Nitrogen and phosphorous budgets during a farming cycle of the Manila clam Ruditapes philippinarum: An in situ experiment

Aquaculture 261 (2006) 98 – 108 www.elsevier.com/locate/aqua-online

Nitrogen and phosphorous budgets during a farming cycle of the Manila clam Ruditapes philippinarum: An in situ experiment Daniele Nizzoli ⁎, Marco Bartoli, Pierluigi Viaroli Department of Environmental Sciences, Parma University, Parco Area delle Scienze 33/A, 43100 Parma, Italy Received 4 April 2006; received in revised form 27 June 2006; accepted 27 June 2006

Abstract In the Sacca di Goro lagoon a farming cycle of the Manila clam (Ruditapes philippinarum) was simulated seeding young molluscs in an unexploited sandy spot. The experimental area (2100 m2) consisted of three sectors: a control (C), almost devoid of clams (∼ 1600 m− 2, ∼ 30 ind m− 2), a low (L) density area (400 m2, ∼ 300 ind m− 2) and a high (H) density zone (∼ 110 m− 2, ∼ 800 ind m− 2). Water chemistry, external freshwater nutrient loads, molluscs filtration rates, biomass, elemental composition and nutrient recycling were analysed. Clam filtration rates and light and dark fluxes of nutrients were measured with intact core incubations. Three replicate cores (i.d. 20 cm) were collected from C, L and H in April, one month after the seeding, June, August and October 2003. External loads were calculated multiplying dissolved and particulate nutrients concentration by freshwater flow from the main lagoon tributaries. Direct excretion, filtration activity of clams and particulate matter deposition resulted in significantly higher ammonium (NH+4 ) and soluble reactive phosphorus (SRP) effluxes to the water column at L and H. For the entire farming cycle, particulate nitrogen (PN) uptake by clams from the water column was 1.7 (C), 9.1 (L) and 16.3 (H) mol m− 2, whilst total dissolved nitrogen (TDN) fluxes were − 0.3 (C), 1.6 (L) and 6.9 (H) mol m− 2. Particulate phosphorus (PP) uptake from the water column was 0.1 (C), 0.6 (L) and 1.0 (H) mol m− 2, whilst total dissolved phosphorus (TDP) efflux was 0.2 (C), 0.5 (L) and 0.8 (H) mol m− 2. At the end of the farming cycle, harvested N as mollusc flesh was negligible for C, 0.4 mol m− 2 for L and 1.8 mol m− 2 for H. Harvested P as mollusc flesh was negligible for C, 0.02 mol m− 2 for L and 0.04 mol m− 2 for H. Farmed areas seem to have a great potential for fast coupling between sedimentation (filter feeder mediated biodeposition) and benthic recycling. At the lagoon level, mollusc farming probably attenuates the export of particulate matter to the open sea. Our results show that a minor fraction of biodeposited N (∼ 6%) and P (∼ 3%) was exported as a commercial product at the end of the farming cycle, whilst a larger fraction was incorporated in the sediments or recycled as dissolved inorganic or organic forms. © 2006 Elsevier B.V. All rights reserved. Keywords: Ruditapes philippinarum; Farming; Nitrogen; Phosphorous; Coastal lagoon

1. Introduction

⁎ Corresponding author. Tel.: +39 0521 905976; fax: +39 0521 905402. E-mail address: [email protected] (D. Nizzoli). 0044-8486/$ - see front matter © 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.aquaculture.2006.06.042

Molluscs farming is a fast growing industry, which in 2002 represented 23.5% of the total aquaculture production with an annual yield of 10.7 million tonnes. In Europe, the production of bivalve filter feeders accounts for up to the 80% of total marine aquaculture and sustains

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many local economies with an estimated annual revenue of about 900 million . Within the European Union, Italy is the main producer of the Manila clam Ruditapes philippinarum with an annual crop of about 50,000 tons (FAO, 2003). Most of the farming activities are concentrated in the highly productive coastal lagoons of the Northern Adriatic Sea (Rossi and Paesanti, 1992). Suspension-feeders are recognised to have important functional roles in aquatic ecosystems, as they affect both water column and bottom substrate (Prins et al., 1998; Newell, 2004). Nevertheless, the extent of the impact of intensive farming on trophic conditions and nutrients availability in coastal lagoons is still controversial (Newell, 2004). Several studies reported, for areas where molluscs are farmed suspended in the water column (oysters and mussels), high oxygen demand and nutrient recycling rates, as a consequence of organic enrichment to the sediment or due to the direct metabolic activity in the water column (Dahlbäck and Gunnarsson, 1981; Baudinet et al., 1990; Hatcher et al., 1994; Mazouni et al., 1996; Nizzoli et al., 2006b). Other works emphasize the role of the intense filtration and biodeposition as a tool to control seston concentrations and to sequester nutrients which would be removed from the system when the shellfish are harvested, buried in the sediment or lost through denitrification (Kaspar et al., 1985; Newell et al., 2002). To date, despite the increasing importance of clam cultivation, few works addressed its impact, in particular in the Mediterranean basin, even if there are evidences

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for high potential in nutrients and sediment metabolism alteration. Magni et al. (2000) extrapolated benthic fluxes from single clam incubations, finding that benthic fluxes induced by Ruditapes were one order of magnitude higher of the diffusive fluxes modelled from sediment profiles. Bartoli et al. (2001a) demonstrated that high densities of Ruditapes increased dark sediment effluxes of ammonium by 6.5 times and of phosphorus by 4.6 times. To our knowledge, there are no data available on a full farming cycle in order to quantify the overall effect of massive clam rearing on nitrogen and phosphorous dynamics. In this work we simulated a full (seven months) farming cycle of R. philippinarum in the Sacca di Goro lagoon (Italy), which is one of the most important sites for clam production. Physical and chemical parameters of the water column, sediment–water nutrient fluxes, clams filtration rates and elemental composition were quantified in order to evaluate the effects of Ruditapes farming on the net N and P balances. Results are discussed with respect to rearing densities, external nutrient loads, benthic–pelagic coupling and the implication of filter feeders cultivation at the whole lagoon scale. 2. Material and methods 2.1. Study site This study was performed in the Sacca di Goro lagoon (44°82′N, 12°27′E), a microtidal basin located in the

Fig. 1. Map of the Sacca di Goro lagoon indicating the position of the experimental area (V), the licensed surface exploited for Ruditapes philippinarum farming activity and the main fresh water sources (PV = Po di Volano; PG = Po di Goro; GI = Giralda Channel; BO = Bonello Channel).

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southernmost part of the Po River Delta (Northern Italy) (Fig. 1). This lagoon has been extensively studied in the frame of four European projects and a detailed description of its main features can be found in Viaroli et al. (2006). The Sacca di Goro has been exploited for R. philippinarum farming since 1985, with a production peak of about 15000 tons year− 1 (1990) and sudden declines (down to 8000 tons year− 1, 1996–1998) due to macroalgal blooms and associated dystrophic events (Viaroli et al., 2006). At present, between 8 to 10 km2 (approximately 1/3 of the total lagoon surface) are licensed for clam farming, with the most exploited areas located in the central part of the lagoon in front of the main connection with the sea. Clams are generally seeded at densities between 500 to 1000 individuals m− 2 (Castaldelli et al., 2003) but their distribution is not homogeneous so that densities up to 1000–2000 ind m− 2 are not unusual. The experimental site was located in a sandy area along the inner sand barrier, close to the sea mouth (Fig. 1). This zone is one of the most heavily exploited areas of the Sacca di Goro because of the optimal sedimentary and hydrodynamic characteristics. Sand was devoid of clams due to recent deposition of material collected from digging operations performed in order to maintain water circulation within the basin and to reclaim new areas for Ruditapes farming. Surface sediments were thus well oxidised and less then 2% (DW) in carbon, and 0.05% (DW) in nitrogen and phosphorus content (Nizzoli et al., in press). 2.2. Experimental set-up The experimental set-up has been previously described in Nizzoli et al. (in press). In brief we simulated a 7-month farming cycle in a field with a total surface of 2100 m2. The field was divided into three zones seeded, on 13 March 2003, at different clam densities. Sector H (110 m2) was seeded with 1500 ind m− 2, sector L (400 m2) with 500 ind m− 2 and sector C (1600 m2) was not seeded and used as a control. Samplings were carried out on 16 April, 18 June, 8 August and 15 October. At each site, on each sampling date, three replicate cores for flux determinations were randomly collected using transparent plexiglass tubes (40 cm height; i.d. 20 cm). Additionally, at the sampling site and from the main fresh water tributaries (Po di Volano, Giralda and Romanina channels and the Po di Goro river) (Fig. 1), 5 l of water were collected for determining dissolved inorganic (NH4+, NO2−, NO3−) and organic (DON) nitrogen, soluble reactive (SRP) and dissolved organic (DOP) phosphorus, particulate nitrogen (PN) and phosphorous (PP).

2.3. Fluxes determination Dissolved inorganic and organic nitrogen and phosphorous fluxes were determined both in the light and in the dark by means of incubation of intact sediment cores. After sampling, cores were transported, within 1 h, to the Goro hatchery laboratory (CRIM: Centre for Molluscs Reproduction) and submersed in a 300 l reservoir containing lagoon water. The tank was placed outside, under natural light and temperature; the cores were left to stabilise for about 12 h before the beginning of the experiments. During the 12 h the water inside each core was continuously mixed by a small aquarium pump whilst the reservoir was continuously renewed with lagoon water at a rate of 100 l h− 1 to maintain a constant supply of particulate matter for clams and the other components of the benthic community. Incubations started after lowering the level of the water in the tank just below the top of the tubes and sealing the cores with floating Plexiglas lids. Samples for solutes determinations were withdrawn from each core at the beginning and at the end of the incubations. Solute fluxes across the sediment–water interface were calculated according to Nizzoli et al. (in press). Net daily water–sediment fluxes were calculated from the sum of light and dark exchange rates multiplied for the light and dark periods. Net exchanges across the interface over the entire farming cycle (180 days) have been then calculated assuming constant daily rates for the time lags between sampling periods. After the incubations the sediment in each core was sieved through a 500 μm mesh net and the recovered clams were incubated in the dark for the measurement of filtration rates. Incubations were performed following the same experimental set-up previously described. At the beginning and at the end of the incubations 500 ml of water were collected from each core and immediately filtered using pre-combusted glass fibre filters (GF/F) for the determination of total suspended particulate matter. Weight specific clearance rates were calculated according to Nakamura (2001). Fluxes of particulate nitrogen and phosphorus were calculated multiplying filtration rates for ambient nitrogen and phosphorous concentrations. At the end of the incubation, clam density and biomass within each core were determined. Clams were weighed individually for total fresh weight. After 24 h at 70 °C, dry flesh and shell weight were also determined. The dried material was analysed for total carbon, nitrogen and phosphorus content. 2.4. Methods for water and clam analysis Ammonium was determined with the blue indophenol method according to Bower and Holm-Hansen

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(1980). Nitrate was determined after cadmium reduction as nitrite via diazotation (APHA, 1975). Phosphate was determined spectrophotometrically according to Valderrama (1977). Dissolved organic nitrogen and phosphorus were determined as phosphate and nitrate respectively, following persulfate oxidation (Valderrama, 1981). Particulate phosphorus and nitrogen were determined as phosphate and nitrate respectively, following persulfate oxidation and extraction of the filters (Valderrama, 1981). Total carbon and nitrogen content in the clam flesh and shell was determined using a CHNS-O EA 1108 Elemental Analyzer (Carlo Erba). Total phosphorous in clam flesh and shell was determined following the method of Aspila et al. (1976). 2.5. Statistical analysis Data were analysed with a two way ANOVA, with a posteriori comparison of the means performed using a post hoc Tukey test (Sokal and Rolf, 1995).

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Table 2 Daily loads from freshwater tributaries carried during the investigated period to the Sacca di Goro lagoon SRP DOP TDP NH+4 NO−3 April June August October

1.4 3.6 2.9 1.2

0.1 0.0 0.1 0.2

1.6 3.6 3.0 1.4

DON TDN PN

26.1 217.6 2.1 23.9 38.8 49.4 19.6 42.0 48.8 14.7 49.6 5.2

PP

243.8 38.5 2.5 117.1 142.0 9.0 114.1 61.3 3.9 65.8 12.7 0.8

Units are 103 mol d− 1.

and declined in spring and autumn (∼4 and ∼0.2 μM respectively). In the considered period freshwater input was on average 1300 ⁎ 103 m3 d− 1 with a summer peak up to 1500 ⁎ 103 m3 d− 1. TDN load (Table 2) was mainly composed of NO 3− and decreased from April (243 ⁎ 103 mol d− 1) to October (65 ⁎ 103 mol d− 1). PN external load peaked in summer with 142 ⁎ 103 mol d− 1 in June. TDP and PP loads had a summer maximum (∼ 3.0 ⁎ 103 and from ∼ 3.9 to 9 ⁎ 103 mol d− 1 respectively); the dissolved pool was mainly composed of SRP.

3. Results 3.2. Clams biomass and elemental composition 3.1. Water characteristics and external nutrient loads Water temperatures varied from 15 °C (April) to 27 °C (August). Total dissolved nitrogen concentrations (TDN) peaked in April (118 μM) and decreased in the following months. In spring NO3− was the dominant dissolved form (80% of TDN) whereas in summer NH4+ and dissolved organic nitrogen (DON) became quantitatively more important (Table 1). Total dissolved phosphorus concentrations (TDP) exhibited an opposite trend with summer maximum (2.5–3.4 μM) and spring and fall minima mostly driven by SRP evolution (Table 1). Particulate nitrogen (PN) and phosphorous (PP) concentrations peaked in summer (∼10 and ∼0.6 μM respectively) Table 1 Water temperature (t) and concentrations of particulate nitrogen (PN), particulate phosphorous (PP), seston N:P ratio, ammonium (NH+4), nitrate (NO−3 ), dissolved organic nitrogen (DON), reactive soluble phosphorous (SRP) and dissolved organic phosphorous (DOP) measured in the water column of the experimental area during the four samplings t

PN

°C μM April June August October

PP

N:P

μM

15 3.72 0.22 16.90 25 10.82 0.61 17.70 27 9.69 0.69 14.04 17 4.26 0.26 16.40

NH+4

NO−3

DON SRP DOP

μM

μM

μM

μM

μM

11.48 96.64 10.70 0.70 0.30 16.63 6.03 32.50 1.84 0.70 10.63 6.98 37.10 2.12 1.30 19.07 20.68 17.20 0.81 0.50

Each value represents a single measurement.

The seeded clams had an average length of 16.0 ± 4.4 mm equivalent to a total fresh weight of 1.03 ± 0.79 g ind− 1. At the end of the farming cycle, clams attained 33.9 ± 0.3 mm and 11.3 ± 0.2 g ind− 1. The average density measured at the two cultivated fields was always lower compared to the seeding value and was between 343 ± 221 (field L) and 889 ± 297 (field H) ind m− 2. During the whole farming cycle, clam density remained constant in the field H, whereas in October, at site L, a slight decrease was detected, when empty shells were found on the sediment surface. Clams mortality was probably a consequence of the development of a thick macroalgal bed at site L with sediment anoxia. A few clams (mean density 45 ± 39 ind m− 2) were also harvested at field C. At the beginning of the experiment total clam biomass (fresh flesh and shell) was 0.6 ± 0.3 (L) and 1.4 ± 1.3 (H) kg m− 2 and raised up to 2.3 ± 4.0 (L) and 9.3 ± 0.7 (H) kg m− 2 at the end of the farming period (Fig. 2). The average elemental composition of dry flesh was C = 31.4 ± 3.0 mmol g− 1, N = 8.4 ± 0.6 mmol g− 1 and P = 0.2 ± 0.0 mmol g− 1, with no differences among sites (ANOVA, P N 0.05) (Table 3). The elemental composition of dry shells was C = 10, N = 0.2, P b 0.01 mmol g− 1. 3.3. Filtration rates and particulate nutrient fluxes Weight specific filtration rates (Table 4) ranged from 1 to 3 L g− 1 h− 1, with the highest values (2.8 ± 0.9 L g− 1

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D. Nizzoli et al. / Aquaculture 261 (2006) 98–108 Table 4 Weight specific filtration rates and PN and PP clearance rates (PN and PP uptake was calculated multiplying weight specific filtration rates for ambient PN and PP concentrations) Filtration rates −1

L April June August October

g− 1DW

−1

h

2.8 ± 0.9 1.2 ± 0.2 1.5 ± 0.3 1.4 ± 0.3

PN clearance −1

μmol g

DW

PP clearance −1

h

10.6 ± 3.6 12.6 ± 2.8 14.2 ± 2.8 6.3 ± 1.0

μmol g− 1DW h− 1 0.6 ± 0.2 0.7 ± 0.2 1.0 ± 0.2 0.4 ± 0.1

All data are referred to dry weight (DW) and are expressed as mean±standard deviation (n=6).

3.4. Dissolved phosphorous and nitrogen fluxes During the experiment TDP was net released from the benthic system to the water column (Fig. 4). Average release for the whole experimental area followed a general seasonal pattern with highest rates (between 2.5 ± 2.2 and 5.1 ± 3.5 mmol m− 2 d− 1) measured in summer samplings (ANOVA, P b 0.001). TDP regeneration increased progressively from June to August and from low to high densities, with 0.9 ± 0.4 (June) and 2.3 ± 1.3 (August) mmol m− 2 d− 1 at field C and 4.5 ± 2.5 (June) and 8.8 ± 3.7 (August) mmol m− 2 d− 1 at field H (Fig. 4). At field H fluxes resulted always significantly higher compared to field C (ANOVA, P b 0.001), whilst differences between C and L or L and H were statistically not significant (ANOVA, P N 0.05). SRP fluxes measured at C were negligible in April and October and slightly higher (0.4–0.6 mmol m− 2 d− 1) in the two summer samplings. Accordingly, the net SRP regeneration was from 6 to 7 folds higher at H compared to C. DOP fluxes were not statistically different among stations (ANOVA, P N 0.05). TDP fluxes, integrated over the rearing period, were 0.45 in the field L and 0.82 mol m− 2 in the field H, respectively 2 to 4 folds higher than those calculated for C (Table 5). NO3− fluxes (Fig. 5) were mostly directed towards the sediment and no significant differences were found between control and farmed fields (ANOVA, P N 0.05). Average fluxes resulted higher in April (−7.2 ± 14.7 mmol

Fig. 2. Average Ruditapes philippinarum total biomass (flesh plus shell fresh weight) (A) and flesh biomass (dry weight) (B) measured in C, L and H in the four samplings. Error bars represent one standard deviation (n = 3).

h− 1) measured in April (ANOVA, P b 0.01). Considering the standing biomass in each experimental field, the weight specific filtration rates corresponded to an amount of processed water ranging from 0.5 to 15 m3 m− 2 d− 1, which was equivalent to a subtraction from the water column of 50–120 mmol m− 2 d− 1 as N and 2.5–8 mmol m− 2 d− 1 as P (Fig. 3). The highest rates were measured in June and August (ANOVA, P b 0.001) in the farmed fields L and H. PN and PP uptake, integrated over the entire farming cycle, ranged from 9 (L) to 16 (H) mol m− 2 and from 0.6 (L) to 1 (H) mol m− 2 respectively (Table 5). Table 3 Average clam biomass and content of C, N and P in shell and flesh Biomass (g

April June August October

DW

ind− 1)

C (mmol g

−1 DW)

N (mmol g

−1 DW)

P (mmol g

−1 DW)

Shell

Flesh

Shell

Flesh

Shell

Flesh

Shell

Flesh

0.73 ± 0.17 3.92 ± 1.48 6.06 ± 2.67 5.36 ± 0.78

0.08 ± 0.02 0.51 ± 0.15 0.48 ± 0.17 0.26 ± 0.03

10.7 ± 0.3 10.8 ± 0.5 10.9 ± 0.2 10.7 ± 0.4

33.35 ± 1.27 32.48 ± 0.64 32.95 ± 1.20 26.90 ± 0.24

b0.2 b0.2 b0.2 b0.2

8.38 ± 0.46 8.17 ± 0.30 9.12 ± 0.26 7.75 ± 0.21

b0.01 b0.01 b0.01 b0.01

0.28 ± 0.03 0.40 ± 0.02 0.33 ± 0.01 0.27 ± 0.02

All data are referred to dry weight (DW) and are expressed as mean ± standard deviation (n = 3).

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Fig. 3. Mean daily water–sediment fluxes of PN (A) and PP (B) measured in C, L and H in the four samplings. Error bars represent one standard deviation (n = 3).

m− 2 d− 1) and October (−12.1 ± 6.4 mmol m− 2 d− 1) when NO3− availability in the water column was highest (20 and 90 μM, respectively). NO2− fluxes (data not shown) were usually negligible at all sampling sites. A net daily ammonium production was measured at all sampling fields and at all sampling dates with the exception of October when fluxes were extremely variable and not significantly different from zero (Fig 5). NH4+ efflux was significantly higher in the two summer samplings (ANOVA, P b 0.001). At C, ammonium efflux was negligible in April and raised up to 20.0 ± 4.7 mmol m− 2 d− 1 (June) and 11.7 ± 4.3 mmol m− 2 d− 1 (August). These fluxes were always significantly lower than those measured in the farmed sectors (ANOVA, P b 0.01). At L and H fields ammonium recycling was from 1.6 to 1.9 (June) and from 2.6 to 5 (August) folds higher compared to control sediments. NH4+ efflux was slightly higher at H compared to L, but differences were not significant (ANOVA, P N 0.05). Dissolved organic nitrogen fluxes

Table 5 Total amount of nutrients (as particulate and dissolved forms) subtracted (−) or released (+) to the water column from the benthic system −2

PN (mol m ) PP (mol m− 2) TDN (mol m− 2) TDP (mol m− 2)

C

L

H

− 1.7 ± 0.7 − 0.1 ± 0.0 − 0.3 ± 0.5 +0.20 ± 0.13

− 9.1 ± 2.3 (5.3) − 0.6 ± 0.2 (6.0) +1.6 ± 1.0 +0.45 ± 0.20 (2.2)

− 16.3 ± 3.8 (9.5) − 1.0 ± 0.2 (10.0) +9.1 ± 3.6 +0.82 ± 0.19 (4.1)

Values were integrated over the 7 months farming cycle from daily fluxes measured at sites C, L and H. Numbers in parenthesis refer to the stimulation of fluxes in farmed sites compared to control. All data are expressed as mean ± standard deviation (n = 3).

(data not shown) represented a minor fraction of TDN (b 20%) compared to ammonium and were affected by an extremely high variability which masked seasonal patterns and differences between sites. In the control site TDN fluxes integrated over the farming period resulted close to zero, whilst R. philippinarum rearing determined a net nitrogen release up to ∼1.6 mol m− 2 at L and ∼6.9 mol m− 2 at H (Table 5). 4. Discussion 4.1. Effects of R. philippinarum rearing on particulate and dissolved nitrogen and phosphorous fluxes In agreement with the previous studies we measured an increase in the downward fluxes of particulate nutrients and a parallel enhancement of dissolved forms release to the water column in the farmed zones. Weight specific filtration rates measured in this study were in the lower part of reported ranges (1.2 and 6 l h− 1 g− 1) for R. philippinarum (Goulletquer et al., 1989; Sorokin and Giovanardi, 1995; Nakamura, 2001). Such lower rates were probably the net result of different cooccurring factors, as clearance rates are known to be influenced by several environmental and physiological parameters such as temperature, oxygen and suspended particulate concentrations (Goulletquer et al., 1989; Lei et al., 1996; Hatcher et al., 1997). In particular optimal temperatures for R. philippinarum filtration activity ranged between 12 and 20 °C (Goulletquer et al., 1989). In our experiment water temperature was above this range in summer (N 25 °C), whilst in October, an optimal water temperature occurred at low chlorophyll-a concentration.

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Fig. 4. Mean daily water–sediment fluxes of TDP (A), SRP (B) and DOP (C) measured in C, L and H in the four samplings. Error bars represent one standard deviation (n = 3).

The PN and PP uptake estimated from fluxes at sites L and H over the whole farming period was 5 and 9 times greater than that measured at C. Clam densities at L and H were in the range of those usually found in the southern sandy fields of the Sacca di Goro, it is thus realistic to presume that in this area a major fraction of suspended particulate matter was retained in the lagoon by filter feeders that were responsible for faster transfer of particulate N and P from the water column to the benthic system. Over the entire farming cycle, TDP internal loading at L and H was from 2 to 4 times greater than that measured at C. Similarly, at the control site, sediments resulted a net sink for dissolved nitrogen whilst at L and H a considerable amount of nitrogen was released, mostly as ammonium. On average TDN and TDP effluxes increased between April to August, according

to temperature and biomass rise, and decreased in October. Higher fluxes measured in farmed areas can be explained by bacterial mineralisation of biodeposited labile faeces and pseudo faeces that in R. philippinarum populations can be extremely high and attain 1.4 g d− 1 ind− 1 (Jie et al., 2001). However, nutrients recycling due to direct excretion can be also significant. Magni et al. (2000) demonstrated that R. philippinarum excreted both inorganic N and P and that direct excretion can account up to the 90% of benthic fluxes. In October, macroalgae (Ulva spp) developed over the sediment surface and controlled both N and P release from the sediment of L and H, keeping benthic fluxes very low and similar in all the three fields. Assimilation by benthic primary producers is known to reduce the net release to the water column acting as a temporary sink for nutrients released from the sediment (Dalsgaard,

Fig. 5. Mean daily water–sediment fluxes of NH+4 (A) and NO−3 (B) measured in C, L and H in the four samplings. Error bars represent standard deviation (n = 3); negative bars mean a consumption from the sediment whilst positive bars represent a release.

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2003; Bartoli et al., 2001b). The effect of Ulva spp on benthic fluxes was evident in particular at site L, where algal biomass peaked at 850 g DW m− 2 (Nizzoli et al., in press). To date, there is a general consensus over the role of bivalve filter feeders as important mediators of N cycling but their role on P cycling as well as on the N and P stoichiometry is less clear. Some authors demonstrate that bivalve filter feeders do not influence significantly phosphorous recycling (Doering et al., 1987; Dame et al., 1991); others measured a preferential recycling of nitrogen compared to phosphorous in high densities oyster reefs (Sornin et al., 1986), mussels beds (Prins and Smaal, 1994) and mussel suspended ropes (Nizzoli et al., in press). Magni et al. (2000) measured a balanced release of nitrogen and phosphorous with incubation of single Ruditapes individuals, whilst Prins et al. (1995) observed a decrease in the N to P ratio of seston in relation to increasing mussel biomass. In this work NH4+ and SRP fluxes resulted significantly correlated in fields C (Kendall T P b 0.02) and H (Kendall T P b 0.001) but not in field L (Fig. 6). N to P ratio of seston varied between 14 to 18 (Table 1) and approximate to the Redfield ratio for phytoplankton (Redfield et al., 1963). NH4+ to SRP fluxes on the contrary indicate that in the control field there was a preferential release of inorganic N compared to inorganic P (average N:P ratio 31). At L and H the effect of clams farming on NH4+ and SRP stoichiometry was not univocal. At site H in fact, the N to P ratio of inorganic nutrients released to the water column was very similar to that of the filtered material and comprised between 11 and 19, whilst at site L a preferential release of N was observed. As SRP is retained by sediments due to interactions with Fe and Ca it is likely that P net efflux is not linearly correlated with clams biomass. At low

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clams densities in fact, a slow SRP release can result from sediment retention, whilst at high clam densities, the saturation of P exchange sites results in a net efflux and in a more balanced NH4+ to SRP regeneration ratio. However further studies are needed in order to clarify this relationship, because not only the total amount of nutrient recycled to the water column but also its N and P stoichiometry could influence the primary productivity and the structure of primary producers community (Cloern, 2001). 4.2. N and P exported as clams biomass and nutrients recycling: a comparison Within aquaculture activities, the farming of bivalve filter feeders is considered to have little impacts on the natural environment because filtration activity imposes a top down control on seston and, as bivalves grow, a fraction of particulate matter is converted into biomass. Thus, at the end of farming cycles, crop harvesting results in a net removal from the system of the nutrients incorporated into the biomass (Kaspar et al., 1985). Bivalves as Mytilus edulis have indeed been proposed as biofilters for water purification in fish aquaculture facilities (Troell and Norberg, 1998). At the end of our farming cycle, the net amount of nitrogen and phosphorous stored in biomass and potentially exported from the system with clams harvesting was easily calculated from elemental composition of clams flesh and shell (Table 3), individual weight and shell to flesh ratio. Seeding at the beginning of the cycle determined a N and P input to the system of 0.8 ± 0.2 and 0.02 ± 0.00 μmol ind− 1 respectively whilst clams harvesting accounted for a N and P export of about 2.7 ± 0.3 and 0.07 ± 0.01 μmol ind− 1. Consequently the potential net N and P export corrected for clams densities, gave

Fig. 6. Daily NH+4 versus SRP fluxes measured in each single core collected in control (left) and farmed sites (right) from April to August. Dashed lines represent the average N:P ratio of seston.

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0.36 (L) and 1.81 (H) mol m− 2 as N and 0.01 (L) and 0.04 (H) mol m− 2 as P. These quantities accounted for a small percentage of the total amount of nutrients cycled by the clam + sediment system in the farming period (Fig. 6). In particular, N export represented on average ∼ 6% of the filtered material and ∼ 30% of that recycled to the water column. The potential export of P represented only ∼ 3% of both the filtered and recycled matter (Fig. 7). Our results suggest that clams farming has the potential to control particulate matter in the water column but it also fasten its conversion into dissolved solutes. With respect to nutrient abatement the clams + sediment system seems thus only a temporary nutrients

Fig. 7. Fluxes of N (upper) and P (lower) subtracted from the water column (filtered PN and PP), released from sediments (TDN and TDP) and potentially exported as crop (N and P in shell and flesh) at C, L and H. Units are mol m− 2 farming period− 1 where the latter is 7 months.

sink probably due to elevated excretion rates and lability of the settled material. Furthermore, as demonstrated here, the amount of N and P removed with harvested clam is of little significance compared to overall incoming (discussed later in the text) and cycled loads. 4.3. Implications of mollusc farming on the ecological status of the Sacca di Goro Semi-enclosed coastal lagoons as the Sacca di Goro are transitional ecosystems where high loads of dissolved and particulate nutrients are processed during their transport from terrestrial to marine environments (Giordani et al., 2005). Coastal lagoons are considered sites with high primary production and mineralisation rates as well as sites where sedimentation and burial subtract material from the water column. However, the fate of incoming nutrients is difficult to predict accurately due to the number of theoretically co-occurring biogeochemical processes (assimilation, sedimentation, adsorption, denitrification, chemical reduction, bioturbation, etc.). During the 7 months of the experiment duration the loading delivered by freshwater sources was 328 t TDN, 15 t TDP, 187 t PN and 25 t PP, accordingly with the previous studies (Viaroli et al., 2006). The amount of N and P cycling at the lagoon level through filtration, assimilation, regeneration and burial pathways can be tentatively estimated assuming a clam market size of 10 g wet weight (shell length ∼3 cm) and a crop of 6000 tons. This crop has been calculated as half of the total biomass produced in the Sacca di Goro (∼12 000 tons y− 1) and is realistic for the period (April–October) covered by our farming cycle. Results from this study indicate that during its growth R. philippinarum subtracts approximately 0.23 g N ind− 1 and 0.03 g P ind− 1 and recycles approximately 0.15 g N ind− 1 and 0.02 g P ind− 1. For 6000 tons of clam produced, approximately 137 tons of PN and 19 tons of PP are subtracted from the water column, about 91 tons of TDN and 12 tons of TDP are recycled back whilst ∼16 t of N and ∼0.9 t of P are net removed by harvesting. The comparison of these budgets with dissolved and particulate external loads indicates that the amount of particulate matter processed by clams is of the same order of magnitude of that transported in the lagoon by freshwater inputs; the regenerated fraction is about 30 and 90% of external TDN and TDP loads. This is somehow a rough comparison as it does not consider nutrients supplied by autochthonous phytoplankton growth or entered in the lagoon with tidal currents from the surrounding sea. Moreover, the extrapolation at the ecosystem level of the results of our experiment, carried out in a small area with fixed densities, is a difficult task and

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must be considered with some caution. Nevertheless our results and speculations stress the importance of clams cultivation and its effects on the alteration of nutrient dynamics within the lagoon; in particular, clams seems to act as regulators of the ratio between particulate and dissolved forms. Our results clearly demonstrate that cultivated molluscs modify rates of material fluxes across the sediment–water interface and that in the Sacca di Goro a significant fraction of the particulate external loads is retained and recycled as dissolved nutrients through the sediment+ clams combined action. The main actions realised in recent years to mitigate the Sacca di Goro eutrophication are the reduction of fresh water inputs, diverting many lagoon tributaries to the sea, the dredging of tidal channel to improve water circulation, and the macroalgae collection to avoid dystrophy and clams production collapses (Simeoni et al., 2000). Results presented in this paper indicate that clams farming has the potential to retain nutrients in coastal areas with negative implications for sediment and water quality. The retention of particulate matter in cultivated areas and the alteration of particulate to dissolved nutrient ratio can affect negatively water and sediment quality and stimulate primary productivity. In particular, a positive feedback between bivalves farming and the growth of nuisance macroalgae, not controlled by top down grazing and favoured by a bottom up release of nutrients, should be considered with special attention. Based on these results we suggest to carefully consider the width of farmed surface in coastal lagoons in order to minimize the impact of the internal recycling mediated by clams farming. Acknowledgements This research was supported with funding from the Italian Ministry for Agriculture and Forestry (Ricerca N°6C.76 Prof Giulio De Leo) and partially with funding from the DITTY E.U. project (Development of an Information Technology Tool for the Management of European Southern Lagoons under the influence of riverbasin runoff) contract n° EVK3-2001-00226. The authors are indebted to Edoardo Turolla of the Centre for Mollusc Research, Goro for the provision of laboratory facilities, useful discussion and criticism and helpful information on clam and farming practises in the Sacca di Goro. References APHA, 1975. Standard Methods for the Examination of Water and Wastewaters, 14th ed. A.P.H.A, Washington, USA. 1193 pp. Aspila, K.I., Agemian, H., Chau, A.S.Y., 1976. A semiautomated method for the determination of inorganic, organic and total phosphate in sediments. Analyst 101, 187–197.

107

Bartoli, M., Nizzoli, D., Viaroli, P., Turolla, E., Castaldelli, G., Fano, E.A., Rossi, R., 2001a. Impact of Ruditapes philippinarum farming on nutrient dynamics and benthic respiration in the Sacca di Goro. Hydrobiologia 455, 203–212. Bartoli, M., Castaldelli, G., Nizzoli, D., Gatti, L.G., Viaroli, P., 2001b. Benthic fluxes of oxygen, ammonium and nitrate and coupled and uncoupled denitrification rates in three eutrophic coastal lagoons with different primary producers. In: Faranda, M., Guglielmo, L., Spezie, G. (Eds.), Structures and Processes in the Mediterranean Ecosystems. Springer Verlag. Baudinet, D., Alliot, E., Berland, B., Grenz, C., Plante-Cuny, M.R., Plante, R., Salen-Picard, C., 1990. Incidence of mussel culture on biogeochemical fluxes at the sediment–water interface. Hydrobiologia 207, 187–196. Bower, C.E., Holm-Hansen, T., 1980. A salicylate–hypoclorite method for determining ammonia in seawater. Canadian Journal of Fisheries and Aquatic Sciences 37, 794–798. Castaldelli, G., Mantovani, S., Welsh, D.T., Rossi, R., Mistri, M., Fano, E.A., 2003. Impact of commercial clam harvesting on water column and sediment physicochemical characteristics and macrobenthic community structure in a lagoon (Sacca di Goro) of the Po River Delta. Chemistry and Ecology 19, 161–171. Cloern, J.E., 2001. Our evolving conceptual model of the coastal eutrophication problem. Marine Ecology. Progress Series 210, 223–253. Dahlbäck, B., Gunnarsson, L.H.A.H., 1981. Sedimentation and sulfate reduction under a mussel culture. Marine Biology 63, 269–275. Dalsgaard, T., 2003. Benthic primary production and nutrient cycling in sediments with benthic microalgae and transient accumulation of macroalgae. Limnology and Oceanography 48, 2138–2150. Dame, R.F., Dankers, N., Prins, T.C., Jongsma, H., Smaal, A.C., 1991. The influence of mussel beds on nutrients in the western Wadden Sea and eastern Scheldt estuaries. Estuaries 14, 130–138. Doering, P.H., Kelly, J.R., Oviatt, C.A., Sowers, T., 1987. Effect of the hard clam Mercenaria mercenaria on benthic fluxes of inorganic nutrients and gases. Marine Biology 94, 377–383. FAO (Food and Agricolture Organization of the United States), 2003. The State of World Fisheries and Aquaculture 2002. Food and Agriculture Organization of the United States, Rome. Giordani, G., Viaroli, P., Swaney, D.P., Murray, C.N., Zaldívar, J.M., Marshall Crossland, J.I. (Eds.), 2005. Nutrient fluxes in transitional zones of the Italian coast. LOICZ Reports and Studies, vol. 28. LOICZ, Texel, the Netherlands. ii+157 pages. Goulletquer, P., Heral, M., Deslous-Paoli, J.M., Prou, J., Garnier, J., Radez, D., Boromthanarat, W., 1989. Ecophysiologie et bilan energetique de la palourde japonaise d'elevage Ruditapes philippinarum. Journal of Experimental Marine Biology and Ecology 132, 85–108. Hatcher, A., Grant, J., Schofield, B., 1994. Effects of suspended mussel culture (Mytilus spp.) on sedimentation, benthic respiration and sediment dynamics in a coastal bay. Journal of Experimental Marine Biology and Ecology 115, 219–235. Hatcher, A., Grant, J., Schofield, B., 1997. Seasonal changes in the metabolism of cultured mussels (Mytilus edulis l.) from a Nova Scotian inlet: the effects of winter ice cover and nutritive stress. Journal of Experimental Marine Biology and Ecology 217, 63–78. Lei, J., Payne, B.S., Wang, S.Y., 1996. Filtration dynamics of the zebra mussel, Dreissena polymorpha. Canadian journal of fisheries and aquatic sciences 53, 29–37. Jie, H., Zhang, Z., Zishan, Y., Widdows, J., 2001. Differences in the benthic–pelagic particle flux (biodeposition and sediment erosion) at intertidal sites with and without clam (Ruditapes philippinarum)

108

D. Nizzoli et al. / Aquaculture 261 (2006) 98–108

cultivation in eastern China. Journal of Experimental Marine Biology and Ecology 261, 245–261. Kaspar, H.F., Gillespie, P.A., Boyer, I.C., MacKenzie, A.L., 1985. Effects of mussel aquaculture on the nitrogen cycle and benthic communities in Kenepuru Sound, Marlborough Sounds, New Zealand. Marine Biology 85, 127–136. Magni, P., Montani, S., Takada, C., Tsutsumi, H., 2000. Temporal scaling and relevance of bivalve nutrient excretion on a tidal flat of the Seto Inland Sea, Japan. Marine Ecology. Progress Series 198, 139–155. Mazouni, N., Gaertner, J.C., Deslous-Paoli, J.M., Landrein, S., Geringer d'Oedenberg, M., 1996. Nutrient and oxygen exchanges at the water–sediment interface in a shellfish farming lagoon (Thau, France). Journal of Experimental Marine Biology and Ecology 203, 92–113. Nakamura, Y., 2001. Filtration rates of the Manila clam, Ruditapes philippinarum: dependence on prey items including bacteria and picocyanobacteria. Journal of Experimental Marine Biology and Ecology 266, 181–192. Newell, R.I.E., 2004. Ecosystem influences of natural and cultivated populations of suspension-feeding bivalve molluscs: a review. Journal of Shellfish Research 23, 51–61. Newell, R.I.E., Cornwell, J.C., Owens, M.S., 2002. Influence of simulated bivalve biodeposition and microphytobenthos on sediment nitrogen dynamics: a laboratory study. Limnology and Oceanography 47, 1367–1379. Nizzoli, D., Welsh, D.T., Fano, E.A., Viaroli, P., 2006a. 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., Viaroli, P., in press. Oxygen and ammonium dynamics during a farming cycle of the bivalve Tapes philippinarum. Hydrobiologia. Prins, T.C., Smaal, C.S., 1994. The role of the blue mussel Mytilus edulis in the cycling of nutrients in the Oosterschelde estuary (The Netherlands). Hydrobiologia 282/283, 413–429. Prins, T.C., Escaravage, V., Smaal, A.C., Peeters, J.C.H., 1995. Nutrient cycling and phytoplankton dynamics in relation to mussel grazing in a mesocosm experiment. Ophelia 41, 289–315.

Prins, T.C., Smaal, C.S., Dame, R.F., 1998. A review of the feedbacks between bivalve grazing and ecosystem processes. Aquatic Ecology 31, 349–359. Redfield, A.C., Ketchum, B.H., Richards, F.A., 1963. The influence of organisms on the composition of seawater. In: Hill, M.N. (Ed.), The Sea, vol. II, pp. 26–77. Rossi, R., Paesanti, F., 1992. Successful clam farming in Italy. Proc. XXIII Annual Shellfish Conference, 19 and 20 May 1992, Fishmongers' Hall, London, pp. 62–68. Simeoni, U., Dal Cin, R., Fontolan, G., Tessari, U., 2000. Morfogenesi ed evoluzione dello Scanno di Goro (Delta del Po). In: Simeoni, U. (Ed.), Studi costieri, vol. 2, pp. 5–20. Sokal, R.R., Rolf, F.J., 1995. Biometry. W.H. Freeman and Company, New York. Sorokin, Y.I., Giovanardi, O., 1995. Trophic characteristics of the Manila clam (Tapes philippinarum Adams and Reeve). Journal of Marine Science 52, 853–862. Sornin, J.M., Feulillet, M., Heral, M., Fardeu, J.C., 1986. Influence des cultures D'huitres Crassostrea gigas sur le cycle du phosphore en zone intertidale: role de la biodeposition. Oceanologica Acta 9, 313–322. Troell, M., Norberg, J., 1998. Modelling output and retention of suspended solids in an integrated salmon-mussel culture. Ecological Modelling 110, 65–77. Valderrama, J.C., 1977. Methods used by the Hydrographic Department of National Board of Fisheries, Sweden. In: Grasshof, K. (Ed.), Report of the Baltic Intercalibraton Workshop. Annex, Interim Commission for the Protection of the Environment of the Baltic Sea, pp. 13–40. Valderrama, J.C., 1981. The simultaneous analysis of total nitrogen and total phosphorus in natural waters. Marine Chemistry 10, 109–122. Viaroli, P., Giordani, G., Bartoli, M., Naldi, M., Azzoni, R., Nizzoli, D., Ferrari, I., Zaldívar Comenges, J.M., Bencivelli, S., Castaldelli, G., Fano, E.A., 2006. The Sacca di Goro lagoon and an arm of the Po river. In: The Handbook of Environmental Chemistry (Ed. in Chief: O. Hutzinger) Volume 5. Water pollution: estuaries. Volume editor: P.J. Wangersky Chapter: pp. 197–232.