminotransferase activities and amino acid excretion of bivalve molluscs and brachiopods . Comparative Biochemistry and Physiology 26 : 697 – 705 .

Hammen , C.S. , Miller , H.F. , Jr. , and Geer , W.H. 1966 . Nitrogen excretion of Crassostrea virgin-ica . Comparative Biochemistry and Physiology 17 : 1199 – 1200 .

Hart , R. 2003 . Dynamic pollution control . Ecological Economics 47 : 79 – 93 .

Hartstein , N.D. , and Rowden , A.A. 2004 . Effect of biodeposits from mussel culture on macroinver-tebrate assemblages at sites of different hydro-dynamic regime . Marine Environmental Research 57 : 339 – 357 .

Hartstein , N.D. , and Stevens , C.L. 2005 . Deposition beneath long - line mussel farms . Aquacultural Engineering 33 : 192 – 213 .

Hatcher , A. , Grant , J. , and Schofi eld , B. 1994 . Effects of suspended mussel culture ( Mytilus spp.) on sedimentation, benthic respiration and sediment nutrient dynamics in a coastal bay . Marine Ecology Progress Series 115 : 219 – 235 .

Haven , D.S. , and Morales - Alamo , R. 1966 . Aspects of biodeposition by oysters and other inverte-brate fi lter feeders . Limnology and Oceanography 11 : 487 – 498 .

Heasman , K.G.M. , Pitcher , G.C. , McQuaid , C.D. , and Hecht , T. 1998 . Shellfi sh mariculture in the Benguela system: raft culture of Mytilus gallo-provincialis and the effect of rope spacing on food extraction, growth rate, production, and condition of mussels . Journal of Shellfi sh Research 17 : 33 – 39 .

Heisler , J. , Glibert , P. , Burkholder , J. , Anderson , D. , Cochlan , W. , Dennison , W. , Gobler , C. , Dortch , Q. , Heil , C. , Humphries , E. , Lewitus , A. , Magnien , R. , Marshall , H. , Stockwell , D. , and Suddleson , M. 2008 . Eutrophication and harmful algal blooms: a scientifi c consensus . Harmful Algae 8 : 3 – 13 .

Helcom . 2007 . An Approach to Set Country - Wise Nutrient Reduction Allocations to Reach Good Marine Environment of the Baltic Sea . Helcom BSAP Eutro Expo/2007. Helsinki Commission , Helsinki, Finland .

Henderson , A. , Gamito , S. , Karakassis , I. , Pederson , P. , and Smaal , A. 2001 . Use of hydrodynamic and benthic models for managing environmental impacts of marine aquaculture . Journal of Applied Ichthyology 17 : 163 – 172 .

Henriksen , K. , Rasmussen , M.B. , and Jensen , A. 1983 . Effect of bioturbation on microbial nitro-gen transformations in the sediment and fl uxes of ammonium and nitrate to the overlaying water . Environmental Biogeochemistry 35 : 193 – 205 .

H é ral , M. 1993 . Why carrying capacity models are useful tools for management of bivalve culture . In: Dam , R.F. (ed.), Bivalve Filter Feeders in Estuarine and Coastal Ecosystem Processes . Springer Verlag , Heidelberg, Germany , pp. 455 – 477 .

Herman , P.M. , and Scholten , H. 1990 . Can suspen-sion feeders stabilize estuarine ecosystems? In:


Gibson , R.N. (ed.), Trophic Relationships in the Marine Environment . Aberdeen University Press , Aberdeen, Scotland , pp. 104 – 116 .

Hilborn , R. , Armstrong , D. , Friedman , C. , Naish , K. , Orensanz , J. , Ruesink , J. , Vadopalas , B. , Feldman , K. , Valero , J. , Cheney , D. , Suhrbier , A. , Christy , A. , and Davis , J. 2004 . Comprehensive literature review and synopsis of issues relating to geoduck ( Panopea abrupta ) ecology and aquaculture production . Draft of Deliverable, January 12. Prepared for the Washington State Department of Natural Resources, Olympia, WA, 123pp.

Hopkins , J.S. , Hamilton , R.D. , II , Sandifer , P.A. , and Browdy , C.L. 1993 . The production of bivalve molluscs in intensive shrimp ponds and their effect on shrimp production and water quality . World Aquaculture 24 : 74 – 77 .

Howarth , R.W. 2008 . Coastal nitrogen pollution: a review of sources and trends globally and region-ally . Harmful Algae 8 : 14 – 20 .

Howarth , R.W. , Boyer , E.W. , Pabich , W.J. , and Galloway , J.N. 2002 . Nitrogen use in the United States from 1961 – 2000 and potential future trends . Ambio 31 : 88 – 96 .

Iglesias , J. 1981 . Spatial and temporal changes in the demersal fi sh community of the Ria de Arosa (NW Spain) . Marine Biology 65 : 199 – 208 .

Inglis , G.J. , and Gust , N. 2003 . Potential indirect effects of shellfi sh culture on the reproductive success of benthic predators . Journal of Applied Ecology 40 : 1077 – 1089 .

Inglis , G.J. , Hayden , B.J. , and Ross , A.H. 2000 . An overview of factors affecting the carrying capacity of coastal embayments for mussel culture . Report NIWA (National Institute of Water and Atmospheric Research, Ltd.), New Zealand.

Ito , S. , and Imai , T. 1955 . Ecology of oyster bed I. On the decline of productivity due to repeated cultures . Tohoku Journal of Agricultural Research 4 : 9 – 26 .

Jamieson , G.S. , Chew , L. , Gillespie , G. , Robinson , A. , Bendell - Young , L. , Heath , W. , Bravender , B. , Tompkins , A. , Nishimura , D. , and Doucette , P. 2001 . Phase 0 review of the environmental impacts of intertidal shellfi sh aquaculture in Baynes Sound . Canadian Science Advisory Secretariat Research Document 2001/125, ISSN 1480 - 4883, Canada (www document). www.dfo - mpo.gc.ca/csas/

Jaramillo , E. , Beltran , C. , and Bravo , A. 1992 . Mussel biodeposition in an estuary in southern Chile . Marine Ecology Progress Series 82 : 85 – 94 .

Jie , H. , Zhang , Z. , Zishan , Y. , and Widdows , J. 2001 . Differences in the benthic - pelagic particle fl ux (biodeposition and sediment erosion) at intertidal sites with and without clam ( Ruditapes philippinarum ) cultivation in Eastern China . Journal of Experimental Marine Biology and Ecology 261 : 245 – 261 .

Jordan , T.E. , and Valiela , I. 1982 . A nitrogen budget of the ribbed mussel, Geukensia demissa , and its signifi cance in nitrogen fl ow in a New England salt marsh . Limnology and Oceanography 27 : 75 – 90 .

Kaiser , M.J. , Spencer , B.E. , and Edwards , D.B. 1996 . Infaunal community changes as a result of commercial clam cultivation and harvesting . Aquatic Living Resources 9 : 57 – 63 .

Kaiser , M.J. , Laing , I. , Utting , S.D. , and Burnell , G.M. 1998 . Environmental impacts of bivalve mariculture . Journal of Shellfi sh Research 17 : 59 – 66 .

Kaspar , H.F. , Gillespie , P.A. , Boyer , I.C. , and MacKenzie , A.L. 1985 . Effects of mussel aqua-culture on the nitrogen cycle and benthic com-munities of Kenepuru Sound, Marlborough Sounds, New Zealand . Marine Biology 85 : 127 – 136 .

Kreeger , D.A. , and Newell , R.I.E. 2001 . Seasonal utilization of different seston carbon sources by the ribbed mussel, Geukensia demissa (Dillwyn) in a mid - Atlantic salt marsh . Journal of Experimental Marine Biology and Ecology 260 : 71 – 91 .

Krom , M.D. , and Berner , R.A. 1981 . The diagenesis of phosphorus in a nearshore marine sediment . Geochimica et Cosmochimica Acta 45 : 207 – 216 .

Kwei Lin , C. , Ruamthaveesub , P. , and Wanuchsoontom , P. 1993 . Integrated culture of the green mussel ( Perna viridis ) in waste water from an intensive shrimp pond: concept and practice . World Aquaculture 24 : 68 – 73 .

La Rosa , T. , Mirto , S. , Favaloro , E. , Savona , B. , Sara , G. , Danovaro , R. , and Mazzola , A. 2002 . Impact on the water column biogeochemistry of a Mediterranean mussel and fi sh farm . Water Research 36 : 713 – 721 .

Labarta , U. , Fernandez - Reiniz , M.J. , and Babarro , J.M.F. 1997 . Differences in physiological ener-


getics between intertidal and raft cultivated mussels Mytilus galloprovincialis . Marine Ecology Progress Series 152 : 167 – 173 .

Langton , R.W. , Haines , K.C. , and Lyon , R.E. 1977 . Ammonia - nitrogen production by the bivalve mollusc Tapes japonica and its recovery by the red seaweed Hypnea musciformis in a tropical mariculture system . Helgolander Wissenschaftliche Meeresuntersuchungen 30 : 217 – 229 .

Leguerrier , D. , Niquil , N. , Petiau , A. , and Bodoy , A. 2004 . Modeling the impact of oyster culture on a mudfl at food web in Marennes - Oleron Bay (France) . Marine Ecology Progress Series 273 : 147 – 162 .

Lindahl , O. , and Kollberg , S. 2009 . Can the EU agri - environmental aid program be extended into the coastal zone to combat eutrophication? Hydrobiologia 629 : 59 – 64 .

Lindahl , O. , Hart , R. , and Hernoth , B. 2005 . Improving marine water quality by mussel farming: a profi table solution for Swedish society . Ambio 34 : 131 – 138 .

Livingstone , D.R. , Widdows , J. , and Fieth , P. 1979 . Aspects of nitrogen metabolism of the common mussel Mytilus edulis : adaptation to abrupt and fl uctuating changes in salinity . Marine Biology 53 : 41 – 55 .

L ó pez - Jamar , E. , Iglesias , J. , and Otero , J.J. 1984 . Contribution of infauna and mussel - raft epi-fauna to demersal fi sh diets . Marine Ecology Progress Series 15 : 13 – 18 .

Lu , L. , and Grant , J. 2008 . Recolonization of inter-tidal infauna in relation to organic deposition at an oyster farm in Atlantic Canada — a fi eld experiment . Estuaries and Coasts 31 : 767 – 775 .

Lu , Y. , Blake , N.J. , and Torres , J.J. 1999 . Oxygen consumption and ammonia excretion of larvae and juveniles of the bay scallop Argopecten irra-dians c oncentricus (Say) . Journal of Shellfi sh Research 18 : 419 – 424 .

Lucas , M.I. , Newell , R.C. , Shumway , S.E. , Seiderer , L.J. , and Bally , R. 1987 . Particle clearance and yield in relation to bacterioplankton and sus-pended particulate availability in estuarine and open coast populations of the mussel Mytilus edulis . Marine Ecology Progress Series 36 : 215 – 224 .

Luckenbach , M.W. , and Harry , V. 2004 . Linking watershed loading and basin - level carrying capacity models to evaluate the effects of land

use on primary production and shellfi sh aqua-culture . Bulletin of the Fisheries Research Agency (Japan) Suppl. 1 : 123 – 132 .

Luckenbach , M.W. , and Wang , H.V. 2004 . Linking watershed loading and basin - level carrying capacity models to evaluate the effects of land use on primary production and shellfi sh aqua-culture . Bulletin of the Fisheries Research Agency Suppl. 1 : 123 – 132 .

Lum , S.C. , and Hammen , C.S. 1964 . Ammonia excretion of Lingula . Comparative Biochemistry and Physiology 12 : 185 – 190 .

Magni , P. , Montani , S. , Takada , C. , and Tsutsumi , H. 2000 . Temporal scaling and relevance of bivalve nutrient excretion on a tidal fl at of the Seto Inland Sea, Japan . Marine Ecology Progress Series 198 : 139 – 155 .

Mallet , A.O. , Carver , C.E. , and Landry , T. 2006 . Impact of suspended and off - bottom Eastern oyster culture on the benthic environment in eastern Canada . Aquaculture 255 : 362 – 373 .

Mallin , M.A. , Williams , K.E. , Esham , E.C. , and Lowe , R.P. 2000 . Effect of human development on bacteriological water quality in coastal watersheds . Ecological Applications 10 : 1047 – 1056 .

Mann , R. 1979 . Some biochemical and physiologi-cal aspects of growth and gametogenesis in Crassostrea gigas and Ostrea edulis grown at sustained elevated temperatures . Journal of the Marine Biological Association of the United Kingdom 59 : 95 – 110 .

Mann , R. , and Glomb , S. 1978 . The effect of temperature on growth and ammonia excre-tion of the Manila clam Tapes japonica . Estuarine and Coastal Marine Science 6 : 335 – 339 .

Mao , Y. , Zhou , Y. , Yang , H. , and Wang , R. 2006 . Seasonal variation in metabolism of cultured Pacifi c oyster, Crassostrea gigas , in Sanggou Bay, China . Aquaculture 253 : 322 – 333 .

Marino , J. , P é rez , A. , and Roma , G. 1982 . The mussel culture Mytilus edulis in the R í a de Arosa Northwestern Spain . Boletin Instituto Espa ñ ol de Oceanografi a 7 : 297 – 308 .

Marinov , D. , Galbiati , L. , Giordani , G. , Viaroji , P. , Norro , A. , Bencivelli , S. , and Zaldivar , J.M. 2007 . An integrated modelling approach for the management of clam farming in coastal lagoons . Aquaculture 269 : 306 – 320 .

Mariojouls , C. , and Sornin , J. - M. 1987 . Sur exploi-tation et d é t é rioration de la qualit é des terrains


conchylicoles: cons é quences sur les syst è mes d ’ exploitation — exemples en France et Japon . Norois 34 : 51 – 61 .

Martin , J.L.M. , Sornin , J. - M. , Marchand , M. , Depauw , N. , and Joyce , J. 1991 . The signifi cance of oyster biodeposition in concentrating organic matter and contaminants in the sediment . In: De Pauw , N. , and Joyce , J. (eds.), Aquaculture and the Environment . Reviews of the International Conference Aquaculture Europe ’ 91, Dublin, Ireland, June 10 – 12, 1991. Special Publication of the European Aquaculture Society, Vol. 14 . Ostend , Belgium , p. 207 (abstract).

Mattsson , J. , and Lind é n , O. 1983 . Benthic macro-fauna succession under mussels, Mytilus edulis L . (Bivalvia), cultured on hanging long - lines . Sarsia 68 : 97 – 102 .

Maurer , D. , Nguyen , H. , Robertson , G. , and Gerlinger , T. 1999 . The Infaunal Trophic Index (ITI): Its suitability for marine environmental monitoring . Ecological Applications 9 : 699 – 713 .

Mazouni , N. 2004 . Infl uence of suspended oyster cultures on nitrogen regeneration in a coastal lagoon (Thau, France) . Marine Ecology Progress Series 276 : 103 – 113 .

Mazouni , N. , Gaertner , J.C. , Deslous - Paoli , J.M. , Landrein , S. , and Geringer d ’ Oedenberg , M. 1996 . Nutrient and oxygen exchanges at the water - sediment interface in a shellfi sh farming lagoon (Thau, France) . Journal of Experimental Marine Biology and Ecology 203 : 92 – 113 .

Mazouni , N. , Gaertner , J.C. , and Deslous - Paoli , J.M. 1998 . Infl uence of oyster culture on water column characteristics in a coastal lagoon (Thau, France) . Hydrobiologia 373 – 374 : 149 – 156 .

Mazouni , N. , Gaertner , J. - C. , and Deslous - Paoli , J. - M. 2001 . Composition of biofouling commu-nities on suspended oyster cultures: an in situ study of their interactions with the water column . Marine Ecology Progress Series 214 : 93 – 102 .

McComb , A.J. (ed.). 1995 . Eutrophic Shallow Estuaries Lagoons . CRC Press , Boca Raton, FL .

McGranahan , G. , Balk , D. , and Anderson , B. 2007 . The rising tide: assessing the risks of climate change and human settlements in low elevation coastal zones . Environment and Urbanization 19 : 17 – 37 .

McKindsey , C.W. , Anderson , M.R. , Barnes , P. , Courtenay , S. , Landry , T. , and Skinner , M. 2006 .

Effects of Shellfi sh Aquaculture on Fish Habitat . Department of Fisheries and Oceans (DFO) — Canadian Science Advisory Secretariat Research Document 2006/011, Fisheries and Oceans Canada , Ottawa, ON, Canada .

Meeuwig , J.J. , Rasmussen , J.B. , and Peters , R.H. 1998 . Turbid waters and clarifying mussels: their moderation of empirical chl : nutrient relations in estuaries in Prince Edward Island, Canada . Marine Ecology Progress Series 171 : 139 – 150 .

Mesnage , V. , Ogier , S. , Bally , G. , Disnar , J.R. , Lottier , N. , Dedieu , K. , Rabouille , C. , and Copard , Y. 2007 . Nutrient dynamics at the sed-iment - water interface in a Mediterranean lagoon (Thau, France): infl uence of biodeposition by shellfi sh farming activities . Marine Environmental Research 63 : 257 – 277 .

Metzger , E. , Simonucci , C. , Viollier , E. , Sarazin , G. , Prevot , F. , and Jezequel , D. 2007 . Benthic response to shellfi sh farming in Thau lagoon: pore water signature . Estuarine, Coastal and Shelf Science 72 : 406 – 419 .

Miller , R.E. , and Wands , J.R. 2009 . Applying the System Wide Eutrophication Model (SWEM) for A Preliminary Quantitative Evaluation of Biomass Harvesting As A Nutrient Control Strategy for Long Island Sound . Hydroqual, Inc. , Mahwah, NJ .

Miron , G. , Landry , T. , Archambault , P. , and Frenette , B. 2005 . Effects of mussel culture hus-bandry practices on various benthic characteris-tics . Aquaculture 250 : 138 – 154 .

Mirto , S. , La Rosa , T. , Danavaro , R. , and Mazzola , A. 2000 . Microbial and meiofaunal response to intensive mussel - farm biodeposition in coastal sediments of the western Mediterranean . Marine Pollution Bulletin 40 : 244 – 252 .

Mojica , R. , Jr. , and Nelson , W.G. 1993 . Environmental effects of a hard clam ( Mercenaria mercenaria ) aquaculture site in the Indian River Lagoon, Florida . Aquaculture 113 : 313 – 329 .

Munroe , D. , and McKinley , R.S. 2007 . Commercial Manila clam ( Tapes philippinarum ) culture in British Columbia, Canada: the effects of preda-tor netting on intertidal sediment characteristics . Estuarine, Coastal and Shelf Science 72 : 319 – 328 .

Murdoch , R. , and Oliver , M. 1995 . Study of Chlorophyll Concentrations Within and Around


Mussel Farms: Beatrix Bay, Pelorus Sound . 1995/6 - WN, National Institute of Water and Atmospheric Research , Wellington, New Zealand .

National Research Council (NRC) . 2000 . Clean Coastal Waters — Understanding and Reducing the Effects of Nutrient Pollution . National Academy Press , Washington, DC .

Navarro , J. 1988 . The effects of salinity on the physiological ecology of Choromytilus chorus Molina, 1782. Bivalvia: Mytilidae . Journal of Experimental Marine Biology and Ecology 122 : 19 – 33 .

Navarro , J.M. , and Gonzalez , C.M. 1998 . Physiological responses of the Chilean scallop Argopecten purpuratus to decreasing salinities . Aquaculture 167 : 315 – 327 .

Navarro , E. , Iglesias , J.I.P. , Camacho , A.P. , Labarta , U. , and Beiras , R. 1991 . The physiological ener-getics of mussels ( Mytilus galloprovincialis Lmk ) from different cultivation rafts in the R í a de Arosa (Galicia, N.W. Spain) . Aquaculture 94 : 197 – 212 .

Naylor , R.L. , Goldberg , R.J. , Mooney , H. , Beveridge , M.C. , Clay , J. , Folk , C. , Kautsky , N. , Lubchenco , J. , Primavera , J. , and Williams , M. 1998 . Nature ’ s subsidies to shrimp and salmon farming . Nature 282 : 883 – 884 .

Naylor , R.L. , Goldberg , R.J. , Primavera , J.H. , Kautsky , N. , Beveridge , M.C. , Clay , J. , Folk , C. , Lubchenco , J. , Mooney , H. , and Troell , M. 2000 . Effect of aquaculture on world fi sh sup-plies . Nature 405 : 1017 – 1024 .

Newell , R.I.E. 2004 . Ecosystem infl uences of natural and cultivated populations of suspen-sion - feeding bivalve molluscs: a review . Journal of Shellfi sh Research 23 : 51 – 61 .

Newell , R.I.E. , and Jordan , S.J. 1983 . Preferential ingestion of organic material by the American oyster, Crassostrea virginica . Marine Ecology Progress Series 13 : 47 – 53 .

Newell , C.R. , and Shumway , S.E. 1993 . Grazing of natural particulates by bivalve molluscs: a spatial and temporal perspective . In: Dame , R.F. (ed.), Bivalve Filter Feeders in Estuarine and Coastal Ecosystem Processes . Springer , Berlin/Heidelberg, Germany , pp. 85 – 148 .

Newell , R.I.E. , Cornwell , J.C. , and Owens , M.S. 2002 . Infl uence of simulated bivalve biodeposi-tion and microphytobenthos on sediment nitro-

gen dynamics: a laboratory study . Limnology and Oceanography 47 : 1367 – 1379 .

Newell , R.I.E. , Fisher , T.R. , Holyoke , R.R. , and Cornwell , J.C. 2005 . Infl uence of eastern oysters on nitrogen and phosphorus regeneration in Chesapeake Bay . In: Dame , R. , and Olenin , S. (eds.), The Comparative Roles of Suspension Feeders in Ecosystems . 47 , NATO Science Series: IV — Earth and Environmental Sciences. Springer , Dordrecht, The Netherlands , pp. 93 – 120 .

Nicholls , R.J. , and Small , C. 2002 . Improved esti-mates of coastal populations and exposure to hazards released . EOS 83 : 301 – 307 .

Niquil , N. , Pouvreau , S. , Sakka , A. , Legendre , L. , Addessi , L. , Borgne , R.L. , Charpy , L. , and Delesalle , B. 2001 . Trophic web and carrying capacity in a pearl oyster farming lagoon (Takapoto, French Polynesia) . Aquatic Living Resources 14 : 165 – 174 .

Nixon , S.W. , Oviatt , C.A. , Garber , J. , and Lee , V. 1976 . Die1 metabolism and nutrient dynamics in a salt marsh embayment . Ecology 57 : 740 – 750 .

Nizzoli , D. , Welsh , D.T. , Bartoli , M. , and Viaroli , P. 2005 . Impacts of mussel ( Mytilus galloprovin-cialis ) farming on oxygen consumption and nutrient recycling in a eutrophic coastal lagoon . Hydrobiologia 550 : 183 – 198 .

Nizzoli , D. , Bartoli , M. , and Viaroli , P. 2006a . Nitrogen and phosphorous budgets during a farming cycle of the Manila clam Ruditapes philippinarum : an in situ experiment . Aquaculture 261 : 98 – 108 .

Nizzoli , D. , Welsh , D.T. , Fano , E.A. , and Viaroli , P. 2006b . Impact of clam and mussel farming on benthic metabolism and nitrogen cycling, with emphasis on nitrate reduction pathways . Marine Ecology Progress Series 315 : 151 – 165 .

Nizzoli , D. , Bartoli , M. , and Viaroli , P. 2007 . Oxygen and ammonium dynamics during a farming cycle of the bivalve Tapes philippina-rum . Hydrobiologia 587 : 25 – 36 .

Nugues , M.M. , Kaiser , M.J. , Spencer , B.E. , and Edwards , D.B. 1996 . Benthic community changes associated with intertidal oyster cultiva-tion . Aquaculture Research 27 : 913 – 924 .

Ogilvie , S.C. , Ross , A.H. , and Schiel , D.R. 2000 . Phytoplankton biomass associated with mussel farms in Beatrix Bay, New Zealand . Aquaculture 181 : 71 – 80 .


Ottmann , F. , and Sornin , J. - M. 1985 . Observations on sediment accumulation as a result of mollusc culture systems in France . In: Labish Chao , N. , and Kirby - Smith , W. (eds.), Proceedings of the International Symposium on Utilization of Coastal Ecosystems: Planning, Pollution, and Productivity , Vol. 1 . University of Rio Grande do Sul , Brazil , pp. 329 – 337 .

P á ez - Osuna , F. , Guerrero - Galv á n , S.R. , and Ruiz - Fern á ndez , A.C. 1998 . The environmental impact of shrimp aquaculture and the coastal pollution in Mexico . Marine Pollution Bulletin 36 : 65 – 75 .

Parsons , G.J. , Shumway , S.E. , Kuenstner , S. , and Gryska , A. 2002 . Polyculture of sea scallops ( Placopecten magellanicus ) suspended from salmon cages . Aquaculture International 10 : 65 – 77 .

Paterson , K.J. , Schreider , M.J. , and Zimmerman , K.D. 2003 . Anthropogenic effects on seston quality and quantity and the growth and sur-vival of Sydney rock oyster ( Saccostrea glomer-ata ) in two estuaries in NSW, Australia . Aquaculture 221 : 407 – 426 .

Paul , M.J. , and Meyer , J.L. 2001 . Streams in the urban landscape . Annual Review of Ecology and Systematics 32 : 333 – 365 .

Pawlowski , M. , Bouwman , L. , Beusen , A. , and Overbeek , C. Past and future nitrogen and phos-phorus balances and feed use in global aquacul-ture: I. Shellfi sh and aquatic plants . Reviews in Fisheries Science .

P é rez - Camacho , A. , Gonzalez , R. , and Fuentes , J. 1991 . Mussel culture in Galicia (N.W. Spain) . Aquaculture 94 : 263 – 278 .

Pfeiffer , T.J. , Lawson , T.B. , and Rusch , K.A. 1999 . Northern quahog, Mercenaria mercenaria , seed clam waste characterization study: precursor to a recirculating culture system design . Aquacultural Engineering 20 : 149 – 161 .

Philippart , C.J.M. , van Aken , H.M. , Beukema , J.J. , Bos , O.G. , Cadee , G.C. , and Dekker , R. 2003 . Climate - related changes in recruitment of the bivalve Macoma balthica . Limnology and Oceanography 48 : 2171 – 2185 .

Pietros , J.M. , and Rice , M.A. 2003 . The impacts of aquacultured oysters, Crassostrea virginica (Gmelin, 1791) on water column nitrogen and sedimentation: results of a mesocosm study . Aquaculture 220 : 407 – 422 .

Pilditch , C.A. , Grant , J. , and Bryan , K.R. 2001 . Seston supply to sea scallops ( Placopecten mag-ellanicus ) in suspended culture . Canadian Journal of Fisheries and Aquatic Science 58 : 241 – 253 .

Pitcher , G.C. , and Calder , D. 1998 . Shellfi sh culture in the Benguela system: phytoplankton and the availability of food for commercial mussel farms in Saldhana Bay, South Africa . Journal of Shellfi sh Research 17 : 15 – 24 .

Plew , D.R. , Stevens , C.L. , Spigel , R.H. , and Hartstein , N.D. 2005 . Hydrodynamic implica-tions of large offshore mussel farms . IEEE Journal of Oceanic Engineering 31 : 95 – 108 .

Prins , T.C. , and Smaal , C.S. 1994 . The role of the blue mussel Mytilus edulis in the cycling of nutri-ents in the Oosterschelde estuary (The Netherlands) . Hydrobiologia 282/283 : 413 – 429 .

Prins , T.C. , Escaravage , V. , Smaal , A.C. , and Peeters , J.C.H. 1995 . Nutrient cycling and phytoplank-ton dynamics in relation to mussel grazing in a mesocosm experiment . Ophelia 41 : 289 – 315 .

Prins , T.C. , Smaal , C.S. , and Dame , R.F. 1998 . A review of the feedbacks between bivalve grazing and ecosystem processes . Aquatic Ecology 31 : 349 – 359 .

Prosch , R.M. , and McLachlan , A. 1984 . The regeneration of surf - zone nutrients by the sand mussel, Donax serra R ö ding . Journal of Experimental Marine Biology and Ecology l80 : 221 – 233 .

Raillard , O. , and M é nesguen , A. 1994 . An ecosys-tem model for estimating the carrying capacity of a macrotidal shellfi sh system . Marine Ecology Progress Series 115 : 117 – 130 .

Reitan , K.I. , Oeie , G. , Olsen , Y. , and Reinerstsen , H. 1999 . Effect of increased primary production in a fjord on growth of blue mussels and scal-lops . Journal of Shellfi sh Research 18 : 726 . (abstract).

Rice , M.A. 2001 . Environmental impacts of shell-fi sh aquaculture: fi lter feeding to control eutro-phication . In: Tlusty , M. , Bengtson , D. , Halvorson , H.O. , Oktay , S. , Pearce , J. , and Rheault , R.B. Jr. Marine Aquaculture and the Environment: A Meeting for Stakeholders in the Northeast . Cape Cod Press , Falmouth, MA , pp. 76 – 86 .


Richard , L. 2004 . Balancing marine aquaculture inputs and extraction: combined culture of fi nfi sh and bivalve molluscs in the open ocean . Bulletin of Fisheries Research Agency (Japan) Suppl. 1 : 51 – 58 .

Romero , P. , Gonz á lez - Gurriar á n , E. , and Penas , E. 1982 . Infl uence of mussel rafts on spatial and seasonal abundance of crabs in the R í a de Arosa, North - West Spain . Marine Biology 72 : 201 – 210 .

Rosenberg , R. , and Loo , L. - O. 1983 . Energy fl ow in a Mytilus edulis culture in western Sweden . Aquaculture 35 : 151 – 161 .

Ryther , J.H. 1954 . The ecology of phytoplankton blooms in Moriches Bay and Great South Bay, Long Island, New York . Biological Bulletin 106 : 198 – 209 .

Sauriau , P.G. , Mouret , V. , and Rinc é , J. - P. 1989 . Organisation trophique de la malacofaune ben-thique non cultiv é e du bassin ostr é icole de Marennes - Ol é ron . Oceanologica Acta 12 : 193 – 204 .

Schl ü ter , L. , and Josefsen , S.B. 1994 . Annual varia-tion in condition, respiration and remineralisa-tion of Mytilus edulis L. in the Sound, Denmark . Helgolander Meeresunters 48 : 419 – 430 .

Scholten , H. , and Smaal , A.C. 1999 . The ecophysi-ological response of mussels ( Mytilus edulis ) in mesocosms to a range of inorganic nutrient loads: simulations with the model EMMY . Aquatic Ecology 33 : 83 – 100 .

Scott , G.I. , Fulton , M.H. , Strozier , E.D. , Key , P.B. , Daugomah , J.W. , Porter , D. , and Strozier , S. 1996 . The effects of urbanization on the American oyster, Crassostrea virginica (Gmelin) . Journal of Shellfi sh Research 15 : 523 – 524 .

Shaw , K.R. 1998 . Prince Edward Island benthic survey . Technical Report of Environmental Science, Vol. 4. Department of Fisheries and Environment, Montague, Prince Edward Island, Canada, 94pp. .

Shpigel , M. , Neori , A. , Popper , D.M. , and Giordin , H. 1993 . A proposed model for “ envi-ronmentally clean ” land - based culture of fi sh, bivalves and seaweeds . Aquaculture 117 : 115 – 128 .

Shumway , S.E. 1990 . A review of the effects of algal blooms on shellfi sh and aquaculture . Journal of the World Aquaculture Society 21 : 65 – 104 .

Shumway , S.E. , and Kraeuter , J.N. (eds.). 2004 . Molluscan shellfi sh research and management: charting a course for the future . Final Proceedings from the Cooperative research and Information Institute (CRII) Workshop, Charleston, SC, January 2000, 156pp .

Shumway , S.E. , Davis , C. , Downey , R. , Karney , R. , Kraeuter , J. , Parsons , J. , Rheault , R. , and Wikfors , G. 2003 . Shellfi sh aquaculture — in praise of sustainable economies and environ-ments . World Aquaculture 34 : 15 – 17 .

Sloth , N.P. , Blackburn , T.H. , Hansen , L.S. , Risgaard - Petersen , N. , and Lomstein , B.A. 1995 . Nitrogen cycling in sediments with different organic loadings . Marine Ecology Progress Series 116 : 163 – 170 .

Smaal , A. , and Prins , T.C. 1993 . The uptake of organic matter and the release of inorganic nutrients by bivalve suspension feeder beds . In: Dame , R.F. (ed.), Bivalve Filter Feeders in Estuarine and Coastal Ecosystem Processes . Springer - Verlag , Heidelberg, Germany , pp. 273 – 298 .

Smaal , A.C. , and Zurburg , W. 1997 . The uptake and release of suspended and dissolved material by oysters and mussels in Marennes - Ol é ron Bay . Aquatic Living Resources 10 : 23 – 30 .

Smaal , A.C. , Vonck , A.P.M.A. , and Bakker , M. 1997 . Seasonal variation in physiological ener-getics of Mytilus edulis and Cerastoderma edule of different size classes . Journal of the Marine Biological Association of the United Kingdom 77 : 817 – 838 .

Smaal , A. , van Stralen , M. , and Schuiling , E. 2001 . The interaction between shellfi sh culture and ecosystem processes . Canadian Journal of Fisheries and Aquatic Sciences 58 : 991 – 1002 .

Smayda , T.J. 1989 . Primary production and the global epidemic of phytoplankton blooms in the sea: a linkage? In: Cosper , E.M. , Bricelj , V.M. , and Carpenter , E.J. (eds.), Coastal and Estuarine Studies No. 35. Novel Phytoplankton Blooms . Springer - Verlag , New York , pp. 449 – 484 .

Smith , J. , and Shackley , S.E. 2004 . Effects of a com-mercial mussel Mytilus edulis lay on a sublitto-ral, soft sediment benthic community . Marine Ecology Progress Series 282 : 185 – 191 .

Snra , R.F. , and Baggaley , A. 1976 . Rate of excretion of ammonia by the hard clam Mercenaria


mercenaria and the American oyster Crassostrea virginica . Marine Biology 36 : 251 – 258 .

Songsangjinda , P. , Matsuda , O. , Yamamoto , T. , Rajendran , N. , and Maeda , H. 2000 . The role of suspended oyster culture on nitrogen cycle in Hiroshima Bay . Journal of Oceanography 56 : 223 – 231 .

Sornin , J. - M. , Feuillet , M. , H é ral , M. , and Deslous - Paoli , J. - M. 1983 . Effets des biod é p ô ts de l ’ hu î tre Crassostrea gigas (Thunberg) sur l ’ accumulation de mati è res organiques dans les parcs du bassin de Marennes - Ol é ron . Journal of Molluscan Studies Supplement 12A : 185 – 197 .

Sornin , J. - M. , Feulillet , M. , H é ral , M. , and Fardeu , J.C. 1986 . Infl uence descultures D ’ huitres Crassostrea gigas sur le cycle du phosphore en zone intertidale: role de la biodeposition . Oceanologica Acta 9 : 313 – 322 .

Sorokin , I.I. , Giovanardi , O. , Pranovi , F. , and Sorokin , P.I. 1999 . Need for restricting bivalve culture in the southern basin of the Lagoon of Venice . Hydrobiologia 400 : 141 – 148 .

Souchu , P. , Vaquer , A. , Collos , Y. , Landrein , S. , Deslous - Paoli , J.M. , and Bibent , B. 2001 . Infl uence of shellfi sh farming activities on the biogeochemical composition of the water column in Thau lagoon . Marine Ecology Progress Series 218 : 141 – 152 .

Spencer , B.E. , Kaiser , M.J. , and Edwards , D.B. 1996 . The effect of Manila clam cultivation on an intertidal benthic community: the early culti-vation phase . Aquaculture Research 27 : 261 – 276 .

Spencer , B.E. , Kaiser , M.J. , and Edwards , D.B. 1998 . Ecological effects of intertidal Manila clam cultivation: observations at the end of the cultivation phase . Journal of Applied Ecology 34 : 444 – 452 .

Srna , R. , and Baggaley , A. 1976 . Rate of excretion of ammonia by the hard clam Mercenaria mer-cenaria and the American oyster Crassostrea virginica . Marine Biology 36 : 251 – 258 .

Stenton - Dozey , J.M.E. , Jackson , L.F. , and Busby , A.J. 1999 . Impact of mussel culture on macro-benthic community structure in Saldahana Bay, South Africa . Marine Pollution Bulletin 39 : 357 – 366 .

Stenton - Dozey , J. , Probyn , T. , and Busby , A. 2001 . Impact of mussel ( Mytilus galloprovincia-lis ) raft culture on benthic macrofauna, in

situ oxygen uptake, and nutrient fl uxes in Saldanha Bay, South Africa . Canadian Journal of Fisheries and Aquatic Sciences 58 : 1021 – 1031 .

Sterner , R.W. 1986 . Herbivores ’ direct and indirect effects on algal populations . Science 231 : 605 – 607 .

Straus , K.M. , Crosson , L.M. , and Vadopalas , B. 2008 . Effects of Geoduck Aquaculture on the Environment: A Synthesis of Current Knowledge . Washington Sea Grant, University of Washington , Seattle, WA .

Strohmeier , T. , Aure , J. , Duinker , A. , Castberg , T. , Svardal , A. , and Strand , Ø . 2005 . Flow reduc-tion, seston depletion, meat content and distri-bution of diarrhetic shellfi sh toxins in a long - line blue mussel ( Mytilus edulis ) farm . Journal of Shellfi sh Research 24 : 15 – 23 .

Sundb ä ck , K. , Miles , A. , and Goransson , E. 2000 . Nitrogen fl uxes, denitrifi cation and the role of microphytobenthos in microtidal shallow - water sediments, an annual study . Marine Ecology Progress Series 200 : 59 – 76 .

Suzuki , T. 2001 . Oxygen - defi cient waters along the Japanese coast and their effects upon the estua-rine ecosystem . Journal of Environmental Quality 30 : 291 – 302 .

Tenore , K.R. , Boyer , L.G. , Call , R.M. , Corral , J. , Garcia - Fernandez , C. , Gonzalez , N. , Gonzalez - Gurriaran , E. , Hanson , R.B. , Iglesias , J. , Krom , M. , L ó pez - Jamar , E. , McClain , J. , Pamatmat , M.M. , Perez , A. , Rhoads , D.C. , de Santiago , G. , Tietjen , J. , Westrich , J. , and Windom , H.L. 1982 . Coastal upwelling in the Rias Bajas, NW Spain: contrasting the benthic regimes of the Rias de Arosa and de Muros . Journal of Marine Research 40 : 701 – 772 .

Viaroli , P. , Bartoli , M. , Bondavalli , C. , Christian , R.R. , Giordani , G. , and Naldi , M. 1996 . Macrophyte communities and their impact on benthic fl uxes of oxygen, sulphides and nutrients in shallow eutrophic environments . In: Caumette , P.J. , Castel , J. , and Herbert , R. (eds.), Developments in Hydrobiology 117. Coastal Lagoon Eutrophication and Anaerobic Processes: C.L.E.A.N. . Kluwer Academic Publishers , Dordrecht, Germany , pp. 105 – 119 .

Viaroli , P. , Bartoli , M. , Giordani , G. , Azzoni , R. , and Nizzoli , D. 2003 . Short term changes of benthic fl uxes during clam harvesting in a coastal


lagoon (Sacca Di Goro, Po River Delta) . Chemistry and Ecology 19 : 189 – 206 .

Vitousek , P.M. , Aber , J. , Howarth , R.W. , Likens , G.E. , Matson , P.A. , Schindler , D.W. , Schlesinger , W.H. , and Tilman , G.D. 1997 . Human altera-tion of the global nitrogen cycle: causes and consequences . Ecological Applications 7 : 737 – 750 .

Wassmann , P. 2005 . Cultural eutrophication: per-spectives and prospects . In: Wassmann , P. , and Olli , K. (eds.), Drainage Basin Inputs and Eutrophication: An Integrated Approach . University of Troms ø , Troms ø , Norway , pp. 224 – 234 . www.ut.ee/ ∼ olli/eutr/

Weise , A.M. , Cromey , C.J. , Callier , M.D. , Archambault , P. , Chamberlain , J. , and McKindsey , C.W. 2009 . Shellfi sh - DEPOMOD: modelling the biodeposition from suspended shellfi sh . Aquaculture 288 : 239 – 253 .

Weiss , E.T. , Carmichael , R.H. , and Valiela , I. 2002 . The effect of nitrogen loading on the growth rates of quahogs ( Mercenaria mercenaria ) and soft - shell clams ( Mya arenaria ) through changes in food supply . Aquaculture 211 : 275 – 289 .

Whiteley , J. , and Bendell - Young , L. 2007 . Ecological implications of intertidal mariculture: observed differences in bivalve community structure between farm and reference sites . Journal of Applied Ecology 44 : 495 – 505 .

Widdows , J. 1978 . Combined effects of body size, food concentration and season on the physiol-ogy of Mytilus edulis . Journal of the Marine Biological Association of the United Kingdom 58 : 109 – 124 .

Wiegner , T.N. , Seitzinger , S.P. , Breitburg , D.L. , and Sanders , J.G. 2003 . The effects of multiple stressors on the balance between autotrophic and heterotrophic processes in an estuarine system . Estuaries 26 : 352 – 364 .

Xie , Q. , and Burnell , G.M. 1995 . The effect of activity on the physiological rates of two clam species, Tapes philippinarum (Adams & Reeve) and Tapes decussatus (Linnaeus) . Biology and Environment: Proceedings of the Royal Irish Academy 95B : 217 – 223 .

Yang , H. , Wang , P. , Zhang , T. , Wang , J. , He , Y. , and Zhang , F. 1999 . Effects of reduced salinity on oxygen consumption and ammonia - N excre-tion of Chlamys farreri . Chinese Journal of Oceanology and Limnology 17 : 208 – 211 .

Yokoyama , H. 2002 . Impact of fi sh and pearl farming on the benthic environments in Gokasho Bay: evaluation from seasonal fl uctuations of the macrobenthos . Fisheries Science 68 : 258 – 268 .

Zeldis , J.R. 2005 . Magnitudes of natural and mussel farm - derived fl uxes of carbon and nitro-gen in the fi rth of Thames . CHC2005 - 048 . Environment Waikato Technical Report 2005/30, 42pp.

Zeldis , J.R. , Howard - Williams , C. , Carter , C.M. , and Schiel , D.R. 2008 . ENSO and riverine control of nutrient loading, phytoplankton biomass and mussel aquaculture yield in Pelorus Sound, NZ . Marine Ecology Progress Series 371 : 131 – 142 .

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