Macrozoobenthos monitoring in relation to dredged sediment disposal: The case of the Marano and Grado Lagoon (northern Adriatic Sea, Italy)

Regional Studies in Marine Science 33 (2020) 100916

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Macrozoobenthos monitoring in relation to dredged sediment disposal: The case of the Marano and Grado Lagoon (northern Adriatic Sea, Italy) Nicola Bettoso, Ida Floriana Aleffi, Lisa Faresi, Alessandro D’Aietti, Alessandro Acquavita

∗

Agenzia Regionale per la Protezione dell’Ambiente del Friuli Venezia Giulia (ARPA FVG), Via Cairoli 14, 33057, Palmanova, Italy

article

info

Article history: Received 1 February 2019 Received in revised form 28 October 2019 Accepted 28 October 2019 Available online 31 October 2019 Keywords: Sediment Environmental management Mediterranean lagoon WFD Ecological quality status

a b s t r a c t Coastal lagoons often require direct human intervention, such as dredging and the disposal of dredged material, to maintain the safety and accessibility of navigation channels. Dredging may have short and long-term effects on macrozoobenthic communities. Because they respond relatively quickly to anthropic and natural stress, benthic invertebrates are considered a biological quality element sensu Water Framework Directive 2000/60/CE (WFD) used to assess the ecological status in transitional water such as coastal lagoons. The monitoring of dredged areas and the subsequent comparison to the ecological quality status (EcoQS) sensu WFD in the Marano and Grado Lagoon has revealed that macrozoobenthic communities are slightly unbalanced and tolerant species prevailed. In this naturally stressed environment, the recovery time after the disturbance caused by dredging was three to six months, and richness and diversity were mostly related to the salinity range and distance from sea inlets. © 2019 Elsevier B.V. All rights reserved.

1. Introduction Macrozoobenthos is one of the Biological Quality Elements (BQEs) considered in the application of the Water Framework Directive (WFD; EU, 2000) as it has the ability to rapidly respond to natural and anthropic stress (Pearson and Rosenberg, 1978; Dauer, 1993; Borja et al., 2000). Coastal lagoons require direct human intervention to prevent the deposition of silt particles and to maintain the accessibility/safety of navigation channels. This is done mainly by dredging and the disposal of dredged material in well-defined areas, and represents one of the concerns in coastal zone management (Bolam et al., 2006). Dredging may cause several short-term effects on benthic assemblages (Quigley and Hall, 1999). Among these one must consider (1) the reduction of richness and abundance, which is directly related to the disturbance event; (2) the change in sediment properties (e.g., grain size; Nayar et al., 2007) modifying relevant habitat features; (3) the re-suspension of fine sediment with related porewaters and associated nutrients, organic matter (OM) and pollutants (Eggleton and Thomas, 2004), which can lead to eutrophication, hypoxic events and increasing toxicity (Vezzone et al., 2018); (4) the loss of habitat and a reduction in ∗ Correspondence to: Agenzia Regionale per la Protezione dell’Ambiente del Friuli Venezia Giulia, Osservatorio Alto Adriatico, Via A. La Marmora 13, 34139 Trieste, Italy. E-mail address: [email protected] (A. Acquavita). https://doi.org/10.1016/j.rsma.2019.100916 2352-4855/© 2019 Elsevier B.V. All rights reserved.

the functioning of the ecosystem. In addition, long-term effects on macrobenthic assemblages are also expected due to increased water circulation and oxygenation, the enhancing of biogeochemical processes (e.g., OM mineralisation) and new available space for larval settlement and recruitment. The spatial and temporal impact of dredging activity on benthic communities has been investigated in several marine environments (Harvey et al., 1998; Bolam and Whomersley, 2003, 2005; Fredette and French, 2004; Bolam et al., 2006; Powilleit et al., 2006; Bolam, 2012), but there are few studies focused on Mediterranean lagoons (Munari and Mistri, 2014; Ponti et al., 2009). The Marano and Grado Lagoon belongs to the extended transitional lagoons network of the northern Adriatic Sea (Italy) (Fig. 1). It is classified as a coastal microtidal lagoon of large dimensions (Italian Decree n. 131/08) and has been protected by the Ramsar Convention since 1971. Following the implementation of the Habitats Directive (92/43/EC) the lagoon was also designated as a Site of Community Importance (SCI — IT3320037). Besides its natural significance, the lagoon and the neighbouring mainland host several socio-economic activities (e.g., industrial sites, commercial harbours, marinas, fishing, fish and clam farming, and tourism Sladonja et al., 2011). The lagoon displays some morphological features originating from a sediment deficit such as the flattening and deepening of tidal flats, an evident reduction in the size of saltmarshes and channels silting up. Thus, there is a

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N. Bettoso, I.F. Aleffi, L. Faresi et al. / Regional Studies in Marine Science 33 (2020) 100916

Fig. 1. The Marano and Grado Lagoon (northern Adriatic Sea) and its main waterways (black lines) and rivers (blue lines). (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

pressing need for effective conservation and management (Bezzi et al., 2015). The lagoon is crossed by several waterways constructed in the past and still in use today. The main channels are the historical ‘‘Litoranea Veneta’’ (1915–1916) and the Aussa-Mare (1970s). The Aussa-Mare waterway is used by mercantile ships to cross from the Porto Buso inlet to Porto Nogaro, an industrial harbour located in the inland Aussa-Corno River. In addition, minor channels aiding navigability, large marinas and associated shipways, especially in the western part of the lagoon (i.e., Lignano Sabbiadoro and Aprilia Marittima) are present (Fontolan et al., 2012). These waterways are not restored regularly but on an as-needed basis. The regional administration adopted a temporary plan to store the dredged sediment in disposal areas such as tidal flats, adjacent to each dredging area and belonging to the same water body and there is currently a debate regarding whether or not to use the dredged sediment to restore salt marshes (Bezzi et al., 2015). The monitoring plan adopted used a BACI approach (Before/ After and Control/Impact; Underwood, 1992), which was successfully applied to dredging in harbours (Guerra-García et al., 2003), estuaries (van der Wal et al., 2011), and lagoons (Ponti et al., 2009; Munari and Mistri, 2014; Piló et al., 2019). According to this method, a well-defined set of physico-chemical and chemical parameters of sediment are measured before and after the work, in both dredging and disposal areas: the content of pollutants (e.g. heavy metals and persistent organic pollutants), ecotoxicological assays and sediment texture are evaluated to check the sediment compatibility between sites before dredging. After this preliminary assessment, there is an ad hoc monitoring program conducted during each of the procedures for every disposal area to ensure there will be no negative impact on the environment. From 2013 to 2015, nine dredging activities in six bodies of water took place in the Marano and Grado Lagoon and involved the relocation of 556,200 m3 of sediment. In this paper

we evaluate the effects of dredging on macrobenthic communities (pre- and post-dredging) by assessing the ecological quality status (EcoQS) via an M-AMBI index (Multivariate-AMBI — http://ambi. azti.es) (Muxika et al., 2007) and specific statistical nonparametric tests, comparing the results with those found via WFD monitoring of the same water bodies. 2. Material and methods 2.1. Study area and field sampling under WFD and dredging operations The Marano and Grado Lagoon is located between the Tagliamento and Isonzo River deltas and extends for approximately 32 km, reaching a width of up to 5 km (A = 160 km2 ). The main freshwater inputs are from small rivers (especially the western sector) and drainage pumps located in the low Friulian plain. The sediment loads, in the form of silty and clayey particles, are mainly due to Isonzo and Tagliamento River inputs and to the erosion of barrier islands (sand materials; Fontolan et al., 2012). Salinity has a wide range moving from river mouths to tidal inlets (from 1 to 33) and the water temperature demonstrates a clear seasonal trend (3.2 to 30.1 ◦ C). A great quantity of anthropogenic nutrients (mainly nitrogen as N-NO3 ) are carried into the lagoon area via river discharges, especially in the western sector, but these inputs are offset by the water exchange with the adjacent open sea, which exerts a dynamic dilution effect (Acquavita et al., 2015). The implementation of the WFD was introduced into Italian law by Legislative Decree no 152/2006, and subsequently by decrees for the characterisation of water body types (WBs; DM 131/2008), monitoring (DM 56/2009) and classification (DM 260/2010) of WBs. The classification is supported by the analysis

N. Bettoso, I.F. Aleffi, L. Faresi et al. / Regional Studies in Marine Science 33 (2020) 100916

3

Fig. 2. Index map of the study area with the monitored Water Bodies sensu Water Framework Directive (sampling stations depicted as grey circles) and the areas subjected to dredging activities. TME (1–4 mesohaline WBs, TPO1-5 polyhaline WBs, TEU1-4 euhaline WBs, FM2-4 are heavily modified WBs subjected to substantial changes due to anthropogenic activities (e.g., significant hydraulic regime changes, ex fish farms). The arrows indicate the main sea inlets. TME: type mesohaline; TPO: type polyhaline; TEU: type euhaline; FM: heavily modified; WBs: water bodies.

of some specific descriptors such as geographic location, geomorphology, amplitude of tides, and surface salinity. In the Marano and Grado Lagoon, water types were defined as a function of salinity: mesohaline (TME) with salinity ranging from 5 to 20, polyhaline (TPO; salinity range = 20 − 30) and euhaline (TEU; salinity range = 30 − 40); 17 WBs were identified in all (Fig. 2). In order to conduct WFD monitoring, a total of 23 sampling sites were selected for benthos collection in spring 2014; this was conducted by means of a 0.047 m2 van Veen grab and four replicates were collected from each site. The collected sediment was sieved (1 mm mesh) and fixed in a 4% buffered formaldehyde solution stained with Bengal Rose immediately after retrieval. Subsequently the fauna was separated, counted and identified to the lowest possible taxonomic level. Dredging and disposal operations were conducted in five areas from 2013 to 2015 (A–E; Table 1, Figs. 2–3). Dredged sediments were pumped through a pipeline into the disposal area to prevent an excess of turbidity; every disposal area was surrounded by silt curtains. In dredging zones, additional macrozoobenthos sample monitoring was carried out inside disposal areas to test the impact (I) and in adjacent stations as a control (C). The control was monitored before (B) starting operations and from 3 to 6 months after the end of disposal (A) depending on the volume of the sediment dredged in every area (Table 1). For example, in Area C, where approximately 240,000 m3 of sediment were dredged, an additional sampling after one year and two control stations were required. On the other hand, in spite of the mandatory BACI approach, a degree of flexibility is permitted depending on the characteristics of the site: the low depth of the disposal site in Area B did not permit sampling inside the impact area (I), thus the samples were collected from two adjacent control stations (Fig. 3). Finally, during the first dredging (51,000 m3 ), sampling before operations was not carried out in Area D due to logistical problems. Sampling and analysis for macrozoobenthos (41 samples) were conducted with the same methodology adopted for WFD monitoring.

Fig. 3. Sampling stations for dredging monitoring (dark and dotted arrows indicate dredging and disposal areas, respectively).

Grain size classification was determined according to Nota’s scale (Nota, 1958) and adapted for the Adriatic Sea by Brambati et al. (1983). In detail, it is based on the sand weight percentage (Sand%) as follows: sand (Sand% > 95), pelitic sand (70 < Sand% < 95), very sandy pelite (30 < Sand% < 70), sandy pelite (5 < Sand% < 30) and pelite (Sand% < 5).

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Table 1 Dredging areas with the related volume of dredged material (per year and area), sampling station code (dredging area-water body-Impact or Control) and scheduled operations conducted. Volume of dredged sediment (m3 )

Dredging area

Year

A

2013 2014 2015

3,500 5,000 5,000

Sampling stations

Sampling before dredging

Sampling after end dredging

A-FM3-I, A-FM3-C A-FM3-I, A-FM3-C A-FM3-I, A-FM3-C

12 June 2013 12 June 2014 27 May 2015

14 October 2013 (90 dd) 10 November 2014 (120 dd) 07 December 2015 (180 dd)

B

2014 2015

10,500 80,000

B-TME1-C1, B-TPO3-C1 B-TME1-C1, B-TPO3-C2

22 April 2014 06 March 2015

09 September 2014 (90 dd) 23 December 2015 (120 dd)

C

2014

240,000

C-TME4-I, C-TME4-C, C-TPO5-C

26 June 2014

05 May 2015 (180 dd) 27 October 2015 (365 dd)

D

2015 2015

51,000 120,000

D-TPO4-C1, D-TPO5-I1 D-TPO4-C1, D-TPO5-I2

08 July 2015

05 May 2015 (180 dd) 19 May 2016 (180 dd)

E

2015

41,200

E-TPO5-I, E-TPO5-C, E-TPO503-C

29 October 2015

07 March 2016 (120 dd)

2.2. Application of ecological indices and data analyses The ecological quality status (EcoQS) was assessed by means of the M-AMBI index (Multivariate-AMBI — http://ambi.azti.es) (Muxika et al., 2007), which is derived from the calculation of the proportion sensitive/tolerant species measured by AMBI (Borja et al., 2000), by the Shannon–Wiener diversity index (H′ ) (Shannon and Weaver, 1949) and by the richness expressed as the number of taxa (S). AMBI is based on the classification of the benthic species in five (I–V) ecological groups (EG), according to their tolerance to pollution (EG-I = species which are very sensitive to organic enrichment, intolerant to pollution; EG-II = species which are indifferent to enrichment; EG-III = species which are tolerant to enrichment, slightly unbalanced environments; EG-IV = second-order opportunistic species, slight to pronounced unbalanced environments; EG-V = first-order opportunistic species, pronounced unbalanced environment), then applying a formula to calculate the AMBI on a scale of increasing disturbance (for further details see Borja et al., 2000, 2003). Italian Decree n. 260/2010 reports the official reference conditions for High status as follows: AMBI = 2.14, H′ = 3.40, S = 28 for poly/mesohaline WBs and respectively 0.63, 4.23, 46 for euhaline WBs. Ecological Quality Ratios (EQR) boundaries between ecological status classes were defined by the same Decree by the following values of MAMBI: High/Good = 0.96, Good/Moderate = 0.71, Moderate/Poor = 0.57 and Poor/Bad = 0.46. The Wilcoxon two-sample paired test was applied to the parameters in order to verify the null hypothesis that the median of the paired differences is 0; in our case the paired samples are related to the comparison of stations B (before) to A (after) and C (control) to I (impact). In accordance with the BACI sampling design, the assemblage multivariate pattern was tested by PERMANOVA after a fourth root transformation and using the Bray– Curtis similarity. Finally the data from all stations in every area were averaged to undertake K-dominance curves and SIMPER analysis, in order to evaluate the similarities and/or differences among dredging areas and WFD water types. Calculations were performed using the PRIMER v7 + PERMANOVA software package developed at the Plymouth Marine Laboratory. 3. Results A total of 132 taxa from 9612 individuals were identified from the dredging monitoring sites, and 125 taxa from 8675 individuals were identified from the WFD samples. Annelida Polychaeta was the dominant taxon in terms of species and abundance in both monitoring programs, followed by molluscs and crustaceans in WFD and dredging samples, respectively, while echinoderms were scarcely represented. The remaining taxa (Porifera, Ascidiacea, Nemertea, Phoronida and Chironomidae as Insecta) were grouped into Other (Fig. 4). Grain-size was sandy pelite (5 <

Fig. 4. Taxa and abundance in dredging and WFD monitoring.

Sand% < 30) and no significant difference was found among the sites investigated (Table 2). The values of richness (S) and diversity (H′ ) were lower after the disposal of dredged material in the impact sites, although the Wilcoxon (W) two sample paired test indicated a significant difference only for richness (S) in the before (B) compared to after (A). The AMBI showed that benthic communities were always unbalanced regardless of dredging. As a consequence, the EcoQS sensu M-AMBI was Moderate in both C and I stations, and decreased from Good to Moderate when comparing B to A stations, although the W test did not show any significant differences (Table 2). In Table 3, the results of PERMANOVA on the benthic community of dredging areas are shown. The two main factors (Time B–A and Site C–I) were not found to be significant, while no significant Time x Site interaction was detected either. The mean values of macrozoobenthic indices and EcoQS for dredging areas compared to WFD sites grouped on the basis of water types are reported in Table 4. EcoQS were Good in Areas B, D and E as well as in euhaline and polyhaline water types. Moderate conditions were recorded in Areas A, C and in mesohaline water types. AMBI confirmed that benthic communities were

N. Bettoso, I.F. Aleffi, L. Faresi et al. / Regional Studies in Marine Science 33 (2020) 100916 Table 2 Average values ± s.d. of Sand (%), richness (S), Shannon–Wiener diversity index (H’), AMBI and M-AMBI index following BACI approach: Before (B) Vs. After (A) and Control (C) Vs. Impact (I). (W Wilcoxon two sample paired test: n.s. not significant; * significant). n %Sand S H′ AMBI M-AMBI

B

A

21 23.8 ± 11 22 ± 11 2.89 ± 1.1 2.79 ± 0.37 0.74 ± 0.15

21 26 ± 11.3 17 ± 6 2.57 ± 0.8 2.93 ± 0.72 0.66 ± 0.16

W

C

I

W

ns * ns ns ns

19 29.6 ± 11 20 ± 10 2.75 ± 0.94 3.02 ± 0.59 0.69 ± 0.18

19 23.4 ± 13.7 18 ± 8 2.69 ± 1.02 2.86 ± 0.63 0.67 ± 0.11

ns ns ns ns ns

Table 5 Cumulative percentage of dominant taxa in water types (eu-, poly- and mesohaline) and dredging areas (A–E) revealed by SIMPER analysis. Euhaline Heteromastus filiformis Abra segmentum Cerastoderma glaucum Pseudoleiocapitella fauveli Leiochone leiopygos Pseudopolydora antennata Nephtys hombergii Notomastus latericeus Leptopentacta elongata Paranthura japonica Loripes lacteus Prionospio fallax Prionospio cirrifera

Table 3 Summary of PERMANOVA on macrozoobenthos assemblages of dredging areas (Bray–Curtis similarity on fourth root transformed data) comparing time (Before Vs. After) and site (Control Vs. Impact). Source

df

Pseudo-F

P (MC)

Time (B–A) Site (C–I) Time x site Residuals

1 1 1 37

1.58 0.77 0.8

0.109 0.644 0.611

Total

40

5

Polyhaline Abra segmentum Heteromastus filiformis Cerastoderma glaucum Leiochone leiopygos Notomastus latericeus Cirriformia tentaculata Paranthura japonica Nephtys hombergii

always unbalanced both in the dredged and WFD sites. On the basis of S and H′ indices, values comparable to polyhaline WBs were found in Area A, whereas in Area C values were similar to a mesohaline condition. The trends of abundance are displayed by the K dominance curves reported in Fig. 5. The steepest and most elevated curve of mesohaline WBs indicated the lowest diversity, Area A fits quite well with the euhaline curve as well as Area D with polyhaline. The curve for Area B was found to be intermediate between poly- and euhaline conditions, instead Area E fit between those meso- and polyhaline. Finally, the curve for C was similar to the mesohaline curve. Species contributing to the average similarity of samples within each group are reported in Table 5 (SIMPER): dominance increased from the euhaline group to Area C according to K dominance curves. Among species, the dominance of the bivalve Abra segmentum in the Lagoon was noteworthy as well as those of the polychaetes Hediste diversicolor and Streblospio shrubsolii in mesohaline WBs.

Mesohaline Abra segmentum Hediste diversicolor Streblospio shrubsolii Gammarus aequicauda

contrib. % 16,21 11,16 10,71 7,93 7,86 5,19 4,44 4,37 3,15 3,04 2,35 2,32 2,24 80.99

28,98 24,95 6,98 5,91 5,35 3,12 2,75 2,61 80.65

Area A Abra segmentum Leiochone leiopygos Gammarus aequicauda Heteromastus filiformis Capitella capitata Chironomidae Cirrophorus furcatus Nephtys hombergii Cerastoderma glaucum Platynereis dumerilii Notomastus latericeus

contrib. % 17,66 17,22 14,15 9,32 4,34 4,06 3,77 3,11 2,92 2,32 1,78 80.64

Area B Abra segmentum Chironomidae Cirriformia tentaculata Gammarus aequicauda Cerastoderma glaucum Streblospio shrubsolii Heteromastus filiformis Paranthura japonica Aphelochaeta filiformis

18,08 16,58 12,68 8,54 6,88 5,87 5,59 4,70 4,18 83.09

Area D Abra segmentum Leiochone leiopygos Heteromastus filiformis Cerastoderma glaucum Paranthura japonica Notomastus latericeus

24,76 19,55 18,26 7,28 5,60 5,46 80.91

Area E Abra segmentum Hediste diversicolor Paranthura japonica Streblospio shrubsolii Heteromastus filiformis 32,02 27,27 18,15 8,59 86.04

42,29 16,88 9,24 8,28 6,04 82.73

Area C Abra segmentum Hediste diversicolor Corophium orientale

30,43 28,93 25,33 84.68

4. Discussion The soft bottom macrozoobenthic communities of the Marano and Grado Lagoon are dominated by Annelida Polychaeta, followed by Mollusca and Crustacea. The percentage of abundance and richness of each main taxa found in dredging and WFD monitoring were comparable to those recorded during the first WFD survey conducted in 2008 (Bettoso et al., 2010). Sediment structure was similar before and after disposal and both in the control and impact sites. The sediment texture seen in this study was mostly sandy pelite, as reported in Acquavita et al. (2014), with coarser fractions in the proximity of the tidal inlets (euhaline water types), while finer sediments were found in the inner sectors of the lagoon (mesohaline water types) as a consequence of

the prevailing tidal sediment transport and of the modest freshwater contribution from rivers, which provide particles enriched by silts and clays (Acquavita et al., 2012). Dredging operations are expected to reduce species diversity by 30–70% and the number of individuals by 40–95% (Newell et al., 1998). The recovery time for macrobenthic communities after disturbance depends on the spatial scale, hydrodynamic conditions, bottom grain size and the structure of the community affected (Kaiser and Spencer, 1996; Pranovi et al., 1998). For example, in the Ceuta Harbour (North Africa), natural recolonisation processes proceeded rapidly after dredging, and about 6 months were required for the disturbed

Table 4 Mean values ± s.d. of macrozoobenthic indices and resulting Ecological Quality Status (EcoQS) in dredging areas and water types sensu WFD Directive. Dredging areas

n. S H′ AMBI M-AMBI EcoQS

WFD water types

A

B

C

D

E

Euhaline

Polyhaline

Mesohaline

12 25 ± 10 3.2 ± 1.1 2.7 ± 0.9 0.63 ± 0.17 Moderate

8 20 ± 5 3.1 ± 0.6 2.8 ± 0.5 0.83 ± 0.10 Good

9 11 ± 3 1.9 ± 0.8 2.9 ± 0.5 0.62 ± 0.11 Moderate

6 25 ± 9 3.2 ± 0.4 3.2 ± 0.6 0.84 ± 0.13 Good

6 18 ± 3 3 ± 0.3 3.1 ± 0.2 0.76 ± 0.06 Good

8 35 ± 10 3.7 ± 0.8 2.9 ± 1.1 0.73 ± 0.14 Good

10 24 ± 9 2.8 ± 0.6 2.9 ± 0.7 0.82 ± 0.15 Good

5 12 ± 6 2.2 ± 0.5 2.7 ± 0.8 0.68 ± 0.04 Moderate

6

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Fig. 5. K-dominance curves for water types and dredging areas.

area to re-establish a macrobenthic community structure similar to that of the control area (Guerra-García et al., 2003). In a Mediterranean coastal lagoon, macrobenthic communities were resilient, recovering within months rather than years (Munari and Mistri, 2014) because in this ecosystem the magnitude and frequency of natural perturbations are high (Bolam and Rees, 2003; Fredette and French, 2004). Similarly, soft bottom benthic communities found in the Marano and Grado Lagoon are well-adapted to perturbations such as salinity, temperature changes, oxygen deficiency, sediment resuspension and fishing activity (Cognetti and Maltagliati, 2000). These peculiarities of the Mediterranean lagoon macrozoobenthos were previously reported by Pérès and Picard (1964), who defined the Eurythermal and Euryhaline Lagoon biocenosis (LEE). The LEE biocenosis is typical of unstable environments where wide salinity and temperature variations occur yearly due to river floods, rainfalls and a high rate of evaporation in the summer (Bettoso et al., 2010). It is therefore not surprising that despite dredging and disposal, AMBI indicated an unbalanced benthic community health for the whole lagoon, where the EG-III ecological group prevailed (Borja et al., 2003), represented by tolerant species. This explains the relatively rapid recolonisation, the comparable M-AMBI values and the similar community structure sensu PERMANOVA, before dredging and three to six months after the work was completed, both in impact and control sites. In every dredging area, the macrozoobenthos appears to exhibit characteristics similar to that belonging to the same water type on the basis of salinity range and distance from sea inlets according to Bettoso et al. (2010): Areas B, D and E belong to polyhaline WBs and a Good EcoQS was found. Conversely, Area A is euhaline but showed a Moderate EcoQS as it is located in a heavily modified WB (FM3) where a strong modification of the hydrological regime occurred. Finally, Area C had Moderate status because it is situated in the mesohaline belt, far from the inlets, with Moderate to Poor EcoQS and low values of both S and H′ (Bettoso et al., 2010). The bivalve Abra segmentum, the nereid polychaete Hediste diversicolor and the spionid polychaete Streblospio shrubsolii represented at least 60% of the relative abundance in the inner zones, where typical marine species cannot survive, with the exception of those able to tolerate a wide range of chemical and physical parameters, such as polychaetes belonging to Capitellidae and Spionidae families (Bettoso et al., 2013). A. segmentum is typically euryhaline, frequent in oligohyperhaline waters and tolerates a wide range of salinity (from 3 to 41) (Marazanof, 1969; Kevrekidis, 2004) and is common and frequently abundant in Mediterranean coastal lagoons, where it

can be dominant in terms of abundance and biomass (Kevrekidis and Kasapis, 2009). Nevertheless, this species was not among the dominant species in other smaller Adriatic lagoons where similar studies were performed (Ponti et al., 2009; Munari and Mistri, 2014). H. diversicolor, which is widely distributed in Mediterranean lagoons (Bazaïri et al., 2003; Nicolaidou et al., 2005), is a typical inhabitant of European brackish water, where it preferentially lives in muddy sediments, showing a high tolerance to levels of contaminants (Volpi et al., 1999) and salinity (it prefers low salinity values; Guerzoni and Tagliapietra, 2006). In addition, this species is able to rapidly inhabit a defaunated area after dredging by active adult immigration (Bonsdorff, 1980, 1983). S. shrubsolii is a typical lagoon species also widely distributed in other Mediterranean lagoons (Mistri et al., 2002; Rossi and Lardicci, 2002; Bazaïri et al., 2003; Dauer et al., 2003) and it was among the first colonisers of the newly deposited sediments from the dredging activity in Sacca di Goro, Po Delta (Munari and Mistri, 2014) ‘‘most likely’’ taking advantage of reduced competition and new space availability for larval settlement and recruitment after dredging (Ponti et al., 2009). Noteworthy was the presence, in both WFD and dredging monitoring, of the non-indigenous isopod Paranthura japonica being among the dominant species, which had previously been identified in the Venice Lagoon in 2012 by Marchini et al. (2014). The lagoon has a high proportion of ‘‘r-strategists’’ with high reproductive potential, mainly constituted by opportunistic polychaetes (e.g. Heteromastus filiformis) (Boesch, 1974; Dauer, 1984) which quickly colonise disturbed habitats (Boesch, 1977; Pearson and Rosenberg, 1978; Dauer, 1984). This condition is comparable to that of estuarine communities. Because such environments are subject to regular disturbance under natural conditions prior to dredging, the ecological succession recovers to the colonisation phase, but does not proceed to the development of K-selected slow-growing equilibrium species within the community (Newell et al., 1998). The recovery of the ‘‘normal’’ community in disturbed deposits such as the fine sediments of the Marano and Grado Lagoon can be achieved within a few months of the cessation of dredging. These results agree with the estuarine quality paradox concept (see Elliott and Quintino, 2007), as it is often not possible to distinguish between the response due to anthropogenic stress and the natural fluctuation of environmental parameters. Therefore it is a paradox to use macroinvertebrates as biomonitors to detect pollution or dredging impacts for estuarine and lagoon quality assessment, since most species are well adapted to these highly variable environments (Zettler et al., 2013). Nevertheless, an innovative approach comprising the beta

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diversity component (partitioned into turnover and nestedness) as an effective measure of ecosystem resilience should also be considered (Piló et al., 2019). 5. Conclusion Dredging activities were performed to maintain the safety and accessibility of navigation channels and marinas in the Marano and Grado Lagoon. Due to the urgency of the situation, a provisional plan was adopted to relocate the dredged sediments to disposal areas near every dredging zone. The monitoring of macrozoobenthos in disposal areas following the BACI approach and the comparison with the ecological status of this biological quality element (sensu WFD) in the whole lagoon can be summarised as follows:

− Benthic macroinvertebrate communities in the lagoon are slightly unbalanced and tolerant species prevailed, mostly polychaetes; − The soft bottom macrozoobenthos recovered 3–6 months after dredging; − The ecological quality status (EcoQS) of macrozoobenthic communities before disturbance and after recovery is mostly driven by salinity range and distance from sea inlets, regardless of dredging and disposal activities. Beyond the serious necessity of the work and the resilience properties of this lagoon environment, the possibility of using dredged material to restore salt marshes should be taken into consideration in order to better preserve the lagoon ecosystem. Declaration of competing interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgements The research activity was supported by all the staff of ARPA FVG, Italy involved in sampling operations. The authors are very grateful to Karry Close for her careful reading of the manuscript. Thanks to anonymous reviewers for their contribution to the improvement of manuscript. References Acquavita, A., Aleffi, I.F., Benci, C., Bettoso, N., Crevatin, E., Milani, L., Tamberlich, F., Toniatti, L., Barbieri, P., Licen, S., Mattassi, G., 2015. Annual characterization of the nutrients and trophic state in a Mediterranean coastal lagoon: The Marano and Grado Lagoon (northern Adriatic Sea). Reg. Stud. Mar. Sci. 2, 132–144. Acquavita, A., Covelli, S., Emili, A., Berto, D., Faganeli, J., Giani, M., Horvat, M., Koron, N., Rampazzo, F., 2012. Mercury in the sediments of the Marano and Grado Lagoon (northern Adriatic Sea): Sources, distribution and speciation. Estuar. Coast. Shelf Sci. 113, 20–31. Acquavita, A., Falomo, J., Predonzani, S., Tamberlich, F., Bettoso, N., Mattassi, G., 2014. The PAH level, distribution and composition in surface sediments from a Mediterranean Lagoon: The Marano and Grado Lagoon (northern Adriatic Sea, Italy). Mar. Pollut. Bull. 81, 234–241. Bazaïri, H., Bayed, A., Glémarec, M., Hily, C., 2003. Spatial organisation of macrozoobenthic communities in response to environmental factors in a coastal lagoon of the NW African coast (Merja Zerga, Morocco). Oceanol. Acta 26, 457–471. Bettoso, N., Acquavita, A., D’Aietti, A., Mattassi, G., 2013. The Marano and Grado Lagoon: a brief synopsis on the aquatic fauna and fisheries resources. Ann. Ser. Hist. Nat. 23 (2), 135–142. Bettoso, N., Aleffi, I.F., Faresi, L., Rossin, P., Mattassi, G., Crivellaro, P., 2010. Evaluation on the ecological status of the macrozoobenthic communities in the Marano and Grado Lagoon (northern Adriatic Sea). Ann. Ser. Hist. Nat. 20 (2), 193–206.

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