Mercury and Methylmercury Dynamics in Sediments ...

Water Air Soil Pollut (2016) 227:475 DOI 10.1007/s11270-016-3179-2

Mercury and Methylmercury Dynamics in Sediments on a Protected Area of Tagus Estuary (Portugal) Rute Cesário & Carlos Eduardo Monteiro & Marta Nogueira & Nelson J. O’Driscoll & Miguel Caetano & Holger Hintelmann & Ana Maria Mota & João Canário

Received: 8 September 2016 / Accepted: 22 November 2016 # Springer International Publishing Switzerland 2016

Abstract The Tagus Estuary is one of the most Hgcontaminated estuaries in SW Europe. Sediment cores were sampled at two low Hg-contaminated sites inside the natural park, Alcochete (ALC) and Vale Frades (VF), and analyzed for mercury and methylmercury. Concentrations of Hg and MeHg in sediments were below 1 μg g−1 and 4.4 ng g−1, respectively. While in summer organic matter and/or excess SO42− promotes Hg methylation, in winter, Hg availability is the sole driver for methylation. Diffusive fluxes in the sediment/ water interface show a sink of Hg species in the ALC site (ca. 170 mg year−1 of Hg and 60 mg year−1 of MeHg), while in the VF area, a sink of MeHg (ca. 1900 mg year −1) as well as a source of Hg (ca. Electronic supplementary material The online version of this article (doi:10.1007/s11270-016-3179-2) contains supplementary material, which is available to authorized users. R. Cesário : C. E. Monteiro : A. M. Mota : J. Canário (*) Centro de Química Estrutural, Instituto Superior Técnico, Universidade de Lisboa, Av. Rovisco Pais, Lisbon, Portugal e-mail: [email protected] R. Cesário : C. E. Monteiro : M. Nogueira : M. Caetano IPMA-Instituto Português do Mar e Atmosfera, Av. Brasília, 1449-006 Lisbon, Portugal N. J. O’Driscoll Department of Earth and Environmental Science, Irving Environmental Science Center, Acadia University, Wolfville, NS, Canada H. Hintelmann Department of Chemistry, Trent University|, 1600 W Bank Dr., Peterborough, ON K9J7B8, Canada

2000 mg year−1) is observed. The morphology and hydrodynamic regime of the Tagus Estuary seem to influence Hg dynamics even in areas with low levels of Hg contamination. Keywords Mercury . Methylmercury . Sediment cores . RAMSAR site . Tagus Estuary

1 Introduction Mercury (Hg) pollution is a significant global issue due to its ability to undergo long-range transport (Asmund and Nielsen 2000; Jewett et al. 2003) and adversely affect remote ecosystems and human health (e.g., Liu et al. 2012). Mercury biogeochemistry has become an important topic in the environmental sciences. All forms of Hg are toxic, but the Hg-organic compounds raised increased concern, namely methylmercury (MeHg), which is a potent neurotoxin (Clarkson and Magos 2006; Grandjean et al. 2010) and endocrine-disrupting chemical (Tan et al. 2009), highly toxic to humans and other organisms (Crespo-López et al. 2009). Biotic and abiotic Hg methylations are the mechanisms responsible for producing MeHg in the aquatic environment. The first one involves the microbial mediation by anaerobic bacteria (Parks et al. 2013), and the second relates with the chemical methylation through the transfer of methyl groups from other metal-alkyl species such as methyl-Sn complexes (Celo et al. 2006). It is widely accepted that biotic processes dominate over the chemical, although the abiotic processes

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in mercury methylation have been shown to be a significant factor (Castro et al. 2010; Ebinghaus and Wilken 1993; Weber 1993). Estuarine sediments act as a reservoir for past and present Hg inputs due to affinity of inorganic species for organic matter (Gagnon et al. 1996; Mason and Lawrence 1999). Mercury speciation and the bioavailability of divalent forms (Hg2+) to microbes (rather than the total Hg pool) are primary factors governing the methylation in sediments, widely recognized as a willing site for methylation (Gilmour et al. 1992; Hammerschmidt et al. 2004; Hines et al. 2004; Ullrich et al. 2001). Additionally, the microbial community has also an important role in sediments (Mauro et al. 1999) where sulfate-reducing bacteria (SRB) are known to be the main MeHg producers in sediments (Compeau and Bartha 1985; Gilmour et al. 1992; King et al. 1999). However, ironreducing bacteria (IRB) (Fleming et al. 2006; Kerin et al. 2006; Yu et al. 2012) and methanogens (Hamelin et al. 2011; Wood et al. 1968; Yu et al. 2013) can also be involved in the methylation processes, although to a lesser extent. It is commonly accepted that in situ Hg methylation occurs in anoxic conditions both in sediments and water column (e.g., Eckley and Hintelmann 2006; Furutani and Rudd 1980) due to the activity of several bacteria populations. Because these microorganisms are anaerobes, benthic sediments in coastal and estuarine systems often provide the most suitable environment for MeHg production (Choi et al. 1994; Sunderland et al. 2004). Besides the microbial activity of SRB, other processes contribute to the increase of MeHg in sediments, such as the amount and characteristics of dissolved organic matter (DOM) (Ravichandran 2004), dissolved sulfur species (Gagnon et al. 1996; Hammerschmidt et al. 2004; Lambertsson and Nilssons 2006), and sorption and dissolution processes involving Fe and Mn oxyhydroxides (Bloom and Lasorsa 1999; Gagnon et al. 1997). The interplay of organic matter oxidation and SO42 − 2− /S redox cycle is dominant under suboxic-anoxic conditions (Merritt and Amirbahman 2009) and plays an important role in mercury methylation. If high levels of sulfate can lead to increments in microbial activity, otherwise the increase of S2− leads to precipitation of inorganic Hg as a HgS mineral, cinnabar (red coloration), or meta-cinnabar (black coloration) due to its extremely low solubility (Randall and Chattopadhyay 2013). Therefore, the presence of S2− in anoxic conditions decreases Hg availability for methylation.

Water Air Soil Pollut (2016) 227:475

Seasonal variations of Hg and MeHg in sediments appear to be mainly related to temperature, redox status, and seasonal changes in productivity and nutrient availability (Ullrich et al. 2001). High Hg methylation rates in aquatic systems were usually found in warmer months (Callister and Winfrey 1986; Hintelmann and Wilken 1995; Jackson et al. 1982; Korthals and Winfrey 1987). Moreover, an increase in temperature promotes microbial activity (Randall and Chattopadhyay 2013; Ullrich et al. 2001). The chemistry of mercury in sediments and its bioavailability have been actively studied in estuarine ecosystems, with special focus in areas where Hg contamination is moderate to high. Nonetheless, few studies have focused on how processes affect the fate and bioavailability of Hg in the sediment, its methylation, and the possible release of Hg and MeHg into the water column, occurring in low Hg-contaminated areas. Accordingly, the present study contributes to the understanding of the high variability reported in the literature, concerning biogeochemical processes within the Hg cycle in low contaminated areas of mesotidal estuaries. The work was conducted in two areas of Tagus Estuary (Portugal) with low levels of mercury contamination and subjected to different hydrodynamic effects. The main aims are to report (1) the processes involving Hg partition and transformations in bottom sediments and (2) the estimation of Hg species mobility and exchanges between sediment and water column.

2 Material and Methods 2.1 Site Description The Tagus Estuary (Fig. 1) is one of the most important estuaries in Europe, covering 325 km2 with ∼40% of intertidal area. It is a vertically well-mixed estuary with semidiurnal mesotidal regime (Taborda et al. 2009; Vaz et al. 2011). This estuary is well studied from hydrodynamic and sedimentary points of view (Mil-Homens et al. 2014; Taborda et al. 2009; Vale and Sundby 1987; Vaz et al. 2011). The combination of wind forcing, rising tide, and water exchange with the main channels controls the resuspension of particulate matter (Freire et al. 2007; Vale and Sundby 1987), which is associated to the distribution of Hg (Canário et al. 2008b). Mercury contamination in the Tagus Estuary was first reported by Figuères et al. (1985). High levels


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Fig. 1 Locations of sampling sites in Tagus Estuary: Alcochete (ALC) and Vale Frades (VF). Both sampling sites were located inside the Natural Reserve of Tagus Estuary (RNET)

Iberian Peninsula

Natural Reserve Tagus Estuary Lisbon

VF ALC

Atlanc Ocean

of Hg in sediments, suspended matter, and water were attributed to two industrial areas located in the north (North Channel) and south margins (Barreiro) in the vicinity of the Natural Park (RNET), a RAMSAR site. Notwithstanding the inactivation of the most industrial units in these areas, still the extent of the impact of Hg contamination in Tagus Estuary remains unclear, even though knowledge of Hg and MeHg contamination levels in those intertidal sediments was well documented (Canário et al. 2005, 2007a; Micaelo et al. 2003; Monteiro et al. 2016; Vale 1990). In this study, two sites located inside RNET were selected due to its high ecological value (Saldanha 1980; Dias and Marques 1999): Alcochete (ALC) (lat. 38° 45′ 40.2″ N; long. 8° 56′ 14.0″ W) and Vale Frades (VF) (lat. 38° 46′ 57.8″ N; long. 8° 55′ 54.4″ W), both in the vicinity of urban and agricultural areas. Their environmental conditions are different due to hydrodynamic and tidal effects, as well as the existence of nearby tributaries (e.g., Enguias and Sorraia rivers). In fact, the ALC site is located at the opening of the main bay of the estuary and is more affected by the tides, while the VF site is located in the inner part of the protected area of the park in a confluence of channels and therefore more protected from the tides. VF is indeed a nursery area for fish and for aquatic fauna (Dias and Marques 1999).

According to previous works, low/moderate Hg contamination was found in ALC ([Hg] < 1 μg g−1 and [MeHg] < 5 ng g−1) (Canário et al. 2005, 2007a; Monteiro et al. 2016), while VF was considered the most pristine area within the natural park ([Hg] < 0.5 μg g−1 and [MeHg] < 5 ng g−1) (Monteiro et al. 2016). 2.2 Sampling and Sample Processing In July 2010 and February 2011, two sampling campaigns were performed at each site (ALC and VF) in the summer and winter periods. Air temperatures were around 22 °C in summer, while during the winter fieldwork, temperatures were around 15 °C. GPS coordinates were used to ensure that the sediment cores collected at both periods were approximately in the same area (±1 m), therefore minimizing spatial variations within each site. Four sediment subcores were collected in each site, using a PVC corer (7 cm diameter). To avoid oxidation of reduced species, these cores were sliced in a N2 glove box inside a mobile laboratory immediately after sampling. Cores were divided into 1-cm-thick layers for the top 10 cm, and 2-cm layers for depths between 10 and 20 cm. Deeper sediments were sliced in 5-cm sections until the end of the core (summer cores—40 cm at ALC

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and VF; winter cores—30 cm at ALC and 40 cm at VF). For each layer, temperature (°C), pH, and potential redox (EH (mV)) were also measured in situ. Temperature was measured using a multiparametric probe (MultiLine P3, WTW), and pH was measured using a portable Methrom 704 pH meter with temperature control and glass combined electrode, calibrated with two buffer solutions (4.00 ± 0.02 and 7.00 ± 0.02 at 25 °C, Crison Instruments SA). Redox potential (EH) was also measured with a combined electrode Ag/AgCl-platinum, calibrated with a buffer for 220 ± 0.05 mV vs. Ag/ AgCl at 25 °C (Mettler Toledo). Values were corrected to the hydrogen reference electrode. Afterward, slices were hermetically sealed in leak-proof aciddecontaminated 250 mL polypropylene copolymer (PPCO) centrifuge flasks (Nalgene), refrigerated on ice chests, and transported in the dark to the laboratory within 2 h. Overlying water was collected at low tide at each site in the sediment/water interface, before core collection, using a decontaminated syringe. In the laboratory, pore waters were extracted from the bulk sediments by centrifugation at 8000 rpm for 45 min at +4 °C and immediately filtered through 0.45 μm cellulose acetate syringe filters in N2 chamber. Filtrate subsamples were stored in glass flasks, teflon flasks, and polypropylene tubes and acidified to pH < 2, respectively, for reactive dissolved mercury (RHgD) analysis (HNO3, Merck—Hg free), total Hg and MeHg analysis (HCl, Merck—Hg free), and total dissolved metals, chloride, and sulfate determinations (bidistilled HNO3). 2.3 Analytical Methods 2.3.1 Sediment Analysis Sediments were oven dried at 40 °C, disaggregated, and homogenized. Total determinations of Si, Al, Fe, and Mn were performed by mineralization of sediment samples with a mixture of acids (HF, HNO3, HCl), according to the method described by Rantala and Loring (1975). Metal concentrations were obtained by flame atomic absorption spectrometry (F-AAS) in a Perkin Elmer AAnalyst 100 using direct aspiration into a N2O-acetylene flame (Si and Al) or air-acetylene flame (Fe and Mn). Total carbon and total nitrogen contents were measured using a CHN Fissons NA 1500 Analyser, calibrated with sulfanilamide standards. Organic carbon was estimated by the difference between total carbon and


inorganic carbon. Total carbon was determined directly from dry sediments and the inorganic was measured after heating the samples at 450 °C for 2 h, in order to remove the organic fraction (Verardo et al. 1990). Total mercury (THgP) concentrations were determined by AAS using a silicon UV diode detector LECO AMA-254, according to Costley et al. (2000). Methylmercury was determined by alkaline digestion (KOH/MeOH), organic extraction with dichloromethane (DCM) preconcentration in aqueous sulfide solution, back-extraction into DCM, and quantification by atomic fluorescence spectrometry coupled with gaseous chromatography (GC-AFS) using an Agilent Chromatograph and a Millennium Merlin PSA detector (Canário et al. 2004). 2.3.2 Water Analysis Sulfate (SO42−) concentrations were determined by turbidimetry method using a Hitachi U-2000 Spectrophotometer providing a light path of 5-cm length (APHA 1995). Total dissolved Fe and Mn were determined by F-AAS using air-acetylene flame. Dissolved organic carbon (DOC) analysis was performed by high temperature catalytic oxidation (HTCO) using a commercial Shimadzu TOC-5000A analyzer (Benner and Strom 1993). Reactive dissolved mercury concentrations (RHgD) were determined using cold vapor atomic fluorescence spectroscopy (CV-AFS) on a PSA Merlin mercury system (Beckvar et al. 1996). Reactive mercury is reduced to elemental mercury (Hg0) by SnCl2 2% in a 10% HCl solution (Hg free). Volatile Hg0 is then purged from solution by an argon flux. For the quantification of RHgD concentrations, peak height corresponding to the maximum absorption is converted to concentration from a calibration curve. Total dissolved mercury (THgD) was determined following U.S. EPA method 1631 (US-EPA 2002) by CVAFS with a PSA Merlin mercury system. Methylmercury (MeHgD) was determined following U.S. EPA method 1630 (US-EPA 1998) by distillation of 50 mL subsamples, after the addition of 1% C5H9NS2⋅NH3 as a complexing agent. Mercury was ethylated with NaB(C2H5)4, purged with argon, and collected on TenaxTM traps, mercury species are separated with a GC, thermally desorbed, and pyrolytically reduced to Hg0 for detection of MeHg with a Brooks Rand Model III CV-AFS.


2.3.3 QA/QC and Statistical Analysis Analytical quality control was performed for all parameters with daily calibrations, including procedural blanks and sample replicates. Certified reference materials (CRM) and international certified standards (ICS) were checked and used to ensure the analytical accuracy of the procedures. For all metals investigated, the obtained concentrations were consistently within the ranges of the certified values (n = 8; p < 0.05). For solid sediments, analytical results were quality checked by analyzing Al 7.89 ± 0.75% (certified MESS-3, from the National Research Council Canada (NRCC), 8.59 ± 0.23%); Si 25.4 ± 2.9% (certified MESS-3, 27.7%); Fe 3.78 ± 0.56% (certified MESS-3, 4.34 ± 0.11%); Mn 293 ± 24 μg g−1 (certified MESS-3, 324 ± 12 μg g−1); Hg 2.84 ± 0.07 μg g−1 (certified PACS-2, from NRCC, 3.04 ± 0.20 μg g−1) and 0.083 ± 0.003 μg g−1 (certified MESS-3, 0.091 ± 0.009 μg g−1); and MeHg 76.2 ± 2.8 ng g−1 (certified BCR-580, from the Institute for Reference Materials and Measurements (IRMM), 75.5 ± 3.7 ng g−1) and 5.59 ± 0.35 ng g−1 (certified IAEA405, from the Mel International Atomic Energy Agency (IAEA), 5.49 ± 0.53 ng g−1). In water obtained, concentrations of THgD were 2.17 ± 0.3 ng L−1 (certified BCR579, from the Institute for Reference Materials and Measurements (IRMM), 1.9 ± 0.5 ng L−1). Recoveries of Hg ranged from 90 to 109% and for MeHg from 92 to 103%. Replicate samples were also used to assess variability of the data. Blanks were repeated every 20 samples in order to evaluate cross contaminations and to ensure that equipments are operating in the same conditions. The limits of detection (LOD) were calculated as three times the standard deviation from the blanks. Precision was better than 6.0% (n = 8), expressed as a percentage relative standard deviation (Miller and Miller 2010). Data were tested for normality, assessed using Shapiro-Wilks statistical test. The nonparametric Wilcoxon paired-sampled (T) test was also applied since some variables did not present a normal distribution. Spearman correlation coefficients were also computed (Miller and Miller 2010). 2.4 Calculations of Diffusive Fluxes In order to evaluate if diagenetic processes can lead to significant Hg and MeHg postdepositional redistribution and/or diffusion to the overlying water, diffusive

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fluxes of MeHgD and THgD across the sediment/water interface (SWI) were estimated applying Fick’s first law of diffusion (Berner 1980): ϕDw dC ð1Þ J ¼− dz θ2 where J is the diffusion flux, ϕ and θ are respectively porosity and tortuosity, Dw is the ionic/molecular diffusion coefficient in water at 25 °C, and dC/dz is the concentration gradient. We used dz = 1 cm (i.e., the average depth of the uppermost pore water sample). According to Berner (1980), porosity was calculated using the expression: ϕ¼

mw mw þ

ms 2:65

ð2Þ

where mw is the pore water mass and ms is the dried sediment mass. Tortuosity was estimated from porosity using Boudreau’s formulation (Boudreau 1996): θ2 ¼ 1−ln ϕ2 ð3Þ Dw, 25 °C = 1.2 × 10−5 cm2 s−1 was used for MeHgD (Hammerschmidt et al. 2004a), which assumes that MeHgD in pore water exists entirely as CH3HgSH0 (Dyrssen and Wedborg 1991) For THg D , D w, −6 cm2 s−1 was used, assuming that 25 °C = 2 × 10 inorganic Hg is bound to macromolecules in the colloidal size range, as reported by Gill et al. (1999). All Dw were corrected for temperature (t, °C) using the following relationship: Dw; 25 0C ¼ Dt ð1 þ 0:048 ð25 –t ÞÞ

ð4Þ

The temperatures used for the calculations of the fluxes were the ones measured in the field, i.e., 22 °C in summer and 15 °C in winter periods.

3 Results and Discussion 3.1 Sediment Characteristics Concentrations of Si normalized to Al content (Si/Al) were used as a proxy for grain size (Loring 1991). Si/Al ratios in the ALC ranging from 2 to 10 indicated that sediments are a mixture of fine and coarser particles. Sediments in VF showed that Si/Al ranging between 3


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and 6 indicated the presence of more fine-grained particles including silt and clay (Stumm and Morgan 1995). Redox potential (EH) in ALC in both seasons was positive in the topmost layers decreasing sharply to negative values at 7-cm depth (see Supplementary Information—SI-Fig. 1a). Below this layer, EH presented nearly constant values in each season (average −135 ± 18 mV in summer and −206 ± 10 mV in winter). Vertical profiles of EH in VF showed more anoxic conditions and were similar in both sampling periods (SI-Fig. 1b) ranging between −265 and −166 mV. Sediment cores from ALC presented a progressive pH increase with depth in both seasons (SI-Fig. 1c) with higher values in winter. A different pattern was observed in VF with a progressive pH increase with depth in winter but only from the top layer sediment until 7-cm depth; however, in summer, a near constant distribution of pH was measured (7.2 ± 0.1) (SI-Fig. 1d). Organic carbon content (Corg) decreased over depth in ALC for both seasons (Fig. 2a), ranging between 1.2 and 2.0% in summer and from 0.6 to 3.0% in winter, while in VF, minor differences were found in depth (average 1.8 ± 0.1% for both seasons) (Fig. 2b). Fig. 2 Concentrations of organic carbon (Corg, %) (a, b) in solid phase of sediments and of dissolved organic carbon (DOC, μM) (c, d) in sediment pore waters from Alcochete (ALC) and Vale Frades (VF) in summer and winter periods

DOC distribution in sediments showed high variability without any consistent trend between sites. Higher values of DOC were found in summer for both sites, ranging from 46 to 188 mg L−1 in ALC (Fig. 2c) and from 58 to 122 mg L−1 in VF (Fig. 2d). In the cold season, DOC concentrations decreased to an average value of 53 ± 21 mg L−1 in ALC and 25 ± 9.5 mg L−1 in VF. The DOC values obtained in this study are in agreement with those found in pore waters from noncolonized sediments by Canário (2004) in the same estuary; in Aveiro Ria, Portugal, by Otero et al. (2007); or in pore waters of the Yangtze Estuary, China (Wang et al. 2013). 3.2 Mercury and Methylmercury The Tagus Estuary has been for a long time considered one of the most contaminated by Hg in Europe (Canário et al. 2005; Figuères et al. 1985). However, less attention has been given to less contaminated areas and how the dynamics of the estuary may transport and influence other areas, with particularly focus on the Tagus Estuary Natural Reserve, a RAMSAR site. Thus, by sampling Solids

Dissolved DOC (mg L-1)

Corg (%) 1.0

2.0

3.0

0

4.0 0

5

5

Depth (cm)

Depth (cm)

ALC

0.0 0

10 15

100

200

300

10 15

20 20

25

a)

30

summer

c)

25

winter

1.0

2.0

winter

summer

DOC (mg L-1)

Corg (%) 0.0

3.0

0

4.0

100

200

300

0

0

Depth (cm)

VF

5 10

Depth (cm)

475

15 20 25

5 10 15

30 35

b)

40 summer

winter

d)

20 summer

winter


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sediment cores in two different areas in the reserve, the present study is essential not only for assessing the status of Hg contamination in the area but also, as stated above, inferred the transport from the contaminated areas and ultimately for ecological assessment of the RAMSAR site.

retention by salt marsh vegetation that surrounds this site with a lower hydrodynamic regime and also less affected by tidal propagation, compared to ALC (Canário et al. 2005, 2010; Guerreiro et al. 2015). Unlike VF cores, significant positive correlations between THgP concentrations and Si/Al ratios were found in ALC in both periods (r = 0.79, p < 0.05 in summer and r = 0.63, p < 0.05 in winter), pointing the influence of the nature of particles on Hg distribution in ALC cores. No correlations were found between THgP and Corg in both sites and periods suggesting that the variation of Hg in all sampling cores is more related with the mercury input history than the fluctuations of natural organic matter content. This is in line with other previous works in other areas of the Tagus Estuary (e.g., Canário et al. 2003) or in Vigo Ria in Spain (Canário et al. 2007b). In sediments from the ALC summer core, significant negative correlations between Fe, Mn, and THgP [r = −0.82, p < 0.05 for Fe and r = −0.70, p < 0.05 for Mn] (vertical profiles of Fe and Mn in SI-Fig. 2) were also observed. The reduction of Fe and Mn oxyhydroxides through organic matter oxidation in the sediment solid phase promotes Hg release into pore waters (Canário

3.2.1 Mercury in the Sediment Solid Phase Vertical distributions of total mercury (THgP) concentration in the solid fraction of the sediment in the ALC site were similar in both seasons (Fig. 3a, b). Also, in both seasons, a peak was observed between 13 and 15 cm depth, which is probably related with the highest historical contamination period reported during the 1980s (Figuères et al. 1985). In VF, Hg levels were almost constant (0.48 ± 0.04 μg g−1) with depth in the summer period (Fig. 3c). In the cold season, Hg levels (Fig. 3d) showed a major variation (0.54 ± 0.07 μg g−1). The historical Hg contamination peak found in ALC was not observed in VF. This may be due to the relative distance to the Hg anthropogenic source in the south shore of the estuary (Barreiro) and by the particle

Summer

Fig. 3 Concentrations of total mercury (THgP, μg g−1) and methylmercury (MeHgP, ng g−1) in solid phase of sediments from Alcochete (ALC) (a, b) and Vale Frades (VF) (c, d) in summer and winter periods

Winter

THgP (µg g-1)

THgP (µg g-1)

0.0 0.2 0.4 0.6 0.8 1.0

0.0 0.2 0.4 0.6 0.8 1.0 0

5

5

10

10

15 20

a)

25

Depth (cm)

Depth (cm)

ALC

0

30

15 20

b)

25 30

THgP

35

THgP

35

MeHgP

40

MeHgP

40 0.0 1.0 2.0 3.0 4.0 5.0

0.0 1.0 2.0 3.0 4.0 5.0

MeHgP (ng g-1)

MeHgP (ng g-1) THgP (µg g-1)

THgP (µg g-1)

0.0 0.2 0.4 0.6 0.8 1.0

0.0 0.2 0.4 0.6 0.8 1.0 0

5

5

10

10

15 20 25 30 35

c) THgP MeHgP

Depth (cm)

Depth (cm)

VF

0

15 20 25 30 35

d) THgP MeHgP

40

40 0.0 1.0 2.0 3.0 4.0 5.0 MeHgP (ng g-1)

0.0 1.0 2.0 3.0 4.0 5.0 MeHgP (ng g-1)

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et al. 2003). In fact, high concentrations of Fe and Mn, as the resulting products of early diagenetic processes, were found in pore waters, which can be eventually released to the overlying water by diffusion (Canário et al. 2007c). Particulate methylmercury (MeHgP) profiles presented the highest concentrations above 7 and 15 cm depth in ALC and VF, respectively, for both periods (Fig. 3). In the summer, the higher values of MeHgP were determined, ranging between 0.7 and 4.1 ng g−1 in ALC (Fig. 3a) and from 0.23 to 4.3 ng g−1 in VF (Fig. 3c). It should be noted that MeHgP peaks in ALC, in both seasons (Fig. 3a, b), closer to the SWI, contrasting with the THgP peaks, corresponding to in situ Hg methylation, as observed by Schafer et al. (2010). Below 7 cm depth (and 15 cm for VF in summer, Fig. 3c), the levels were below the LOD (0.1 ng g−1) suggesting minor or even the absence of Hg methylation at the bottom sediment. It should be noticed that MeHg concentrations found in these low contaminated areas are higher than those reported for the Guadiana Estuary (below 0.27 ng g−1) by Canário et al. (2007c). Higher values have been obtained in Aveiro Ria (12.4 ng g −1 ) (Ramalhosa 2002) and Sado Estuary (2.3–5.9 ng g−1) (Canário et al. 2007c), both in Portugal that also have record of historical Hg contamination. The proportion of MeHg in solid fraction (MeHgP/ THgP in %) in both sites and seasons was below 1% (Table 1) that is in line with the values found by Canário et al. (2005, 2007a) in other sites from the Tagus Estuary. Similarly, lower values were reported for the Grado Lagoon (0.001–0.20%) by Covelli et al. (2008), Marano and Grado Lagoon (0.01–0.15%) by Emili et al. (2012), Venice Lagoon (0.05–0.3%) by Bloom et al. (2004), Thau Lagoon in France (0.02–0.80%) by Muresan

et al. (2007), or also in Soca/Isonzo River in the northern Adriatic Sea ( 0.05), suggesting that MeHgP distribution is not influenced by the sediment nature. Furthermore, concentrations of MeHgP were apparently not related with the concentrations of THgP in the sediment since no significant correlation was found (p > 0.05). Therefore, other factors like organic or inorganic complexing agents (Mikac et al. 1999; Ullrich et al. 2001) or sorption/desorption MeHg mechanisms in sediments (Canário et al. 2008a) should be considered. No correlations were observed between MeHgP and organic carbon suggesting that Hg methylation may be more influenced by the availability of labile organic substrates, a fraction of the total organic carbon (Ravichandran 2004).

Table 1 Average concentrations of mercury species in solids [THgP (μg g−1), MeHgP (ng g−1), and MeHgP/THgP (%MeHgP)] and in dissolved fraction [THgD, RHgD, and MeHgD (all in ng L−1), RHgD/THgD (% RHgD), and MeHgD/THgD (%MeHgD)] of

sediments and partitioning coefficients (log KD) of THg and MeHg (L kg−1) from Alcochete (ALC) and Vale Frades (VF) in summer and winter periods

Site

Period

Solids THgP (μg g−1)

MeHgP (ng g−1)

MeHgP (%)

THgD RHgD (ng L−1) (ng L−1)

ALC Summer 0.53 ± 0.17 0.97 ± 1.46 0.19 ± 0.28 15 ± 10 Winter VF

0.43 ± 0.13 0.15 ± 0.13 0.04 ± 0.03 24 ± 11

Summer 0.48 ± 0.04 Winter

Log KD (L kg−1)

Dissolved fraction

1.5 ± 1.3

0.32 ± 0.29 36 ± 10

0.54 ± 0.07 0.22 ± 0.33 0.04 ± 0.05

8±7

MeHgD (ng L−1)

RHgD (%)

MeHgD (%)

THg

MeHg

30 ± 16

27 ± 26

4.7 ± 0.3 2.3 ± 0.6

5.4 ± 5.8

2.4 ± 1.5

2.2 ± 1.6

0.48 ± 0.69

0.95 ± 0.61

4.4 ± 2.0

2.7 ± 1.9

12 ± 5.8 4.1 ± 0.1 2.5 ± 0.4

4.4 ± 3.7

1.6 ± 1.2

48 ± 19

23 ± 18

10 ± 9.6 3.0 ± 5.1 4.3 ± 0.2 3.1 ± 0.6 5.0 ± 0.4 2.4 ± 0.8


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3.2.2 Mercury in the Sediment Pore Waters

winter cores were lower than the ones found by Canário et al. (2008b) in contaminated waters from the same estuary although in agreement with the ones reported earlier by the same author in other sites of the Tagus Estuary (Canário 2004). Higher MeHgD levels were found in summer cores, with the highest concentrations observed in the upper sediment layers above 5 cm depth in ALC (maximum concentration of 6.9 ng L−1, Fig. 4e) and 15 cm depth in VF (maximum concentration of 7.7 ng L−1, Fig. 5e). In the winter season, MeHg D was below the LOD (0.02 ng L−1) in depths higher than 5 and 15 cm at ALC (Fig. 4f) and VF (Fig. 5f), respectively. Seasonal variations of MeHgD vertical distribution were also observed in both sites. MeHgD data in this study (below 8 ng L−1) were similar to those found in the Gironde Estuary, France (below 5 ng L−1) (Schafer et al. 2010), and in Marano and Grado Lagoons, Italy (below 8 ng L−1) (Covelli et al. 2008; Hines et al. 2012). In both sites, the methylation horizon moved upwards in the sediment during winter, being marked in ALC from the SWI to 5 cm depth. Strong positive correlations were found between MeHgD and MnD in ALC and VF in winter (r = 0.88, p < 0.05 and r = 0.78,

Total dissolved mercury (THgD), reactive dissolved mercury (RHg D ), and dissolved methylmercury (MeHgD) vertical profiles for both sites and periods (Figs. 4 and 5) showed seasonal variations. The highest THgD and lowest RHgD concentrations were found in ALC winter (Fig. 4b, d) and in VF summer cores (Fig. 5a, c). Vertical distributions for THgD and RHgD were similar in shape for ALC in summer (Fig. 4a, c) and VF in winter (Fig. 5b, d). The strong correlations between THgD and RHgD (r = 0.850, p < 0.05 for ALC in summer and r = 0.980, p < 0.05 for VF in winter) suggest an equilibrium between the Hg species (e.g., Faganeli et al. 2003). Otherwise, no correlations (p > 0.05) were found in ALC in winter and VF in summer, which presented respectively much lower values of % RHgD, below 37% and 8%, respectively. THgD concentrations obtained in this study (below 62 ng L−1) were much lower than the ones found previously in other areas of the Tagus Estuary (Canário 2004). Interestingly, RHgD with maximum concentrations around 15 ± 3 ng L−1 in ALC summer and VF THgD (ng L-1) 40

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Fig. 4 Concentrations of total mercury (THgD, ng L−1) (a, b), reactive mercury (RHgD, ng L−1) (c, d), and methylmercury (MeHgD, ng L−1) (e, f) and contribution of MeHgD to THgD (%)

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(g, h) in sediment pore waters from Alcochete (ALC) in summer and winter periods. The bare point corresponds to the overlaying water and the full points correspond to pore waters


Page 10 of 17 THgD (ng L-1) 40

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Fig. 5 Concentrations of total mercury (THgD, ng L−1) (a, b), reactive mercury (RHgD, ng L−1) (c, d), and methylmercury (MeHgD, ng L−1) (e, f) and contribution of MeHgD to THgD (%)

(g, h) in sediment pore waters from Vale Frades (VF) in summer and winter periods. The bare point corresponds to the overlaying water and the full points correspond to pore waters

p < 0.05, respectively). Since Fe and Mn oxyhydroxides can represent important particulate Hg carrier phases in estuarine environments (Laurier et al. 2003; Muresan et al. 2007; Schafer et al. 2010; Turner et al. 2004), their reductive dissolution releases the associated Hg species into the sediment pore waters together with dissolved Fe and Mn (SI-Fig. 3) Thus, it can be hypothesized that an increase in MeHgD in this cores is coupled with Hg released from reactive oxyhydroxides, increasing Hg bioavailability for methylation. Additionally, the observed higher FeD and MeHgD concentrations in the top layers of the ALC summer core and VF cores in both seasons suggest that iron-reducing bacteria may also promote the production of MeHgD. According to Mason and Fitzgerald (1990), RHgD may be a proxy of the Hg available for methylation, Hg0 formation, and other conversion processes. Reactive dissolved Hg includes volatile Hg species and Hg species easily reduced by SnCl2, as inorganic and organic Hg labile complexes (Beckvar et al. 1996). The observed decoupling between MeHgD and RHgD in both seasons for ALC and in summer VF core indicates that Hg availability is not the sole driver for Hg methylation. In these cases, Hg methylation was mediated by

different factors that influenced the rate process, where RHgD mobilization in the sedimentary column by downward diffusion proceeded without having reached equilibrium and/or MeHg is due to in situ Hg methylation. In VF winter sediments, significant positive correlations were found between MeHgD and RHgD (r = 0.681, p < 0.05) and between MeHgD and THgD (r = 0.711, p < 0.05). This indicates that total, reactive, and methyl-Hg species were in equilibrium in VF winter core and also coupled with Mn oxyhydroxides, governing the Hg availability for methylation. Vertical distributions of MeHgD/THgD (%MeHgD) are shown in Figs. 4 and 5 for the ALC and VF sites, respectively. Despite the occurrence of %MeHgD between 30 and 95% in some layers, mean values in pore waters for both seasons were less than 30% (Table 1), as often found in the literature (e.g., Bratkič et al. 2013; Gagnon et al. 1997; Muresan et al. 2007; Schafer et al. 2010). Higher values of %MeHgD in sediments were also found in Truckee River (67 ± 26%) (PizarroBarraza et al. 2014) and in Carson River (20–80%) (Bonzongo et al. 1996), both in the USA. In the ALC and VF summer upper layers (above 6 cm), both MeHg D and %MeHg D values were


Page 11 of 17 475

generally high and related with an excess of sulfate, thus suggesting in situ Hg methylation by SRB. In ALC and VF winter cores, significant positive correlations (r = 0.64, p < 0.05 and r = 0.80, p < 0.05, respectively) between MeHgD and SO42− concentrations were found. Therefore, it can be hypothesized that an increase of dissolved sulfate concentrations (SI-Fig. 4) may favor MeHgD production through the enhancement of the SRB activity (Gilmour et al. 1992). In summer cores, higher MeHgD values are coupled to a slight decrease of pH and with the consequent increase of DOC protonation, suggesting that Hg is more available to methylating bacteria (Barkay et al. 1997; Miskimmin et al. 1992). The increase of DOC values in ALC from winter to summer was coupled to a similar increment of %MeHgD. This was probably related to the increase of temperature in summer and a slight pH decrease, also. Contrarily, in VF, an increase of DOC values and a reduction of %MeHgD were observed between the two seasons, probably due to complexation of Hg mercury with DOC that limited the amount of bioavailable Hg for uptake, reducing MeHg production (Ravichandran 2004). The correlations between DOC and MeHg concentrations were not significant (p > 0.05); therefore, the previous trends indicate a seasonal pattern related to the patchiness of sediments within the estuary, linked with the nature of DOC, in particular with the humic substances. This indicates that an increase in %MeHgD is inversely proportional to the concentrations of humic substances. Since these sites are surrounded by agricultural and forest areas, the large input of DOC, largely formed by humic substances, must be taken into account. Moreover, the relation observed between the elevated %MeHgD and the highest values of RHgD (Table 2) in VF winter sediments also Table 2 Diffusive fluxes (J) of THgD and MeHgD (ng m−2 day−1) in dissolved fraction of sediments from Alcochete (ALC) and Vale Frades (VF) in summer and winter periods (negative fluxes indicate diffusion from pore water to overlying water) Site

Period

T (°C)

J (ng m−2 day−1) THgD

ALC VF

MeHgD

Summer

22

4.51

2.37

Winter

15

−1.62

−1.30

Summer

22

−7.80

−8.43

Winter

15

−3.76

18.5

suggested that this mercury specie was the major factor in the methylation of mercury, in this period. 3.3 Partition of Mercury in Sediments The controls on Hg bioavailability to methylating bacteria and the availability of MeHg to organisms at the base of the food web (Rothenberg et al. 2008) may be explained by the partition of Hg species between solid/ dissolved phases. In the summer period, a significant negative correlation (r = −0.667, p < 0.05) between THgP and THgD was found in the layers below 6 cm depth in ALC sediments. These results may be due to dissolution/ desorption and precipitation/adsorption processes out of equilibrium (Muresan et al. 2007). In the winter period, ALC sediments presented a significant positive correlation between THg P and THg D (r = 0.551, p < 0.05). This behavior is in agreement with the observations of Muresan et al. (2007), who reported that total Hg in solid phase and pore waters of sediments were in equilibrium. In sediments from VF for both periods, no significant correlations (p > 0.05) were found between THgP and THgD. In contrast to ALC sediments, this suggests that THg in solid phase and pore waters of VF sediments were not in equilibrium. Significant positive correlations between MeHgP and MeHgD were found in ALC sediments in summer (r = 0.694, p < 0.05) and winter (r = 0.937, p < 0.05), suggesting equilibrium conditions between both fractions of MeHg. A similar pattern was observed at VF in summer, below 6 cm depth (r = 0.799, p < 0.05), where the equilibrium between solid and dissolved MeHg was evident. No correlation was found between MeHgP and MeHgD in sediments of VF in winter, in contrast to what was shown before for THg between solid phase and pore waters. The water-solid partitioning of Hg is described in terms of log KD, where KD defines the ratio between solid and dissolved phase constituents, and is expressed in liters per kilogram (Hammerschmidt et al. 2004). Low KD values are associated with an enhanced Hg release into pore waters from the solid phase, while higher values are indicative of a stronger binding of Hg species to the sedimentary matrix. Log KD showed different values between periods, for ALC and VF (Table 1). The highest values of log KD (THg) were found in summer at ALC (5.3 L kg−1) and in winter at VF (5.7 L kg−1). Instead, the lowest values were found

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at VF in the warm season (3.9 L kg−1) and in ALC in the coldest one (4.1 L kg−1). The decrease of log KD (THg) values at VF in summer and at ALC in winter reflected the enhanced release of Hg into pore waters. A different pattern was observed for log KD of MeHg, where the highest values (3.7 L Hg−1, in both sites) were observed in winter, and lowest in the summer (3.2 Hg−1, in both sites) periods. Seasonal log KD (THg) values for ALC and VF are lower than those reported by Emili et al. (2014) for the Gulf of Trieste, northern Adriatic Sea, and in agreement with those reported by Emili et al. (2012) for Marano and Grado Lagoons, Italy. According to Marvin-Dipasquale et al. (2009), the decrease of organic matter content may be coupled with lower log KD values. This behavior was not observed for THg log KD values but only for MeHg log KD values. The absence of correlations between log KD (THg) and organic carbon in both sites and periods can be explained by the possibility that only a limited fraction of organic matter (OM) other than organic carbon is relevant for Hg complexation (Hollweg et al. 2009). However, the role of dissolved organic matter cannot be neglected (Ravichandran 2004), once DOM is known to participate in pore water Hg speciation and thus to affect Hg partition. Thus, log KD values in sediments for THg may result from the simultaneous occurrence of the following processes: (1) microbial degradation of particulate organic matter (POM) that contributes to liberate Hg to both the water column (e.g., Cossa and Noël 1987) and sediment pore water, (2) the diagenetic destruction of Hg with high affinity for OM and sulfur in the pore water or solid phase, and (3) the presence of Hg-rich colloids (e.g., Rothenberg et al. 2008). A significant positive correlation (r = 0.710, p < 0.05) was found between Corg and log KD (MeHg) in ALC summer core. A similar positive correlation (r = 0.620, p < 0.05) was also observed in VF summer sediments. Additionally, in winter sediments of VF, log KD (MeHg) was significantly positively correlated with DOC (r = 0.721, p < 0.05). These findings suggested that OM plays an important role in MeHg partition, which may also be linked with changes in temperatures. 3.4 Diffusive Fluxes of Mercury Species The estimated diffusive fluxes for both THgD and MeHgD in our samples are presented in Table 2. The obtained THgD and MeHgD diffusive fluxes were


generally higher (in absolute value) in VF than in ALC for both periods, due to high THgD and MeHgD concentration gradients in the sediment/water interface. Thus, active transport processes between these interfaces occur in the topmost sediment layers. Diffusive fluxes of THgD and MeHgD in the warmer period were generally higher with the exception of MeHgD flux in VF in winter (up to 5 times higher than the corresponding in summer). More active biogeochemical processes occurring in the topmost sediment layer and in the water column, which tend to create higher gradients, may explain summer diffusive fluxes. On the other hand, winter MeHgD diffusive flux at VF may be a consequence of the decrease in mobility of MeHgD, due to its association in the solid fraction of the sediment and/or by substantial decrease in Hg methylation at this site. The balance between methylation and demethylation processes must also be considered. An interesting result is the different abilities of sediments to export or retain Hg and MeHg in ALC and VF sites, in both seasons. Diffusive fluxes of THgD and MeHgD from pore waters to overlying waters were observed in winter in the ALC site. The opposite was observed in the summer period. These results indicate that ALC sediments were both a source of Hg for the water column (in winter) and an Hg sink (in summer). Therefore, temperature seems to be a main driver in the export/retention of Hg in ALC sediments. In the VF site, diffusive THgD fluxes indicated that sediments export Hg during summer and winter periods. However, MeHgD diffusive fluxes showed opposite behavior. In summer, sediments exported MeHg to the water column but were a sink in winter. The temperature seems to stimulate THgD and MeHgD exportation in summer. Low temperatures promote less THgD exportation from sediments and also the retention of MeHgD in sediments. The results estimated in our sites for THgD diffusive fluxes were in the same range as those found in Gironde Estuary, France (Schafer et al. 2010), and Mugu Lagoon, USA (Rothenberg et al. 2008). The results for MeHgD diffusive fluxes were similar to those estimated both in the Gulf of Trieste, Italy (Covelli et al. 1999), and in Marano and Grado Lagoons, Italy (Covelli et al. 2008; Emili et al. 2012). Assuming that each period (summer and winter) combined is on average half of the year and that sediments were covered by water 12 h a day (semidiurnal tidal cycle), we calculated the total area of each site (ALC 6.6 × 105 m2; VF 2.08 × 106 m2) and estimated


that ALC is a sink of both HgD and MeHgD with an average Bdepositon^ of ca. 170 mg of Hg and 60 mg of MeHg per year. However, the VF area is also a sink of MeHg with an average deposition of ca. 1900 mg year−1 but is a source of HgD with an exportation of ca. 2000 mg year−1. Although molecular diffusion played an important role in Hg/MeHg transport in the sediment/water interface, we cannot exclude the bioturbation effect that can enhance this benthic flux by the irrigation of infauna (e.g., Fanjul et al., 2011) and the Hg transport through an advection mechanism (e.g., Zagar et al. 2007). In fact, in a recent study made by Cesário et al. (paper in preparation), molecular diffusion only accounted for less than 4% of the Hg transported from the sediment to the water column. This result therefore indicates that the estimated amounts of Hg/MeHg transport annually in these sites may be underestimated and that Hg is more mobile even in remote areas of the Tagus Estuary.

4 Conclusions Mercury sediment dynamics were evaluated in a less exposed area (ALC and VF sites) of the Tagus Estuary, inside a RAMSAR Natural Park (RNET). In spite of the relatively remote location of the sampling sites, far from the historical anthropogenic sources of Hg, this study clearly pointed out the contamination resulting from the internal estuarine water circulation. This contamination was slightly higher in ALC, more influenced by the water circulation in the estuary (max 0.87 μg g−1 Hg), as compared to VF in the inner part of the natural park (max 0.73 μg g−1 of Hg). In both sites, the highest temperature during warmer periods with the consequent enhancement of microbial activity appears to increase the %MeHg within a narrow temperature range. While particle size and nature appeared to drive Hg distribution in ALC, this is not true in VF for Hg and in both sites for MeHg. MeHg distribution in the sediment column is more related to in situ methylation and its distribution due to internal transport/retention processes. Interestingly, it should be noted that these sites appear to have more favorable environmental conditions to methylate available Hg, particularly in sediment pore waters (where %MeHg reached 95%). ALC sediments were both a source of Hg and MeHg to the water column (in winter) and sink (in summer), whereas VF sediments

Page 13 of 17 475

were a source of Hg in both periods. However, those sediments (VF site) have the capacity to export MeHg in summer, although acting as sink in winter. Considering the estimated annual Hg and MeHg budgets in both sites (deposition of ca. 170 mg year−1 of Hg and of ca. 60 mg year−1 of MeHg in ALC; exportation of ca. 2000 mg year − 1 of Hg and deposition of ca. 1900 mg year−1 of MeHg), ALC sediments behaved as a sink of both Hg and MeHg, while VF sediments were a sink of MeHg but a source of Hg. This study suggests that the morphology and hydrodynamic regimes of the Tagus Estuary influence Hg processes and dynamics in two more confined and adjacent areas of the estuary. Further studies concerning the Hg hydrodynamics and particulate transport from contaminated areas inside the Tagus Estuary should be conducted in order to assess Hg transport and its effects on less contaminated areas with direct consequences to wildlife and ecosystem services. Acknowledgments This work was performed under the projects PROFLUX—Processes and Fluxes of Mercury and Methylmercury in a Contaminated Coastal Ecosystem, Tagus Estuary, Portugal (PTDC/MAR/102748/2008), and PLANTA—Effect of saltmarsh plants on mercury methylation, transport and volatilization to the atmosphere (PTDC/AAC-AMB/115798/2009), both funded by the Portuguese Foundation for Science and Technology (FCT). The authors would like to acknowledge Reserva Natural do Estuário do Tejo (RNET) for the permission and support to authorize this work inside the protected areas. Rute Cesário would also like to acknowledge FCT for the funding of her PhD grant (SFRH/BD/86441/2012).

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