Impacts of maintenance channel dredging in a northern Adriatic coastal lagoon. II: Effects on macrobenthic assemblages in channels and ponds

Estuarine, Coastal and Shelf Science 85 (2009) 143–150

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Impacts of maintenance channel dredging in a northern Adriatic coastal lagoon. II: Effects on macrobenthic assemblages in channels and ponds Massimo Ponti*, Andrea Pasteris, Roberta Guerra, Marco Abbiati ` di Bologna, Via S. Alberto, 163, 48100 Ravenna, Italy Centro Interdipartimentale di Ricerca per le Scienze Ambientali (C.I.R.S.A.), Universita

a r t i c l e i n f o

a b s t r a c t

Article history: Received 5 December 2008 Accepted 24 June 2009 Available online 4 July 2009

Coastal lagoons are ephemeral habitats whose conservation requires human intervention, such as maintenance dredging of inner channels. Dredging can reduce the abundance of benthic species due to the removal of individuals with the sediment, modify sediment properties, and resuspend fine sediment, nutrients and pollutants, which can lead to eutrophication, hypoxic events and increasing toxicity. Both direct effects in the dredged channel and possible indirect effects in surrounding shallow areas could be expected. This study assesses the effects of the channel maintenance dredging, performed between October 2004 and August 2005, on the invertebrate assemblages both in channels and adjacent ponds in the northern Adriatic coastal lagoon of Pialassa Baiona. The lagoon is affected by eutrophication, chemical and thermal pollution from wastewater treatment and power plants. Three impacted sites were located in the dredged channel and three in the adjacent interconnected shallow water ponds, while three nonimpacted sites were located in a channel and in a pond far from the dredged area. Replicate samples were collected from each site one time before and one time after the dredging operations. Despite the extent of the intervention, effects of the dredging on macrobenthic assemblages were detected only within the dredged channel, while in the surrounding ponds no clear and unequivocal effects were found. In particular the dredging could have promoted the increase of the abundance of the polychaete Streblospio shrubsolii in the southern and central parts of the dredged channel and the increase in abundance of the amphipod Corophium insidiosum in the northern side, compared to the controls. Instead, species diversity was reduced in the central and northern parts of the dredged channel. These effects on the macrobenthic invertebrate assemblages could be related to the observed changes of sediment characteristics, contamination and toxicity. Overall, direct effects on benthic assemblages in the dredged channels were more detectable than the possible secondary effects in the surrounding shallow ponds, where the higher spatial heterogeneity can mask any relevant effects. Ó 2009 Elsevier Ltd. All rights reserved.

Keywords: dredging coastal lagoon benthic infauna impact assessment BACI sampling design Mediterranean Italy northern Adriatic Sea

1. Introduction Conservation and management of coastal lagoons often requires direct human intervention. To prevent siltation and to maintain the hydrodynamic features of lagoonal systems, periodical dredging of lagoon openings and channels is needed. Dredging activities may have four principal short-term effects on benthic assemblages (Quigley and Hall, 1999 and references therein): (1) reduction of number of species richness and abundances directly related to the disturbance event; (2) change in sediment properties (e.g. grain size) modifying relevant habitat features; (3) resuspension of fine

* Corresponding author at: Laboratori di Scienze Ambientali, Universita` di Bologna, Via S. Alberto 163, 48100 Ravenna, Italy. E-mail address: [email protected] (M. Ponti). 0272-7714/$ – see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.ecss.2009.06.027

sediment and associated nutrients, organic matter and pollutants, which can lead to eutrophication, hypoxic events and increasing toxicity; and (4) habitat loss and ecosystem function reduction. Over long-term some other effects on macrobenthic assemblages could be also expected due to: (1) increased water circulation and oxygenation; (2) acceleration of geochemical processes (e.g. organic mater mineralization); and (3) new available space for larval settlement and recruitment. Dredging operations are generally performed in channels, and seldom in the shallow water bottoms. Resuspension of fine sediments due to the dredging operations could increase turbidity and sedimentation rates for hundreds of meters in the surrounding areas. Albeit unintended, sediment fallout my affect benthic communities (Wilber et al., 2007). To assess the impact of any human activities on the structure of benthic invertebrate assemblage, spatial and temporal scales of

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variability of the assemblages has to be taken into account (Boyd et al., 2003; Josefson et al., 2008; Dauvin, 2008). This study assesses the effects of maintenance channel dredging performed from October 2004 to August 2005 on macrobenthic assemblages in both channels and ponds in the northern Adriatic Italian coastal lagoon of Pialassa Baiona, using a sampling design based on the ‘‘before/after and control/impact’’ (BACI) approach (Green, 1979; Underwood, 1992). Changes in the abundance of numerically dominant taxa, species diversity and assemblage structure in response to dredging in the muddy sandy sediments were compared with appropriate control sites both in channels and ponds. Observed changes were related to the concomitant alteration of sediment characteristics, contaminant concentrations and toxicity, which are analysed in detail in Guerra et al. (2009). 2. Materials and methods 2.1. Study area and description of the dredging activities This study was carried out in the Pialassa Baiona (44 300 N, 12 150 E), a eutrophic microtidal lagoonal system located along the northern Adriatic coast of Italy (Fig. 1). This area was declared a wetland of international importance under the Ramsar Convention in 1981, and is included among the sites of Community interest (pSCIs, ‘‘Habitat’’ Directive 92/43/EEC). At present the lagoon is exploited for recreational fishing, Manila clam (Ruditapes philippinarum) harvesting, regulated hunting, leisure and cultural activities such as canoeing, guided walking tours, and bird watching. The lagoonal system includes brackish and fresh water shallow ponds, with an average depth of 1 m, completely or partially isolated by embankments and crossed by a network of artificial channels dug since 1850. Channel depths range from 1 m to 8 m in the landward and seaward sides, respectively. The lagoon is connected to the sea by a channel connecting it to the Ravenna harbour. The lagoon covers an area of about 10 km2, the tidal

range can exceptionally exceed 1 m, vast shallow areas may emerge during low tides. Sediments vary from sandy to muddy (sand range from 12.1% to 89.5% in weight) according to the occurrence of active sedimentation processes or relict sand dunes. The southern area of the lagoon receives wastewater from urban and industrial sewage treatment plants, and from two thermal power stations (Ponti et al., 2005). Overall water turnover in the lagoon has been estimated in 3 days. The macrobenthic invertebrate assemblages inhabiting muddy and sandy bottoms mainly include spionid and capitellid polychaetes, tubificid oligochaeets, amphipods, bivalves and gastropods. Although Pialassa Baiona is one of the most anthropogenically disturbed coastal lagoons in the Mediterranean Sea, given its proximity to important urban, industrial and harbour areas, it hosts a relatively diverse benthic assemblages (Ponti et al., 2008). Phytoplankton blooms and intense growth of seaweeds (Ulva sp., Enteromorpha sp., Gracilaria sp.) have been frequently observed during the summer, especially in the southern part. Dystrophic crises were often recorded in the summer, which can affect the structure of the benthic assemblages, mainly reducing the abundance of amphipods (Ponti and Abbiati, 2004). From 1958 to 1976 Pialassa Baiona was heavily impacted by industrial pollution. Mercury, polycyclic aromatic hydrocarbons (PAHs), and synthetic polymers were among the most important pollutants which now contaminate the sedimentary compartment (Fabbri et al., 1998, 2000, 2003; Abbondanzi et al., 2005). Superficial sediments are still contaminated, with a total mercury load ranging from 0.13 to 250 mg g1 dry weight (Trombini et al., 2003). The main channels of the lagoon are periodically dredged to favour the water flow and exchange with the sea. From October 2004 to August 2005 Baccarini Channel, one of the main channels that crosses the lagoonal system, was dredged by draglines and excavators placed on a floating pontoon, in order to re-establish original depth, restore embankments and improve water circulation. Dredging intervention was done along 4000 m of the channel,

Fig. 1. Map of Pialassa Baiona lagoon, showing sampling sites located in channels as circles, and ponds as triangles (geographic grid UTM 32T, European Datum 1950).

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starting from the northern side, and about 100,000 m3 of sediment was moved. 2.2. Sampling design and laboratory analyses The sampling design was based on the BACI approach. Three impacted sites were located in the dredged Baccarini Channel (labelled BAC1, BAC3 and BAC5) and three in the adjacent ponds (POL1, POL3 and VEN5; Fig. 1). Each putatively disturbed site, hereafter referred to as ‘‘Impact’’ (I), was compared to three undisturbed ‘‘Control’’ sites (Cs) located in a channel (TBF1, TBF3 and TBF4) and in a pond (RIS1, RIS2, RIS3) not influenced by dredging operations, to consider the natural spatial heterogeneity and avoid spatial confounding in the impact assessment (Underwood, 1992, 1994). Availability of resources dictated that sampling was performed in only one time before and one time after dredging. Admittedly, this is a major drawback in comparison to a strict ‘‘beyond BACI’’ design, which requires replicated dates before and after the disturbance event (Underwood, 1992, 1994). However, the presence of multiple control sites should attenuate the possible confounding effect due to absence of replication in time. This issue will be considered in detail in the discussion. Three replicated samples of the benthic assemblages were collected with a WildcoÒ box corer (sampling area of 0.0225 m2) at each site before (September, 2004) and after (September, 2005) the dredging operations. After a pre-sieving (0.5 mm mesh) in the field, samples were preserved using a buffered solution of 4% formaldehyde. Specimens were identified to the lowest possible taxonomic level, using a binocular microscope, and counted. 2.3. Data analyses Since dredging operations were done in the Baccarini Channel and not in the adjacent ponds, the potential impacts in channel and pond habitats have been analysed separately. Moreover, since the dredging operation lasted a long time with frequent suspension, the physical disturbance could have affected the impacted sites at different times and in different ways. Therefore, the disturbance can not be considered as a single and univocal event; consequently, each putative impacted site was considered individually and compared to the other three control sites. Data were analysed using an asymmetrical model consisting of 2 factors: Time (T, 2 levels, fixed: Before and After, Site (S, random, orthogonal to Time, with 1 putatively Impacted and 3 Control sites), with n ¼ 3 observations per combination of factor levels. If the dredging had an impact, the putatively impacted sites (I) will change over time with a different pattern when compared to control sites (Cs). This difference could be detected as a significant T  S interaction. The T  S mean square was divided into 2 portions: a T  I vs. Cs and a T  Cs interaction. If the T  Cs component is not significant, impact would be detected as a significant T  I vs. Cs interaction, otherwise impact would be detected only when T  I vs. Cs interaction is significantly larger than T  Cs interaction. If there are any interactions between times and sites among controls (T  Cs: P > 0.05) then the appropriate F denominator for the test T  I vs. Cs is the residual (degrees of freedom, df ¼ 16), otherwise the denominator became T  Cs (df ¼ 2) (Underwood, 1992). The latter has lower power, because of lesser degrees of freedom. Species richness (as number of taxa, S) and Shannon’s diversity index (H0 based on log2) were calculated for each replicate sample (Gray, 2000). Differences in species abundance and species diversity indices were analysed by ANOVA, in accordance with the experimental design. Cochran’s C test was used to check the assumption of homogeneity of variances and, when necessary, appropriate transformations were applied. In the case in which

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variances were slightly heterogeneous even after transformations (Cochran’s C test 0.05 > P > 0.01), the analyses were run at a ¼ 0.01, whereas when variances were highly heterogeneous (Cochran’s C test P < 0.01) only not significant results were retained (Underwood, 1997). Bonferroni’s correction for multiple tests was not applied. Similarity between assemblages among the study sites and times were analysed using principal coordinate analysis (PCO, i.e. metric multidimensional scaling) based on Bray–Curtis dissimilarities of square root transformed data (Anderson, 2003; Anderson and Willis, 2003). Each sampling site, before and after the dredging operations, was represented on the PCO bi-dimensional ordination plot by the centroid of three replicates. Differences were assessed by a distance-based permutational multivariate analysis of variance (PERMANOVA) following the same multi-factor experimental design used in the ANOVA (McArdle and Anderson, 2001; Anderson, 2001, 2005), including all contrasts and partitions involving impact and control locations. Following the example provided by Terlizzi et al. (2005), each term in the analysis was coded as a design matrix and tested individually with the appropriate denominator and relevant permutable units using the computer program DISTLM (Anderson, 2004). 3. Results A total of 56 taxa of macrobenthic invertebrates were recorded, and 48 identified to species level. In the channels, the most abundant taxa were the polychaetes Streblospio shrubsolii (Buchanan, 1890) and Capitella capitata (Fabricius, 1780), the amphipod Corophium insidiosum Crawford, 1937, and unidentified tubificid oligochaetes. In addition to these taxa, the gastropod Hydrobia ventrosa (Montagu, 1803), the invasive bivalves Musculista senhousia (Benson in Cantor, 1842), the isopod Idotea balthica (Pallas, 1772), the amphipod Gammarus aequicauda (Martyinov, 1931) and the larvae of the midge Chironomus salinarius (Kieffer, 1921) largely contribute to the total abundance of the macrobenthic invertebrates found in the ponds. 3.1. Dredging effects in the channels The first two axes of the PCO ordination plot clearly represent the overall changes in the macrobenthic invertebrate assemblages in the channels from before to after the dredging (Fig. 2a). The assemblages found in the three control sites (TBF1, TBF3, TBF4) showed consistent temporal patterns among them in terms of the direction and intensity of changes from before to after the intervention, also confirmed by the not significant multivariate T  Cs interaction (Table 1a). On the contrary, assemblages in the dredged sites showed different trends compared to the Cs (Fig. 2a). Multivariate T  I vs. Cs interactions are highly significant at two impacted sites (BAC3 and BAC5) and not significant (even if P ¼ 0.0554) at the third site (BAC1), the patterns of which could be ascribed to the dredging (Table 1a). Among the four most abundant taxa found in the channels, only two species showed variations in abundance, which could be related to the dredging event. The abundance of the polychaete Streblospio shrubsolii decreased in all control sites after the intervention (T  Cs not significant), but increased in the dredged sites BAC1 and BAC3 (T  I vs. Cs highly significant; Fig. 3a and Table 2). At the same time, the abundance of the amphipod Corophium insidiosum increased in the dredged site BAC5 compared to the controls (T  Cs not significant and T  I vs. Cs highly significant; Fig. 3b and Table 2). Species richness remains almost constant in the control sites from before to after the dredging (T  Cs not significant), yet

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did not show change that could be related to the intervention (T  I vs. Cs not significant, Table 1b). It should be noted that in this multivariate analysis the appropriate test for the impact effect was T  Cs. No species living in the ponds showed patterns, which could be related to the dredging impact. The abundance of the polychaete Capitella capitata was drastically reduced in both sites closer to the dredged channel, and in all of the controls (Fig. 4a). Although the ANOVA detected a significant T  I vs. Cs interaction in POL3, this is due to the near absence of this species in this site both before and after the dredging. Absence of C. capitata was in contrast with the reduction in abundance observed in the Cs but it cannot be related to the dredging (Fig. 4a and Table 3). The polychaete Streblospio shrubsolii, which was abundant (197.0  149.4 ind. sample1  SE) in the sampling site VEN5 and very rare in the other sites before the intervention, disappeared in VEN5 after the dredging (Fig. 4b). However this difference was not statistically significant (Table 3). Conversely, the amphipod Gammarus aequicauda increased from zero to 79.7  15.2 ind. sample1 (SE) in POL1, but even these differences are not significant according to the ANOVA test (Table 4). Tubificid oligochaetes, Musculista senhousia, Hydrobia ventrosa, Corophium insidiosum, Idotea balthica, and Chironomus salinarius did not show effects related to the dredging (Table 4). Species richness and Shannon’s diversity index showed heterogeneous trends both in I and Cs sites and no changes are related to the intervention (Figs. 4c,d and Table 3). 4. Discussion

Fig. 2. Principal coordinate ordination plots (PCO) showing different temporal paths (before and after) in impacted and control sites in the channels (a) and in the ponds (b). Each point represents the centroid of the observed similarities.

seemed to decrease in the impacted site BAC5 and increase in BAC1 (Fig. 3c and Table 2). However, but these results cannot be confirmed by the ANOVA test (run at a ¼ 0.01 due to the heterogeneity of variances). Species diversity, in terms of Shannon’s index, showed a consistent increase in the control sites (T  Cs not significant) while significantly decreased in the dredged sites BAC3 and BAC5 (T  I vs. Cs significant; Fig. 3d and Table 2). 3.2. Dredging effects in the ponds The first two axes of the second PCO ordination plot depict well the overall changes in the macrobenthic invertebrate assemblages, which occurred in the ponds from before to after the dredging operation (Fig. 2b). The assemblages found in the three control sites (RIS1, RIS2, RIS3) presented inconsistent temporal shifts in terms of intensity and direction, as highlighted by the highly significant multivariate T  Cs interaction (Table 1b). Overall, the assemblages in the putatively impacted sites, located near the dredged channel,

Any sampling design is a compromise between formal requirements and practical constrains. In particular, in many instances, environmental impacts are assessed using far from optimal sampling designs. For example, impact is often inferred only from samples collected after the event of concern has occurred (after-control-impact, ACI; e.g. Underwood et al., 2003; Rossi et al., 2007). In ACI designs, differences between a putatively impacted site and one or more control sites are interpreted as an evidence of impact. This is questionable, as it is not possible to demonstrate that the differences were not already present prior to the event. Even the beyond BACI design (Underwood, 1992), usually considered among the most rigorous approaches to impact detection, proves to be sub-optimal since the impact ‘‘treatment’’ is not replicated. Moreover, real word applications, despite the efforts of the investigators, are often far from ideal, since replication is kept to the minimum and appropriateness of control sites may be questionable (e.g. Skilleter et al., 2006). Imperfect sampling designs are often used, in full awareness of their limitations, because of the economical and temporal constraints. In the case at hand, resource was such that sampling was undertaken only once before, and once after dredging. Lack of replication in time is the main weakness of this study, and may lead to confounding effects due to the lack of information on temporal trends in Impacted and Control Sites. This point has to be considered in drawing the conclusion on presence or absence of ecological effects of the dredging activities. However, inclusion in the sampling design of multiple control sites attenuate the possible confounding effect, and allows to made remarks on the actual impact of dredging. Significant T  I vs. Cs interaction, together with non-significant T  Cs interaction shows that changes occurred between Before and After time in putatively impacted site versus the mean of the control sites, while differences among control sites remain consistent. This means that the single site (I) has a different pattern of change, compared with the general trend represented by the multiple control sites, which seems fairly unlikely in absence of an actual impact.

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Table 1 Summary of asymmetrical PERMANOVA on benthic invertebrate assemblages (Bray–Curtis dissimilarities on square root transformed data, permutation of raw data) comparing each disturbed site (I) with control sites (Cs), both in the channels (a) and in the ponds (b). Source

df

MS

Pseudo-F

P (perm)

F vs.

(a) Channels TCs ResCs

2 12

3089.5 2038.7

1.515

0.1023

ResCs

T  IBAC1 vs. Cs ResBAC1 þ Cs T  IBAC3 vs. Cs ResBAC3 þ Cs T  IBAC5 vs. Cs ResBAC5 þ Cs

1 16 1 16 1 16

4758.4 2643.1 9362.5 2132.1 6263.9 1623.8

1.800

0.0554

ResBAC1 þ Cs

4.391

0.0004

ResBAC3 þ Cs

3.858

0.0025

ResBAC5 þ Cs

(b) Ponds T  Cs ResCs

2 12

4592.1 1471.0

3.122

0.0002

ResCs

1 2 1 2 1 2

5896.2 4592.1 3340.7 4592.1 2931.7 4592.1

1.284

0.3125

T  Cs

0.727

0.7038

T  Cs

0.638

0.7688

T  Cs

T  IPOL1 vs. Cs T  Cs T  IPOL3 vs. Cs T  Cs T  IVEN5 vs. Cs T  Cs

In the Pialassa Baiona lagoonal systems, macrobenthic invertebrate assemblages are affected by several anthropogenic disturbances, including pollution and summertime dystrophic events that affect the structure and dynamics of the assemblages (Ponti and Abbiati, 2004). Some of the trace metals analysed by Guerra et al. (2009) in the sediments of this lagoon exceeded the effects range low (ERL), i.e. Cr, and effects range medium (ERM), i.e. Hg and Ni, criteria proposed by Long et al. (1995) having potential adverse effects on biota, but Ni and Cr are comparable to the high background values found in surrounding areas. These conditions have likely led to the development assemblages which are highly tolerant and largely adapted to natural and anthropic disturbance

events, such as those which often happens in transitional waters (Ponti et al., 2008). Dredging activities may dramatically affect benthic assemblages by altering species composition, diversity and ecological functioning, however, only a few studies have reported the reduction of species richness and biomass, or effects on secondary production (Quigley and Hall, 1999; Ponti et al., 2007). The consequence of sediment dredging are likely to affect for hundreds of metres the shallow areas surrounding se disturbed sites, according to the extent of the intervention (Quigley and Hall, 1999; Skilleter et al., 2006; Wilber et al., 2007). The dredging operations monitored in this study involved one of the main channels that cross the Pialassa

Fig. 3. (a) Mean abundance (SE) of Streblospio shrubsolii and (b) Corophium insidiosum, (c) species richness (S  SE) and (d) Shannon’s index values (SE) at each site located in the putatively impacted (filled symbols) and in the controls sites (open symbols) in the channels before and after the dredging.

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Table 2 Summary of asymmetrical ANOVA on the abundance of Streblospio shrubsolii, Corophium insidiosum, on species richness and the Shannon’s diversity index comparing each disturbed site (I) with control sites (Cs) in the channels. Source

T  Cs ResCs T  IBAC1 vs. Cs Appr. Den.a T  IBAC3 vs. Cs Appr. Den.a T  IBAC5 vs. Cs Appr. Den.a Transformation a b

Streblospio shrubsolii df

MS

2 12 1 16 1 16 1 16

0.9313 1.0413 16.4653 0.8396 44.9928 1.2685 0.1575 0.7844 Ln(x þ 1)

F

Corophium insidiosum P

df

MS

0.8900

0.4344

19.6109

0.0004

35.4693

0.0000

0.2008

0.6601

2 12 1 16 1 16 1 16

1.0188 0.2682 0.1212 0.2428 0.2446 0.2845 6.2516 0.2022 x^025

F

H0 (Log2)

S P

df

MS

F

P

df

MS

F

P

3.8000

0.0527

0.1969 0.2150

0.1068

0.7481

0.8598

0.3676

0.0731

0.7904

5.0710

0.0387

30.9179

0.0000

5.5846

0.0311

0.4168 0.2691 0.0272 0.2548 1.2069 0.2380 2.0405 0.2308 None

0.2521

1.6667

2 12 1 16 1 16 1 16

1.5500

0.4900

6.2222 3.3333 5.5556 3.3333 0.2223 3.0417 16.0556 2.8750 Noneb

1.8700

0.4992

2 12 1 16 1 16 1 16

8.8410

0.0090

When T  Cs is not significant the appropriate denominator is ResI þ Cs where I is the corresponding putative impacted sites, otherwise is T  Cs. Cochran’s Test P < 0.01.

Baiona lagoonal system and the intervention lasted for nearly 1 year. Therefore it should not be considered as a single and univocal disturbance event, either in space or time. Accordingly each studied impact site, in the channels and in the surrounding shallow ponds, was considered individually and compared with a set of control sites, sampled before and after the dredging. Despite the extent of the intervention, effects of the dredging on macrobenthic assemblages were detected only within the dredged channel, while in the surrounding ponds no clear and unequivocal effects were found (Table 4). In particular the dredging could have promoted the increase in the abundance of the polychaete Streblospio shrubsolii in the southern and central part of the dredged channel (BAC1 and BAC3) and the increase of the amphipod Corophium insidiosum in the northern side (BAC5), compared to the controls. Instead, species diversity was reduced in the central and northern part of the dredged channel (BAC3 and BAC5) in comparison to the controls. These effects on the macrobenthic invertebrate assemblages could be related to the observed changes in sediment catachrestics,

contamination and toxicity, which are summarised in Table 4 (for details see Guerra et al., 2009). Both S. shrubsolii and C. insidiosum could take advantage of reduced competition and new space availability for larval settlement and recruitment after the dredging activities. Despite of a general increasing trend in fine-grained sediment after the dredging, both at impacted and at control sites, a significant additional increase in silt þ clay (<63 mm) sediment was detected in the central part of the dredged channel, while a significant reduction of medium and coarse sand (>250 mm) was observed in the northern site (Table 4). These changes give evidence of bottom alteration inside the dredged channel. Sediment organic matter, measured as loss of weight on ignition (LOI), significantly increased in the southern part of the dredged channel (BAC1) and decreased in the northern side (BAC5), indicating an opposite effect of dredging along the channel may be due to the different time passed since intervention, which started from the north side. Streblospio shrubsolii can be considered an opportunistic polychaete (Sarda´ and Martin, 1993; Lardicci et al., 1997)

Fig. 4. (a) Mean abundance (SE) of Capitella capitata and (b) Streblospio shrubsolii, (c) species richness (S  SE) and (d) Shannon’s index values (SE) at each site located in the putatively impacted (filled symbols) and in the controls sites (open symbols) in the ponds before and after the dredging.

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Table 3 Summary of asymmetrical ANOVA on the abundance of Streblospio shrubsolii, Capitella capitata, on species richness and the Shannon’s diversity index comparing each disturbed site (I) with control sites (Cs) in the ponds. Source

T  Cs ResCs T  IPOL1 vs. Cs Appr. Den.a T  IPOL3 vs. Cs Appr. Den.a T  IVEN5 vs. Cs Appr. Den.a Transformation a

Streblospio shrubsolii

Capitella capitata

H0 (Log2)

S

df

MS

F

P

df

MS

F

P

df

MS

F

P

df

MS

F

P

2 12 1 2 1 2 1 2

1.5080 0.1208 0.3354 1.5080 0.0938 1.5080 2.1111 1.5080 x^0.15

12.4800

0.0012

0.3872

0.3319

3.1616

0.2174

0.0622

0.8263

5.5670

0.0313

0.6530

0.5039

0.7355

0.4815

1.3999

0.3583

3.5943

0.0762

0.0645

0.8232

1.0619 0.2201 3.3573 1.0619 0.7810 1.0619 3.3960 1.0619 None

0.0290

1.6125

2 12 1 2 1 2 1 2

4.8200

0.8025

26.0556 6.6667 42.0139 26.0556 17.0139 26.0556 1.6806 26.0556 None

0.0493

0.0647

2 12 1 2 1 2 1 2

3.9100

0.6836

5.5264 5.3763 0.4172 6.4514 22.5715 4.0545 18.1614 5.0528 Sqrt(x þ 1)

1.0300

0.2224

2 12 1 16 1 16 1 16

3.1980

0.2156

When T  Cs is not significant the appropriate denominator is ResI þ Cs where I is the corresponding putative impacted sites, otherwise is T  Cs.

and the increase in organic matter observed in BAC1 could have favoured its proliferation. The reduction of organic matter in BAC5 could provide suitable conditions for the less tolerant amphipod Corophium insidiosum (Procaccini and Scipione, 1992; Reish, 1993; Kevrekidis, 2004; Prato et al., 2006). Moreover, sediments exposed after dredging were equally or less contaminated by trace metals than removed sediments (Table 4). This environmental improvement could have contributed to the facilitation of the larval and post-larval recruitment in the dredged sites. According to Guerra et al. (2009), in the shallow water ponds connected to the dredged channel the increased thickness of the apparent Redox potential discontinuity (RPD) after the dredging suggests more oxygenated sediments due to enhanced water circulation (Table 4). While sediment toxicity, measured by MicrotoxÒ bioassays, increased at all controls sites, it remained constant at the pond impacted sites (Table 4) and this could be

considered as a positive effect of dredging, possibly related to the increased oxygenation of bottom sediments. In spite of this slight improvement in the environment conditions, negligible effects of the dredging were detectable in the surrounding shallow water ponds. The difficulty in detecting significant variation in assemblages and in species abundance could be largely attributed to the high heterogeneity of the shallow water habitats, especially in this physiographically complex lagoonal system. Spatial and temporal variability, especially in shallow ponds, can hide most of the direct and/or indirect effects related to the dredging activities. Moreover, the ecological effects of the disturbance could differ among sites according to typology, intensity and the duration of the intervention. Overall, direct effects of the dredging activities on the benthic assemblages in the channels have been detected, while less clear is the pattern of possible indirect effects on the assemblages in surrounding shallow ponds.

Table 4 Summary of ANOVA test including results on sediment characteristics (LOI: loss of weight on ignition; RPD: apparent Redox potential discontinuity), contamination, and toxicity tests from Guerra et al. (2009) (ns: not significant, *: P < 0.05; **: P < 0.01; ***: P < 0.001; upward arrow: increasing trend; downward arrow: decreasing trend; horizontal arrow: constant trend). Channels T  Cs

Taxa Tubificid oligochaetes Capitella capitata Streblospio shrubsolii Musculista senhousia Hydrobia ventrosa Corophium insidiosum Gammarus aequicauda Idotea balthica Chironomus salinarius Species diversity S H0 (log2) Sediment characteristics Silt þ Clay (<63 mm) Sand (>250 mm) LOI RPD Sediment contamination Hg Cu Ni Cr Pb Sediment toxicity tests Microtox Corophium insidiosum

Ponds T  I vs. Cs

T  Cs

BAC1

BAC3

BAC5

ns ns ns

ns ns ***[

ns ns ***[

ns ns ns

ns

ns

ns

***[

ns ns

ns ns

ns *Y

ns ns ns ns

ns ns **[ ns

ns ns ns ns * ns *

T  I vs. Cs POL1

POL3

VEN5

*** ns ** ns ** *** ** *** **

ns ns ns ns ns ns ns ns ns

ns */ ns ns ns ns ns ns ns

ns ns ns ns ns ns ns ns ns

ns **Y

* *

ns ns

ns ns

ns ns

*[ ns ns ns

ns ***Y ***Y ns

ns ** *** ns

ns ns ns *[

ns ns ns ***[

ns ns ns ns

*Y ns ns *Y ns

ns *Y ns ns ns

ns ns ns ns ns

*** *** ns ns ns

ns ns ns ns ns

ns ns ns ns ns

ns ns ns ns ns

ns ns

ns ns

ns ns

ns **

***/ ns

*/ ns

*/ ns

150

M. Ponti et al. / Estuarine, Coastal and Shelf Science 85 (2009) 143–150

Acknowledgements This study was undertaken within the research protocol of the Centre for Environmental Sciences of the University of Bologna (CIRSA), Ravenna Municipality, Local Health Agency (AUSL) and Environmental Protection Agency (ARPA-EMR). The work was supported by the strategic projects FINqUER funded by the University of Bologna. The authors would like to thank Giacomo Vasi, for his help with benthic sample processing, and two anonymous referees for their useful comments, which have improved the paper.

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