Tidal cycling of mercury and methylmercury between

Marine Chemistry 130-131 (2012) 1–11

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Tidal cycling of mercury and methylmercury between sediments and water column in the Venice Lagoon (Italy) S. Guédron a, b,⁎, L. Huguet a, D.A.L. Vignati c, B. Liu a, F. Gimbert a, d, B.J.D. Ferrari e, R. Zonta f, J. Dominik a a

Université de Genève – Institut F.-A. Forel, route de Suisse 10, CP416, CH-1290 Versoix, Switzerland ISTerre, Université Grenoble 1, IRD – UMR 5559 (IRD/UJF/CNRS) – BP 53, F-38041 Grenoble, France CNR-IRSA – UOS Brugherio, Via del Mulino 19, I-20861 Brugherio (MB), Italy d Université de Franche-Comté – Laboratoire chrono-environnement, UMR UFC/CNRS 6249 USC INRA, place Leclerc, F-25030 Besançon cedex, France e Cemagref, UR MALY, 3 bis quai Chauveau, F-69336 Lyon cedex 09, France f ISMAR-CNR, Arsenale Tesa 104, Castello 2737 F, I-30122 Venice, Italy b c

a r t i c l e

i n f o

Article history: Received 1 July 2011 Received in revised form 7 November 2011 Accepted 12 December 2011 Available online 24 December 2011 Keywords: Mercury Methylmercury Tidal flushing Sediments Venice Lagoon

a b s t r a c t The sediment of Venice Lagoon regularly undergoes complex redistribution due to tidal forcing, which affects the cycling of contaminants such as mercury (Hg) between the sediment and the water column. We examined the distribution of total Hg (THg) and monomethylmercury (MMHg) in the water column, sediment and pore-water at two sites: VE1 (located in a depositional area adjacent to salt marshes) and VE2 corresponding to a moderately erosive, open area. We obtained instantaneous (using cores and micro-needle samplers) and time-integrated (using peepers) concentrations of the two mercury species in both dissolved and particulate forms. THg and MMHg concentrations were higher in the sediments at site VE1 (621.9 ±213.7 ng g− 1 and 1.25 ± 0.63 ng g− 1 for THg and MMHg, respectively) than in those of the site VE2 (386.9± 92.7 ng g− 1 and 0.53 ± 0.30 ng g− 1). Hg concentrations in sediments were positively correlated with silts and organic matter content. Over two tidal cycles, the concentrations of THg and MMHg varied with the evolution of the tides. During the tidal flooding, both THg and MMHg peaked at the sediment–water interface and a moderate increase of dissolved MMHg was also observed in the water column. These fluctuations were observed during both tides and are suggestively related to advection of mercury species from surficial sediment pore-water to the water column and to desorption from suspended particles. The short-term increase in MMHg concentrations can result from in situ production, release from organic matter degradation, or from oxidative dissolution of redox-sensitive sulfide minerals and iron oxide reduction by micro-organisms; the two latter mechanisms being favored by redox oscillations in the surface sediment layers due to the tidal forcing. The decrease of both dissolved THg and MMHg concentrations at the sediment–water interface after high tide was attributed to a rapid adsorption onto particles. THg concentrations on suspended particles showed little variations during the tidal cycle with a minor peak at tide maximum, while MMHg concentrations on suspended particles slightly increased during ebb tide. MMHg concentrations on suspended particles were double than those in surface sediments, suggesting that tidal flushing may enhance dispersal of particle bound MMHg throughout the lagoon. © 2011 Elsevier B.V. All rights reserved.

1. Introduction The Venice Lagoon is a complex and dynamic system where notable exchanges of materials and energy between the main land (continent) and the Adriatic Sea occur (Ravera, 2000). The lagoon is shallow (average water depth 1.2 m) and consists of intertidal or submerged mudflats, salt marshes and man-made navigation canals (Molinaroli et al., 2009; Pranovi et al., 2004). The sedimentary dynamic strongly depends on the tidal flushing, leading to highly heterogeneous deposition and erosion rates between canals and salt-marshes areas. Sediment

⁎ Corresponding author. Tel.: + 33 4 76 63 59 28; fax: +33 4 76 63 52 52. E-mail address: [email protected] (S. Guédron). 0304-4203/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.marchem.2011.12.003

dynamics is also influenced by human activities related to construction, tourism and clam harvesting activities (Degetto et al., 2005). Venice Lagoon has been recognized as a mercury (Hg) contaminated area (Bloom et al., 2004a; Pavoni et al., 1992). The two chlor-alkali plants of the industrial area of Porto Marghera were identified as the main contributors to the contamination of the lagoon (Bellucci et al., 2002; Bloom et al., 2004b; Zonta et al., 2007). Until recently, antifouling paints used near the town of Chioggia (in the southern part of the lagoon) were another local source of mercury (Berto et al., 2006). Many authors described the Venice Lagoon as a ‘biological reactor’, because of considerable nutrient input, shallow and seasonally warm water, and extensive wetlands. Bloom et al. (2004b) described it as a “methyl Hg incubator”, where even small inputs of inorganic Hg could result in high levels of monomethylmercury (MMHg) production; the

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S. Guédron et al. / Marine Chemistry 130-131 (2012) 1–11

most toxic and bioaccumulating species of Hg (Cheng et al., 2009; Sirot et al., 2008; Wheatley and Wheatley, 2000). However, only few studies performed in the Venice Lagoon did consider the influence of tides on the production and release of MMHg from sediments to the water column. In a study conducted in the Canale SS Apostoli (City of Venice), Bloom et al. (2004b) showed that MMHg concentrations in unfiltered water varied only slightly during a tidal cycle, while total Hg (THg) was inversely correlated with tide height. In addition, the dynamics of total suspended solids (TSS) associated with tidal flushing was suggested as the vehicle for Hg dispersal from the sources to the whole of Venice's Lagoon (Berto et al., 2006) and may play an important role in the adsorption of dissolved Hg species produced in the surficial sediments (Bloom et al., 2004b). In the present study, we considered the influence of the tidal flushing on the short term THg and MMHg concentration changes in the water column and sediment pore water. First, we determined the THg and MMHg concentrations in sediments and pore water in two different areas typical of the sedimentary dynamic of the lagoon, (i.e., a remote site adjacent to salt marshes and open site close to the Burano canal bordering a marshy area). We also used various tools to get instantaneous and time integrated information on the dissolved and particulate THg and MMHg distribution in the water column and the sediment pore waters. Second, we followed temporal variations in particulate and dissolved THg and MMHg concentrations in water column and at the sediment–water interface during two tidal cycles to evaluate the exchanges between the surficial sediments and the water column for both Hg species. 2. Sampling sites, material and methods 2.1. Environmental settings and sampling strategy Two sites located in the northern part of the Venice Lagoon were sampled during two campaigns: October 30th, 2008 to November 4th, 2008 and September 8th to 18th, 2009. These sites were chosen because they were located in a fairly Hg contaminated part of the lagoon (Bloom et al., 2004b; Zonta et al., 2007). The first site (VE1 — 45°30′04.56″N, 12°25′04.31″E) was representative of a low-energy, depositional area and was in a subtidal area adjacent to the salt marshes around Torcello Island, approximately 5 km from the main land (Fig. 1). The second site (VE2 — 45°28′31.41″N, 12°25′50.56″E), representative of a dynamic

and moderately erosive area, was located close to the Burano canal at the border of a salt marsh area and strongly influenced by the incoming tidal currents from the Lido inlet (Fig. 1). In the 2008 prospective campaign, two cores were sampled at VE1 (i.e., VE 1.1 2008 and VE 1.2 2008 on September 30th at high tide) and VE2 (i.e., VE 2.1 2008 and VE 2.2 2008 on October 2nd at low tide). In 2009, the first part of the sampling was dedicated to the physicochemical characterization of sediments, pore-water and overlying water (water overlying sediments in cores) at the two sites (i.e., cores VE 1.3, VE 1.4 on September 8th at high tide and VE 2.3, VE 2.4 on September 11th at low tide and VE 2.5 September 9th at high tide). The second part of sampling was focussed on the dissolved and particulate Hg species dynamics at site VE1 during two tidal cycles on September 17th (tide 1, T0= 8:30 a.m.) and 18th (tide 2, T0= 8:30 a.m.) with a time resolution of one hour. 2.2. Sample collection 2.2.1. Sediment solids, overlying and pore water Sediment cores and associated overlying waters were collected at each site within an area of about 10 m 2 with a hand corer using acrylic tubes (8–10 cm, inner diameter). The hand corer was operated from a small plastic boat facing the water flow direction or, in some cases, by wading from shore. Cores with undisturbed overlying water were delivered to the laboratory within 2 h. Overlying waters were siphoned off, filtered (Sterivex™, nominal cutoff 0.45 μm) and sub-sampled for THg and MMHg analyses. Cores were sliced (2 cm for site VE1 and 3 cm for site VE2) in a Hg free N2-purged glove-box. Pore waters were recovered by centrifugation (4000 rpm, 40 min) and filtered as for overlying waters after returning to the glove-box. Solids were immediately preserved by freezing, while pore waters were preserved as described below. Two separate aliquots from sliced sediments were also analyzed for grain size distribution and percentage moisture. This latter value was used to convert measured concentrations from a wet to a dry weight basis. 2.2.2. Water column and total suspended solids Water samples were collected from a small plastic boat and all sampling and treatment procedures were carried out following ultra-clean techniques (Cossa and Gobeil, 2000). All materials in contact with samples were acid-washed (5 days in 20% HNO3 v/v followed by 5 days in HCl 10% v/v) and rinsed several times with deionized water (Milli-Q®)

Fig. 1. Venice Lagoon map and sampling sites (gray points). The location of the industrial area of Porto Marghera (a historically-known source of Hg to the lagoon) is also indicated.


before use. Polyethylene gloves were used for handling operations. Clean Teflon® (FEP) bottles were stored in double polyethylene bags until use. Aliquots for total dissolved mercury ((THg)D) and dissolved methylmercury ((MMHg)D) analysis were filtered with Sterivex®-HV 0.45 μm sterile filters, collected in FEP bottles, and acidified with HCl 0.5% v/v (Millipore® Seastar) (Parker and Bloom, 2005). Aliquots of suspended material for the determination of total particulate mercury ((THg)P) and methylmercury ((MMHg)P) were obtained by filtration (one filter for each Hg species) through hydrophilic Teflon® membranes (LCR, Millipore®, 0.45 μm pore size, 47 mm diameter; Cossa et al., 1996). Total suspended solids (TSS) were determined gravimetrically. 2.2.3. Time integrated pore water profile (peepers) An integrated Hg/MMHg pore water profile was obtained using dialysis membrane techniques (metacrylate peeper) with a 0.45 μm membrane (HAWP, Millipore®) and 1 cm depth resolution. Peeper was first acid washed, filled with Milli-Q® water, and degassed with Hg-free nitrogen during 15 days (Guédron et al., 2011). The peeper was placed in the sediment at site VE1 during 10 days for osmotic equilibration. Aliquots for (THg)D, (MMHg)D, and (Fe II) analysis were collected in a Hg free N2-purged glove-box, sampled in peeper cells every 1 cm and stored as previously described. Aliquots for dissolved ferrous iron (Fe II) analysis were stored in acid washed (HNO3 10% v/v) HDPE tubes after acidification to pH 1 (HCl 7% v/v, Millipore® Seastar). The sulfide-accumulating zone (SAZ) was identified with sulfide sensitive sellotape (fixed on the peeper during the 10 days of equilibration), through the formation of a surface darkening Ti–S complex (Jezequel et al., 2007). 2.2.4. Time series of pore and surface water (rhizons) The study of temporal variations of Hg species during the two tides was performed at site VE1 using Milli-Q®-rinsed microporous polymer tube samplers (rhizon® SMS, Rhizosphere Research Products) fixed on a metacrylate plate and inserted vertically into the sediment at 2 cm depth (Seeberg-Elverfeldt et al., 2005). The rhizon sampler integrated the pore water between − 2 and 0 cm and the overlying water between 0 and + 8 cm. It should be noted that rhizons preferentially recover water from the sediment macropores, while the other techniques (sediment centrifuging or peepers) sample bulk porewater dominated by water contained in micropores (Harvey et al., 1995). Therefore, the former reflects the concentration in water readily exchangeable with the water column (rapid flushing), while the latter are suitable for the evaluation of molecular diffusion. During the same experiment, the water column was sampled 50 cm above the sediment surface using an HDPE hand pump to get homogenized water column samples. 2.3. Analytical methods 2.3.1. Sediment analysis Sediment samples were freeze-dried and crushed with a rotating agate mill to grain size smaller than 63 μm for Hg analysis (Guédron et al., 2006). Total Hg concentrations ([THgP]) were determined by atomic absorption spectrophotometry following catalytic decomposition and gold amalgamation (Guédron et al., 2009; Roos-Barraclough et al., 2002) with an automatic mercury analyzer (Altec, Model AMA 254) with a relative precision of ± 10% determined from triplicates. Concentrations obtained for repeated analyses of certified reference materials (CRM) were always in the certified range for the standard BEST-1 (0.092 ± 0.009 μg g − 1 — National Research Council of Canada, estuarine sediment). The detection limit, defined as three times the standard deviation of the blank (SDblk), was 0.005 μg g − 1. Monomethylmercury concentrations ([(MMHg)P]) in sediment samples were determined after acid digestion as described for suspended particle (see Section 2.3.2 below).

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The grain size distribution was measured using a particle size analyzer Coulter LS-100 (Beckman® Coulter, Fullerton, CA, USA), following ultrasonic dispersal in de-ionized water (Loizeau et al., 1994). The proportions of three major size classes (clay b 2 μm; silt 2–63 μm; and sand 63–2000 μm), as well as the median grain size, were determined from size distributions. The water content was obtained by the difference in sample weight before and after freezedrying. Organic matter (OM) and carbonate (CaCO3) content were determined by Loss On Ignition (LOI), i.e., by pyrolyzing the samples for 2 h at 550 °C for OM and 950 °C for CaCO3 (Heiri et al., 2001). 2.3.2. Water and particle analyses Samples were analyzed for [(THg)D], [(THg)P] on TSS, [(MMHg)D] and [(MMHg)P] (TSS and solid sediments) by cold vapor atomic fluorescence spectrometry (CVAFS) using a Tekran® (Model 2500) mercury detector after conversion of all mercury species into Hg 0 (Bloom and Fitzgerald, 1988). The principles of the methods used are from the Bloom and Fitzgerald (1988) gold amalgamation method for (HgT)D, from the Liang et al. (1994) ethylation method for (MMHg)P and the Filippelli et al. (1992) hydride-generation method modified by Cossa et al. (2009) for (MMHg)D. Analysis of (HgT)P in suspended particles was performed after HCl/ HNO3 digestion (10 h at 70 °C) in PFA Teflon reactors (Coquery et al., 1997). The accuracy of analyses was checked using CRM ORMS-4 (22.0 ± 1.6 ng L− 1 — National Research Council of Canada) for (HgT)D, CRM BEST-1 for (HgT)P and CRM ERM-AE670 (175.1 ± 62 μmol L− 1 – IRMM – European Commission) for (MMHg)D and (MMHg)P. The measurement error (1SD for triplicate measurements of several samples) was usually about 10% and always better than 15% for [(HgT)D], [(HgT)P], and [(MMHg)D]; and always better than 20% for [(MMHg)P]. The detection limits (3SDblk) were 0.01 ng L − 1 for (HgT)D, 0.004 ng L− 1 for (MMHg)D, and 0.05 ng g− 1 for (MMHg)P. Dissolved ferrous iron concentration ([Fe II]) was measured with the 1,10-phenanthroline method (Tamura et al., 1974) using an UV–VIS spectrometer (Milton Roy, model Spectronic 1201) and the detection limit (3SDblk) was 0.01 mg L − 1. Eh profiles were determined using an Eh probe (PVRTCE, Ponselle Mesure, Viroflay, France) in an independent core sampled at site VE1. 2.4. Estimation of exchanges between porewater and overlying water 2.4.1. Molecular diffusion fluxes Molecular diffusion from sediment to overlying water was estimated by calculation of the diffusion fluxes for Hg(II)D (with Hg(II)D = THg − MMHg) and MMHg at the sediment–water interface using the measured concentration gradient and Fick's first law (Berner, 1980; Rothenberg et al., 2008): ! " 2 FD ¼ −ΦDw=θ ðdC=dzÞ where FD is the diffusive flux of a solute (ng m − 2 d− 1) with concentration C (ng L− 1) measured in core pore and overlying waters at depth z (cm), Φ is the sediment porosity, Dw is the ionic/molecular diffusion coefficient in water, θ is the tortuosity and dC/dz is the gradient of the mean Hg concentration between the overlying water (on the top of the sampled cores) and pore-water in surface sediment (0–2 cm for site VE1 and 0–3 cm for site VE2). Porosity was determined for each surface sediment sample from the wet and dry weights according to Crusius and Anderson (1991) and was 0.74 and 0.51 for site VE1 and VE2, respectively. Tortuosity was estimated from the porosity using Boudreau's formulation: θ 2 = 1 − ln(Φ2) (Boudreau, 1999). We used Dw, 25 °C = 2 × 10 − 6 cm2 s− 1 and 1.2 × 10− 5 cm2 s − 1 for Hg(II)D and MMHg ionic/molecular diffusion coefficient in water, respectively according to Rothenberg et al. (2008).

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2.4.2. Advective flux calculations Tidal flushing tends to push pore-waters from sediment during tidal events and thus export these waters to the water column (Precht and Huettel, 2004; Santos-Echeandía et al., 2010). We calculated the advective daily flux (FA) of dissolved THg and MMHg by introducing the temporal variation of Hg concentration during tidal cycles into the following global mass balance equation: %# & $ FA ¼ Σ Ctþ1 % htþ1 −ðCt % ht Þ =dt where Ct + 1 and Ct are dissolved Hg concentration in the flooding water column at times t + 1 and t, ht + 1 and ht are the water levels at the same times and dt is the time interval between samples. We assume that the calculated flux results from the combination of local advection of pore-water Hg content and surface turbulence associated with wave current affecting unconsolidated surficial sediments. We also consider that the dissolved species exported from the sediment to the water column are rapidly homogenized in the water column by turbulent mixing. The transport of Hg species is calculated only for the tidal flooding because concentrations were relatively stable during the rest of the tidal cycle. Because tidal cycle occurs twice a day, the advective daily flux (FA) was multiplied by two. 2.5. Statistical treatment We used the Mann–Whitney rank sum test (U test) or the Kruskal– Wallis one way analysis of variance by ranks (H test) to compare two or more than two sets of data, respectively. Pairwise multiple comparison according to Dunn's method was then used to identify the specific differences among sets (Webster, 2001). Linear regressions were performed only when normality test, constant variance test and alpha power of the performed regression (α test, p > 0.8) passed. Pearson's correlation coefficient (R) and p values are reported for the computed linear regressions. Mean concentrations (arithmetic means) are always presented with the associated standard deviation (mean ± 1 SD). All statistical analyses were performed using the package Sigmastat® (Systat Software Inc.). 3. Results 3.1. Sedimentological and geochemical parameters of sediments At both sampling sites and considering 2008 and 2009 cores, the percentage clay fraction (b2 μm) never exceeded 1%, indicating winnowing of the finest sediment fraction because of tide-driven and anthropogenic re-suspension of surface sediments. However, the low amount of clay measured in sediments cores could also result (at least partly) from some known limitations of the laser counting technique (Beuselinck et al., 1998; Goossens, 2008). Whichever the case, the selected sites are representative of two contrasting sedimentary environments typical of the Venice Lagoon (Zonta et al., 2007). Silt was the dominant fraction at site VE1 (mean= 57.4 ± 17.4% and showing very large variations with depth and in between profiles; Fig. 2), while at site VE2 grain size was dominated by sands (60.2± 10.8%; Fig. 2). At both sites, OM was significantly correlated with the percentage of silt fraction (R= 0.76, p b 0.001, n = 54) showing that the OM was mainly composed of silt-sized material, while carbonates were strongly associated to the sand fraction (R= 0.77, p b 0.05, n = 54). Sediments from site VE1 were characterized by a larger OM content than at site VE2 (OM = 5.6 ± 2.0% vs. 2.9 ± 1.6%, respectively). OM, carbonate content and grain size reflected the relatively remote situation of site VE1 (which favors the deposition of finer, carbon-rich particles) in contrast to site VE2 where the proximity of the Lido inlet increases tidal erosion. Vertical variations in water content also strongly correlated to the silt content of sediments (R = 0.85, p b 0.01). As for silt content, the water content was significantly larger at site VE1 (39.8 ± 9.4%) than

at site VE2 (26.3 ± 4.3% — U test, p b 0.001) and showed a sharp decrease below the uppermost four centimeters at site VE1. Within-site variability of sediment grain size was lower at VE2 than at VE1. This variability was partly related to the collection year of the cores. Sediments collected at VE1 in 2009 had lower silt and OM, and higher carbonate content than those collected in 2008. A high variability at site VE1 can be due to the uneven bathymetry of this shallow remote area bordered by a multitude of salt marshes which begets turbulent flow and specific local tidal currents leading to heterogeneous patterns of depositions and re-suspension (van Proosdij et al., 2006). 3.2. Mercury speciation in sediments Total mercury concentrations ([THgP]) in solid sediments were in the range of those reported for the northern part of the Venice Lagoon (Bernardello et al., 2006; Berto et al., 2006; Bloom et al., 2004b; Moretto et al., 2003; Zonta et al., 2007). They were higher at site VE1 (mean 621.9±213.7 ng g− 1, n=31) than at site VE2 (386.9±92.7 ng g− 1, n=30) for both years (U test, pb 0.001). [THgP] in sediment cores varied strongly with depth at both sites (Fig. 3). These variations were mainly controlled by the silt/sand proportion and the changes in carbonate and OM contents with depth (Fig. 2). THgP was strongly correlated with the silt size fraction (R=0.70, pb 0.05) and OM (R=0.73, pb 0.05) for all samples. As for [THgP], [MMHg] were significantly higher at site VE1 than site VE2 (U test, pb 0.01), but never exceed 0.45% of THg at both sites. [MMHg] at site VE1 were also correlated with [THgP] (R=0.94, pb 0.05) and OM (R=0.88, pb 0.05), whereas no significant correlation was found between [MMHg] and [THgP], OM or silt content for site VE2. At site VE2, [(MMHg)P] showed similar profiles for all cores, with largest concentrations in the surface sediments (i.e., 0–3 cm, mean=0.92± 0.39 ng g− 1) followed by a sharp decrease below −3 cm to reach the lowest concentrations (0.43±0.17 ng g− 1) at the bottom of cores. 3.3. Mercury speciation and partitioning in overlying and pore-water At site VE1, pore-water [(THg)D] and [(MMHg)D] measured in 2009 cores exhibited large vertical and spatial variations (Fig. 4a). Considering both cores, [(MMHg)D] peaks occurred in three zones below the sediment–water interface (SWI): (i) in the iron-reduction zone (IRZ) (i.e., between the SWI and −4 cm, also corresponding to the redox transition zone), (ii) below the IRZ (i.e., between −4 and −8 cm) and (iii) in the deep anoxic sediments (i.e., between −8 and −11 cm). The presence of such peaks of MMHg in the pore waters recovered from suboxic and anoxic sediment (i.e., in the IRZ and in the sulfate reduction zone (SRZ), respectively) is in accordance with the current consensus that Hg methylation in sediments is due to the activity of sulfate reducing bacteria (SRB — Benoit et al., 2001a; Compeau and Bartha, 1985; Gilmour et al., 1998; Han et al., 2008) and iron reducing bacteria (IRB — Fleming et al., 2006; Han et al., 2008; Kerin et al., 2006). The time-integrated profiles given by the peeper (Fig. 4b) corroborate the snapshot distribution of (THg)D and (MMHg)D observed in pore waters obtained from cores. However, the much higher vertical resolution shows that the (MMHg)D distribution is characterized by a multi-peak feature. A progressive increase in [(MMHg)D] was observed from the overlying waters down to −2 cm, where the first peak occurred. The decreasing Eh in this zone points to MMHg diffusion from the production zone (subsurface sediments) toward the sediment–water interface. Three other peaks were observed in the SAZ, under anoxic conditions, at depths of −7 and −13 cm and −21 cm. At site VE2, [(THg)D] profiles in cores showed a general increase with depth while [(MMHg)D] showed maximum values around −6 cm (Fig. 4a). As for particulate THg, total dissolved mercury concentrations ([(THg)D]) were lower at site VE2 than site VE1 (U test, pb 0.001). In opposition to (THg)D, [(MMHg)D] were significantly higher at site VE2 than at VE1 (U test, p b 0.001). Percentages of methylated mercury were also


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Fig. 2. Silt, organic matter (OM), calcium carbonate (CaCO3) and water content (%) in sediment cores sampled at site VE1 (hexagons) and site VE2 (diamonds). Hourglass symbols refer to 2008, empty symbols to 2009.

higher at site VE2 (0.2 to 42% of THg) than at site VE1 (0.3 to 5.6%). The main methylation zone was identified around −5 cm depth in the highly anoxic sediments (Fig. 4a).

3.4. Tidal cycling of Hg species in suspended solids and water column at site VE 1 The temporal evolution of dissolved Hg species at the SWI (studied with rhizon samplers) showed a sharp increase in [(THg)D] and [(MMHg)D] during the late tidal flooding and during both tides, albeit with a very different amplitude for [(MMHg)D] (Fig. 5). A marked increase in [(MMHg)D] in the water column was also observed for both tidal cycles and it was followed by a gradual decrease at ebb tide. For [(THg)D], a moderate increase was observed only during the second tide (18th September). TSS concentrations ranged from 60 to 180 mg L− 1 during tidal flooding and sharply decreased to around 20 mg L− 1 during ebb tide (Fig. 5). [(THg)P] were constant during the tidal cycle (305± 69 ng g− 1) except at the maximum water level when [(THg)P] were about 50–100% higher than at other times. [(MMHg)P] (ng g− 1) were constant (2.7 ± 0.2 ng g− 1) during tidal flooding and slightly increased after the maximum water level and during ebb tide for both tidal cycles.

4. Discussion 4.1. Availability of mercury for methylation in Venice Lagoon sediments The distribution of THg in solid sediments points to THg inputs from the continental side of the lagoon. This observation is consistent with earlier findings that evidenced a decreasing THg gradient from the inner border of the Lagoon toward the sea inlets. This contamination pattern reflects the increasing distance from the source of Hg contamination (i.e., Marghera industrial site) and the increased tidal erosion and grain size distribution due to the increased proportion of carbonates in sediments toward the sea inlets (Basu and Molinaroli, 1994; Bernardello et al., 2006; Sfriso et al., 2003). MMHg in surface sediments shows similar features. The sheltered location at site VE1 favors the accumulation of fine particles enriched in OM which are a rich substrate for Hg methylators and have strong adsorption capacity for the binding of (MMHg)D as shown by their higher (U test, p b 0.05) log partition coefficients (logKD = log [MMHg]p − log[(MMHg)D]) for MMHg (mean = 6.89 ± 0.42 L kg − 1) than those of the open lagoon sediments (VE2 — MMHg logKD = 6.30 ± 0.43 L kg − 1) which are enriched in sand. The source of MMHg has been identified in the sediment, mainly at the depths corresponding to the reduction of iron oxides (i.e., the IRZ)

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Fig. 3. Total mercury [THg] and mono-methylmercury [MMHg] concentrations (ng g− 1) in sediment cores sampled at site VE1 (hexagons) and site VE2 (diamonds). Hourglass symbols refer to 2008, empty symbols to 2009.

and the oxidation of fresh OM in surficial suboxic sediments, and at the depths of sulfate reduction in deeper anoxic sediments (i.e., the SRZ). For both sites, the absence of correlation between (MMHg)D and (THg)D concentration in both peeper and core profiles corroborates the hypothesis that the availability of dissolved Hg for methylating microorganism is the key limiting factor for MMHg production in the surface sediments in the Venice Lagoon (Han et al., 2007). In surface layer, dissolved organic matter might be the main ligand for dissolved inorganic Hg(II) in aerobic sulfide-free waters (Benoit et al., 2001b; Covelli et al., 2008). In deeper layers, the solubility and availability of Hg for methylating microorganisms are controlled by the precipitation of HgS (Fink, 2002; Mehrotra et al., 2003; Mehrotra and Sedlak, 2005) and/or by the adsorption on and co-precipitation of Hg(II) with FeS(s) (Han et al., 2010; Han et al., 2007; Han et al., 2008) as evidenced by the gradual decrease of [(THg)D] with depth in the SAZ (Fig. 4). Although higher porewater [(MMHg)D] were found in the SAZ, the sharp decrease in water content below −3 cm depth in sediments and the important increase in [(MMHg)D] at the SWI observed during time series (Section 3.4) show that the iron reduction zone is the most important source of MMHg for the water column. 4.2. Tidal cycling of THg and MMHg between sediment and water column The increase in [(MMHg)D] and [(THg)D] (the latter only for tide 2) observed in the water column during the late tidal flooding shows that the release of dissolved Hg species from the sediment adjacent to the salt marshes during tidal flushing is significant enough to affect the water column concentration. This observation is consistent with other

studies which evidenced THg and MMHg exports from salt marsh sediments to the water column during flooding events (Canário et al., 2007; Choe et al., 2004; Langer et al., 2001; Santos-Echeandía et al., 2010). The sharp increase of both dissolved Hg species concentrations measured at the SWI and, to a lesser extent, in the water column during the tidal flooding might be caused by shear stress along lagoon sediments. Shear stress may (i) increase diffusion or advection of dissolved species from surficial sediments and/or (ii) promote in situ production of MMHg or desorption of both Hg species from sediment particles due to the oxidative dissolution of redox sensitive sulfide minerals. With regard to hypothesis (i), the average molecular diffusion fluxes (FD — see Section 2.5) calculated from concentrations measured in cores for both Hg(II) (9.46 ± 0.38 ng m− 2 day− 1) and MMHg (1.09 ± 0.97 ng m− 2 day− 1) are comparable to those obtained in the Lavaca Bay and the Gironde estuary (Gill et al., 1999; Schäfer et al., 2010), but are both around 10 fold lower than benthic fluxes obtained in the open Grado Lagoon (Covelli et al., 2008). However, these fluxes are about 3 orders of magnitude too small to explain the peaks of dissolved Hg and MMHg concentration in the water column during the tidal flooding. Shear stress might therefore play a key role in the advection of dissolved Hg and MMHg from the surficial sediment pore water (PW) to the water column. The increase in Hg and MMHg concentrations observed during the tidal flooding are consistent with findings made in intertidal salt marsh sediments which showed that at tidal flooding PW advection pushed reduced species (ferrous iron, sulfur, manganese…) toward the water column during tidal inundations (Beck et al., 2008; Chow, 2007; Taillefert et al., 2007). This observation is in accordance with the variations of dissolved Fe II concentration at the SWI which


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Fig. 4. a) Cores — dissolved total mercury [(THg)D] and mono-methylmercury [(MMHg)D] concentrations and redox potential (Eh) vertical profiles in the overlying and pore water from sediment cores sampled at site VE1 (hexagons) and site VE2 (diamonds) in 2009. Redox potential was measured separately in additional cores (see Section 2.3.2). b) Peepers — profiles of [(THg)D] (ng L− 1 — white hexagons), [(MMHg)D] × 100 (ng L− 1 — black hexagons), and ferrous iron [(MMHg)D] × 100 (mg L− 1 — plus symbols) in the overlying and pore water obtained with peepers at site VE1 deployed for 10 days (September 7th to 17th 2009). The Sulfide-Accumulating Zone (SAZ) is marked as the shaded gray background band showing the increasing coloration of the sensitive sellotape with depth. The Sediment–Water Interface (SWI — 0 cm) is plotted as a dotted horizontal line.

followed dissolved MMHg during tide 1 (pb 0.05, R = 0.76 — Fig. 5). Calculated daily advective fluxes (FA — see Section 2.5) from surficial sediment to water column for both dissolved Hg(II) (1260 ng m− 2 day− 1) and MMHg (92 ng m− 2 day− 1) are also in accordance with the finding of Precht and Huettel (2004) who evidenced that wave-driven advection was at least 3 orders of magnitude faster than solute transfer by diffusion in a coastal sandy sediment. Assuming that adsorption on sediment is not the key limiting factor controlling the dissolved species concentration in surface porewater (i.e., 0–2 cm interval) during tidal cycles, we can make the hypothesis that the rate of methylation in surface sediment is roughly equal to the ratio of the advective fluxes of dissolved MMHg to dissolved THg. Thereby, the daily methylation rate during this summer period in surface sediments would be equal to 6.8% which is in the range of those obtained with amended anoxic sediment slurries of the Venice Lagoon (Han et al., 2008) or in near-shore marine sediments (Hammerschmidt and Fitzgerald, 2004). This high value could explain the low percentage of MMHg in solid sediments (b0.45%) and suggests that most of produced MMHg in PW of surface sediments is pushed in surface water during tidal flushing. However, this hypothesis does not consider desorption from solids during tidal oscillations and the calculated methylation rate is only a gross estimation. Consequently, other probable causes (ii) have to be considered for the increase in both dissolved Hg species concentrations at the SWI depth including a rapid desorption from sediment particles (Hintelmann and Harris, 2004), in situ micro-organisms production for MMHg, or release from organic matter degradation of both Hg species during redox

transition events (Bouchet et al., 2010; Muresan et al., 2008). Indeed, the wave driven shear stress applied to surface sediment during tidal flooding (i.e., tidal flushing) could enhance oxygen penetration and change the redox state of surface sediments (Precht and Huettel, 2004; Taillefert et al., 2007). Redox oscillations in the surface sediments are known to favor inorganic Hg desorption from particles due to the oxidative dissolution of solid phases such as redox-sensitive MnS and FeS (Domènech et al., 2002; Lewis et al., 2007; Moses and Herman, 1991; Schäfer et al., 2010) and provide further substratum for MMHg formation. In parallel, the rapid recycling of iron oxides in subsurface sediment can cause the peaks of (MMHg)D observed in cores and peeper profiles probably resulting from an increased activity of microbial Fe(III) reduction linked to organic carbon oxidation (Hyun et al., 2009; Kostka et al., 2002). These redox oscillations can also favor MMHg production as evidenced by Bouchet et al. (2010) in a microcosm experiment using a sediment slurry spiked with 199Hg and 201MMHg exposed to oxic–anoxic oscillation with a 4 hour sampling step. The following decrease of [(THg)D] and [(MMHg)D] at high tide and at the beginning of the ebb tide can result from their rapid adsorption onto soluble organic-Fe(III) complexes and/or on Fe(III) (oxy)hydroxides resulting from the oxidation of FeII in surface sediments and surface water (Carey and Taillefert, 2005; Taillefert et al., 2007). The absence of dissolved Hg and MMHg peaks during the ebb tide is attributed to the sheltering effect of salt marshes which dissipate wave energy and reduce the effect of tidal flushing on surface sediments (Lopes et al., 2001; van Proosdij et al., 2006). Considering the total pools of THg

8


Fig. 5. Upper panels (first 6): temporal variations of water level (cm — cross), dissolved ferrous iron (FeII — squares), dissolved total mercury ((THg)D — triangles) and methylmercury ((MMHg)D — diamonds) concentrations monitored in the water column (white symbols) and at the Sediment–Water Interface (SWI — gray symbols) with Rhizon. Bottom panels (last 4): total suspended solids (TSS — X symbols), particulate (black symbols) total mercury ((THg)P) and methylmercury ((MMHg)P) concentrations. All parameters were monitored during the tides of September 17th (tide 1) and 18th (tide 2) 2009 at site VE1.

and MMHg in the water column (expressed in ng L − 1 — Fig. 6), the contribution of the particulate phase is major only during the tidal flooding, whereas at high and ebb tide the proportion of both species in the filterable phase is around 20% for (THg)D and between 20 to more than 50% for (MMHg)D; the maximum percentages being observed at high tide when particles settle. The relatively high and constant (MMHg)D concentrations in the water column from the late rising tide to the ebb tide suggest that released MMHg during the late tidal flooding is not entirely trapped by settling particles at high tide and/or that colloids may be a major carrier phase for (MMHg)D. At ebb tide, the slow decay of [(MMHg)D] in the water column with the increase in particulate

MMHg (ng g− 1) can be attributed to plankton and nekton uptake of dissolved and/or colloidal MMHg, or to demethylation. At site VE2, the average molecular diffusion flux values are significantly lower for Hg(II) (1.05±0.22 ng m− 2 day− 1) due to the lower porosity of sediments than at site VE1 (Φ=0.74 vs. 0.51 for site VE1 and VE2, respectively), while those for MMHg are in the same range (0.83± 0.51 ng m− 2 day− 1) as those of site VE1 due to the high MMHg concentrations measured in the near surface sediment at site VE2. Considering the low adsorption capacity of site VE2 sediments and its more erosive position in the lagoon, open lagoon sediments may be significant contributors for dissolved MMHg release to the water column.


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Fig. 6. Upper panels (first 2): temporal variation of particulate (black symbols) total mercury ((THg)P — triangles) and methylmercury ((MMHg)P — diamonds) concentrations (ng L− 1). Bottom panels (last 2): dissolved vs. total (i.e., dissolved + particulate) ratios (% — crossed symbols) for total mercury and methylmercury monitored in water column during the tides of September 17th (tide 1) and 18th (tide 2) 2009 at site VE1.

4.3. Particles as vehicles for methylmercury dispersal in the Venice Lagoon The distribution of total suspended solids (TSS) with time during the two tides at site VE1 confirms that this remote site is a sedimentation area (Fig. 5). Indeed, the 5 to 8 times lower TSS concentrations during ebb tide than tidal flooding corroborate previous observation on the sheltering effect of marshes which limit the re-suspension of particles during the ebb tide (van Proosdij et al., 2006). In parallel, the co-variation of both [(THg)P] and [(MMHg)P] (expressed in ng L − 1 — H test, p b 0.05; Figs. 5 and 6) with [TSS] during the two tidal cycles shows that salt marshes act as a trap for particles and associated mercury due to the reduced current velocity at ebb tide which permits the deposition of cohesive particles in this landward side of the lagoon (Amos et al., 2010; Molinaroli et al., 2009). When expressed in ng g − 1, [(THg)P] in suspended particles only peaked at the maximum water level due to the deposition of the heaviest particles at high tide (when the current speed reaches its minimum) and to the relative enrichment of fine and Hg-rich particles. On the contrary, the higher [(MMHg)P] (ng g − 1) found during ebb tide suggests that during the tidal flooding, particles coming from the sea dilute [(MMHg)P], whereas after the high tide, when most of the large and heavy particles settled, the remaining particles (small, light, and probably mostly organic) are enriched in MMHg. The comparison of mean Hg concentrations between TSS and surface sediment corroborates this observation since [(MMHg)P] in TSS (3.5 ± 1.7 ng g − 1) were about twice those of surface sediment (1.7 ± 0.6 ng g − 1), while [(THg)P] in surface sediment (676 ± 127 ng g − 1) was about twice those of TSS (330 ± 99 ng g − 1). Furthermore, the percentage of MMHg in the THg fraction on TSS ranged between 0.8 and 1.2%, while it never exceeded 0.45% for sediment particles. Finally, assuming that all the advected dissolved MMHg and Hg (II) are adsorbed onto suspended particles, estimated daily particulate loads are 1.5 and 20.4 ng g − 1 day − 1 for MMHg and Hg(II), respectively representing a mean increase of 42 and 6% for particulate MMHg and Hg(II) concentration, respectively. Even if these particulate loads are

probably overestimated, since demethylation in the water column and adsorption onto settling particles are not considered, tidal flushing could be an important contributor to the suspended particulate load in MMHg. Therefore, sediments act as a net source of MMHg for suspended particles and a sink for particulate inorganic Hg. Such (MMHg)P enrichment of TSS surely has a direct incidence on the dispersal of MMHg in the Venice Lagoon because TSS are well-known vehicles for pollutant transport (Berto et al., 2006). 5. Conclusions In this study, we confirmed that Hg methylation is very active in two areas of the Venice Lagoon with contrasting sedimentation characteristics. Hg methylation in sediments was restricted to the upper eight centimeters in a moderately erosive area (site VE2) and reached 22 cm in a low-energy, depositional area bordering salt marshes (site VE1). MMHg pore water profiles and time series showed that the iron reduction zone is the most important source of MMHg for the water column, although higher porewater MMHg concentrations were found in the sulfate reduction zone. Time-series measurements at site VE1 during two tidal cycles highlighted the strong influence of tidal flushing on the release of both dissolved THg and MMHg from surface sediments to water column during the tidal flooding. The release of MMHg during tidal flushing seems to play an important role in the MMHg enrichment of TSS which may strongly contribute to the dispersal of MMHg contamination in the lagoon. We also showed that MMHg concentration in pore water depends, among other factors, on the sorption capacity of sediments. Indeed, high concentrations of MMHg were measured in low sorption capacity sediments (e.g., calcareous sands) which can also be an important source of dissolved MMHg for the water column due to tidal-driven advection. Finally, tide-driven mobilization of MMHg needs to be considered as an important variable in methylmercury studies in tidal environments as it may considerably influence MMHg availability to aquatic organisms, particularly those feeding on (suspended) particles.

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Acknowledgments This study was supported by the Swiss National Science Foundation (grant no. 200020-117942 and no. IZKOZ2_136134/1). We are greatly indebted to the boat pilot Loris Dametto (CNR-ISMAR) and Philippe Arpagaus (Institut F.-A. Forel) for their dedication during sampling campaigns. We also thank Dr. Davide Tagliapietra (CNR-ISMAR) for his help in field work and advice on the selection of sampling sites, Dr. Andrea Pesce (CNR-ISMAR) for providing logistical support for laboratory work, Dr. Daniel Cossa and Dr. Laurent Oxarango for their insightful discussions and constructive comments in the manuscript preparation.

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