Submarine Groundwater Discharge as a Source of Mercury in the Bay of Puck, the Southern Baltic Sea
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Both groundwater flow and mercury concentrations in pore water and seawater were quantified in the groundwater seeping site of the Bay of Puck, southern Baltic Sea. Total dissolved mercury (HgTD) in pore water ranged from 0.51 to 4.90 ng l−1. Seawater samples were characterized by elevated HgTD concentrations, ranging from 4.41 to 6.37 ng l−1, while HgTD concentrations in groundwater samples ranged from 0.51 to 1.15 ng l−1. High HgTD concentrations in pore water of the uppermost sediment layers were attributed to seawater intrusion into the sediment. The relationship between HgTD concentrations and salinity of pore water was non-conservative, indicating removal of dissolved mercury upon mixing seawater with groundwater. The mechanism of dissolved mercury removal was further elucidated by examining its relationships with both dissolved organic matter, dissolved manganese (Mn II), and redox potential. The flux of HgTD to the Bay of Puck was estimated to be 18.9 ± 6.3 g year−1. The submarine groundwater discharge-derived mercury load is substantially smaller than atmospheric deposition and riverine discharge to the Bay of Puck. Thus, groundwater is a factor that dilutes the mercury concentrations in pore water and, as a result, dilutes the mercury concentrations in the water column.
KeywordsGulf of Gdańsk Loads Mercury Submarine groundwater discharge Seepage water
Submarine groundwater discharge (SGD) is one of the water pathways connecting land and ocean in the global water cycle. It has been recognized as an important factor influencing coastal zones (Burnett et al. 2006; McCoy and Corbett 2009; Szymczycha et al. 2012). Groundwater, similarly to surface water, flows along a hydraulic gradient, thus entering directly into the sea, wherever a coastal aquifer exists. While the contribution of groundwater discharges can be comparatively minor, particularly in the areas dominated by large rivers flows, studies have shown that groundwater is an important fraction of freshwater outflow (Burnett et al. 2006). Moreover, as groundwater is most often enriched with chemical constituents, it can be a source of substantial loads of nutrients, trace metals, and organic compounds (Charette and Sholkovitz 2006; Szymczycha et al. 2012).
The knowledge of SGD nutrient loads as a factor enhancing coastal eutrophication is relatively well established (Valiela et al. 2002). Moreover, there is a growing interest in the concentrations and loads of trace constituents delivered to the coastal zone through SGD (Charette and Sholkovitz 2006; Beck et al. 2010). Studies concerning mercury concentrations in seepage water and their impact on the marine ecosystem have recently indicated the ecological and geochemical importance of SGD. In this respect, some of the recent studies devoted to mercury distribution in SGD-impacted areas indicate enriched concentrations of mercury there. Laurier et al. (2007) reported an increase in total dissolved mercury (HgTD) in seawater and enhanced mercury uptake by mussels in the groundwater-impacted area (the Pays de Caux, France). This phenomenon was explained by high HgTD concentrations in estuarine water that resulted in mercury partitioning to the dissolved or colloidal phase in response to changes in salinity and/or turbidity. Bone et al. (2007) reported that the main driver of mercury flux was the low organic carbon content of the aquifer sediments and emphasized the complexity of mercury transport within groundwater systems. Tiffreau et al. (1995) suggested that mercury will desorb from metal (hydr)oxides at increasing concentrations of chloride through the formation of soluble Hg-Cl complexes. Lamborg et al. (2004) assumed that dissolved organic carbon (DOC) should exert a greater impact on mercury speciation than chloride. Thus, there is a need for data characterizing both mercury concentrations in seepage water and mercury behavior upon the mixing of groundwater and seawater.
Research on mercury concentration levels in the Gulf of Gdańsk has been performed for many years. It has focused on determining mercury concentrations in surface water, sediments, and marine fauna and flora (Pempkowiak et al. 1998; Bełdowski and Pempkowiak 2003, 2007; Kuss and Schneider 2007; Bartnicki et al. 2009; Pohl and Hennings 2008; Bełdowski et al. 2009; Saniewska et al. 2010). The main sources of mercury in the Gulf of Gdańsk and the Bay of Puck were identified to be atmospheric deposition (Bełdowska et al. 2012), river discharge (Pempkowiak et al. 1998), shipyards, harbors, wastewater treatment plants, and the municipal areas of Gdańsk, Gdynia, and Sopot. However, there are no available data concerning mercury fluxes via SGD in the Gdańsk Basin or for the entire Baltic Sea. This paper reports on the results of studies on mercury distribution in seepage water discharged to the Bay of Puck, southern Baltic Sea. In addition, HgTD concentrations in the groundwater and seawater were determined. HgTD flux via SGD into the Bay of Puck was estimated and compared to other mercury fluxes.
2 Materials and Methods
2.1 Study Area
Special care was taken to prevent contamination of collected samples. All components of the groundwater lances used for collecting pore water samples were soaked in a bath, filled with 3 M HNO3, for 1 day. After acid treatment, samplers were rinsed five times with MilliQ water, including ports and tubing. The entire equipment was wrapped in a number of layers of heavy plastic sheeting for transport. The groundwater lances were installed in the sediment and then allowed to equilibrate for ~24 h, with the free sample tube ends. After the equilibration period, the sample tubes were each attached to dedicated, acid-washed, all-PE 50-ml syringes with short sections of acid-cleaned Teflon tubing. Each syringe, in turn, was detached and connected to an acid-washed 25-mm diameter, 0.45-μm pore size, polypropylene syringe filter. A total of 35 ml of pore water was collected from several depths (0, 4, 8, 12, 16, 24, and 30 cm) below sediment–water interface. The collected water samples were divided into separate vials to be analyzed for the following components: HgTD, DOC, and dissolved manganese (Mn). The samples meant for HgTD analysis were transferred into acid-washed borosilicate test tubes with Teflon-lined caps. Samples meant for Mn analysis were transferred into acid-washed 20-ml LDPE bottles and acidified to pH < 2 with supra pure 0.1 N HNO3. Samples used for DOC analysis were transferred to parched glass vials and acidified with 40 μl of concentrated HCl. Groundwater samples from deep (water) wells and seawater samples were collected into vials, prepared in the same manner as the borosilicate bottles for HgTD analysis.
The samples for HgTD analysis were oxidized by BrCl and then pre-reduced with hydroxylamine hydrochloride solution 1 h prior to analysis by CV-AFS (TEKRAN 2600, Canada), according to US EPA method 1631 (US EPA 2002). Quality control included the analysis of blanks (n = 5) and estimating accuracy and precision based on the analysis of water samples (n = 5) (groundwater, seepage water, and seawater) spiked with mercury nitrate in the range of 0.5–2.5 ng Hg l−1. Adequate precision (6 %, given as relative standard deviation (RSD)) and recovery (95 ± 1 %) were obtained throughout the study. Moreover, during each sampling campaign, procedural blank samples (n = 3) were run. The obtained HgTD concentrations of the procedural blank samples were in the range 0.21–0.24 ng Hg l−1 and never exceeded 10 % of concentrations measured in the actual samples. The detection limit of the method used for HgTD analysis was equal to 0.2 ng Hg l−1.
Concentrations of Mn were determined by ICPMS (Elan 9000, Perkin Elmer). Analysis of standard reference material (SLEW 3) and groundwater samples spiked with Mn (5 μg l−1) served as a quality check. Average recovery of Mn was 103 %, and precision given as RSD was 0.3 % (n = 5).
Samples for DOC were analyzed with UV/persulfate oxidation and non-dispersive infrared detection (HyPerTOC, Thermo Electron Corp.). The limit of detection of the method was smaller by an order of magnitude than the measured concentrations. The precision of the results given as RSD was better than 2 % DOC, while recovery tested against the SGW standard was 96 %.
3.1 Salinity Distribution in the SGD-Impacted Area
Samples of 12 pore water profiles were collected in the groundwater-impacted area (GIA) and an additional profile in an area without apparent groundwater impact. The pore water salinity profiles are presented in Fig. 2. Generally, in the GIA profiles, salinity decreased with depth. The GL I 4.11.09 and GL I 5.11.09 profiles present salinity decrease from 7.1 and 6.9 to 2.8 and 2.1, respectively. In profiles GL II 4.11.09 and GL II 5.11.09, salinity decreased from 7 and 7.1 to 4 and 5.2, respectively, while in the GL II 28.02.10 and GL II 1.03.10 profiles, salinity decreased from 7.1 and 6.3 to 0.5 and 0.3, respectively. At the GL I 28.02.10 profile, salinity decreased from 6.7 to 2.3, while on the following day, salinity deceased to 0.3. In May 2010, similarly to November 2009, two groups of salinity profiles can be distinguished. Samples collected at location GL I show a salinity decrease from 7.1 to about 2.6, while samples collected at location GL II show a salinity decrease down to 0.5. The salinity profile in the location apparently unaffected by seepage discharges, GL S 7.05.10, shows constant salinity values of about 7 (Fig. 2). The changes in salinity distribution in GIA are caused by seawater intrusion into the sediment due to wave action and hydrodynamic conditions at the time of sampling (Szymczycha et al. 2012). Seawater can penetrate seepage sites to depths of tens of centimeters due to thermohaline density effects (Bokuniewicz 1992). Previous studies of SGD sites have indicated sediment pore water salinity profiles similar to those described here (Beck et al. 2007a; Beck et al. 2010; Pempkowiak et al. 2010). Therefore, for practical purposes, during this study, pore water samples characterized by salinity <1 were attributed to groundwater. Those characterized by salinity in the range 1 < S < 6.9 were attributed to seepage water (mixture of seawater and groundwater), and the ones with salinity S ≥ 7, to seawater.
3.2 Mercury Distribution in the SGD-Impacted Area
Concentrations of HgTD in the collected water samples
HgTD concentrations (ng l−1)
Number of samples
S > 7
S > 6.9
1 < S < 6.9
S ≤ 1
Groundwater from wells
0.5 < S
DOC distribution (Fig. 3) in the pore water also indicated the mixing of the two end-members (seawater and groundwater). Both end-members of the system represented considerably differing concentrations of organic substances. DOC distribution in pore water profiles was characterized by lower DOC concentrations in the uppermost 15-cm depth in the sediment and higher DOC concentrations in subsurface, lower pore water samples. Average DOC concentrations increased from 4.10 ± 0.05 mg l−1 to 5.24 ± 1.06 mg l−1 in November 2009, from 4.47 ± 0.41 mg l−1 to 5.69 ± 1.17 mg l−1 in February 2010, and from 4.38 ± 0.36 mg l−1 to 6.99 ± 1.28 mg l−1 in May 2010. Dissolved Mn concentrations in pore water profiles (Fig. 3), similarly to DOC, increased with increasing depths. Average Mn concentrations in November 2009 increased from 1.10 ± 0.33 μg l−1 to 87.31 ± 59.22 μg l−1; in February 2010, Mn concentrations increased from 2.53 ± 1.13 μg l−1 to 103.27 ± 5.75 μg l−1; and in May 2010, Mn concentrations increased from 13.94 ± 12.50 μg l−1 to 129.10 ± 39.71 μg l−1. Thus, the selected profiles (GL II 28.02.10, GL I 1.03.10, GL II 1.03.10, GL 5.05.10, GL II 5.05.10, GL 6.06.10, and GL II .6.06.10) all illustrate significant changes in HgTD, DOC, and Mn concentrations starting from a depth of 15 cm in the sediment upward, suggesting a significant influence of the mixing process between seawater and groundwater on the distribution of the measured constituents in pore water of the GIA.
Thus, the groundwater impact on HgTD distribution cannot be fully explained through the observation of geochemical components measured in this study (salinity, DOC, and Mn), suggesting that other processes still drive the distribution of mercury in sediment pore water in the study site. The groundwater–seawater mixing interface should be further investigated to understand the driving processes of mercury distribution.
3.3 The Dissolved Mercury Fluxes to the Study Area Via SGD
HgTD concentrations (Table 1) and the groundwater discharges established earlier (Szymczycha et al. 2012) were used to calculate mercury fluxes via SGD. The SGD rates were equal to 18.4 l day−1 m−2 in November 2009, 3.0 l day−1 m−2 in February 2010, and 3.6 l day−1 m−2 in May 2010. The HgTD fluxes calculated as a product of the groundwater discharge rate and the measured concentrations of the dissolved mercury were equal to 9.70 ± 0.70 ng day−1 m−2 in November 2009, 1.60 ± 0.04 ng day−1 m−2 in February 2010, and 3.00 ± 1.12 ng day−1 m−2 in May 2010. Mercury fluxes were highest in November 2009, corresponding to the high SGD rates.
4.1 Mercury Distribution in the Groundwater-Impacted Area of the Bay of Puck
4.2 The Mercury Fluxes to the Bay of Puck Via SGD
Benthic trace metal release by diffusion from permeable sediments is well documented (Beck et al. 2010). However, during the last decade, it was proven that pore water advection through permeable sediments also represents a significant source of trace metals to coastal waters (Bone et al. 2007; Beck et al. 2007a, 2010). In the Bay of Puck case, the groundwater is characterized by lower HgTD concentrations in comparison with seawater. As a result, the groundwater was a factor that dilutes the mercury concentrations in pore water originating from seawater intrusion. Consequently, after stormy events resulting in seawater intrusion into sediments and enriching pore water with mercury, SGD might be recognized as a significant source of mercury to coastal waters. During this study, we calculated mercury loads via SGD using the characteristic HgTD concentrations in the groundwater (Table 1; 0.63 ± 0.21 ng l−1) and the known groundwater flux to the Bay of Puck. We believe that only the groundwater is a source of mercury to the ecosystem. A groundwater flux (0.03 km3 year−1) was adopted from Korzeniewski (2003). As a result, HgTD fluxes via SGD to the Bay of Puck were equal to 18.9 ± 6.3 g year−1.
According to Boszke (2005), the Bay of Puck is fairly contaminated with mercury. Mercury mass balance (in- and outflows of mercury to the Bay of Puck ecosystem) calculated by Boszke (2005) showed that between 1.1 and 3.8 kg year−1 of Hg enters annually from the atmosphere, whereas the load of Hg carried to the bay with the river water was about seven times lower (0.13–0.44 kg year−1). The budget has proven that the main source of mercury in the system is the atmosphere. The mercury fluxes via groundwater discharge appear to be insignificant compared to both the abovementioned sources. The mercury flux obtained in this study is much lower in comparison with other groundwater-impacted areas, e.g., Etretat and Yport along the Pays de Caux Estuary, France (Laurier et al. 2007); Waquoit Bay in Massachusetts, USA (Bone et al. 2007); Elkhorn Slough in Central California, USA (Black et al. 2009); Stinson Beach in Northern California, USA (Black et al. 2009); Hwasun and Bangdu Bays on Jeju Islands, Korea (Lee et al. 2011); and Malibu Lagoon, CA, USA (Ganguli et al. 2012). In Malibu Lagoon, the average mercury flux via SGD was equal to 82.24 ng day−1 m−2, which was higher than that in Hwasun Bay (74.22 ng day−1 m−2) and lower than in Bangdu Bay (158.47 ng day−1 m−2), Stinson Beach (501.48 ± 320.95 ng day−1 m−2), Elkhorn Slough (601.77 ± 401.18 ng day−1 m−2), Waquit Beach (94.28 ± 381.12 ng day−1 m−2), Etretat (461.36 ± 762.24 ng day−1 m−2), and Yport (126.37 ± 421.24 ng day−1 m−2). This may be caused by both a lower concentration of mercury and lower SGD correlated with lower discharges of suspended rates in the study area in comparison with the other studies or a combination of these factors.
Thus, there is a need to note and discuss reasons for low concentrations of dissolved mercury in groundwater, much lower than concentrations of dissolved mercury in the wet atmospheric deposition. The water that is recharging the aquifer should have been delivering large mercury loads to the groundwater that are not manifested as increased concentrations of dissolved mercury in groundwater. Most of the load must be retained in the unsaturated and saturated zone of the surface soil since the groundwater mercury concentrations are low.
Assuming SGD rate to the Bay of Puck at 0.03 km3 (Korzeniewski 2003), the dissolved mercury concentration in rainwater at 8.6 ng dm−3 (Boszke 2005), and the average dissolved mercury concentration in the seeping groundwater at 0.6 ng dm−3 (this study, Table 1), the load of mercury that is retained in the surface soil in rainwater percolating to the saturated zone is equal to 0.03 km3 × 109 m3/km3 × (8.6−0.6)ng dm−3 × 103 dm3/m3 × 10−6 mg/ng = 0.24 × 103 g. The catchment area of the Bay of Puck is close to 100 km2 (Korzeniewski 2003; Kozerski 2007); surface soil is composed of peat and other organic-rich deposits and clays (Lidzbarski 2000; Kozerski 2007) that are characterized by high affinity and complexing capacity towards mercury (Boszke et al. 2008). Thus, it can be safely assumed that yearly retention of mercury, equal to 2.4 g mercury km−2, is well within a reasonably assumed retention capacity of the catchment. This conclusion agrees well with a very low concentration of mercury in the land-based wells used for extraction of drinking water (this study, Table 1) (Kozerski 2007).
5 Summary and Conclusions
Total dissolved mercury concentrations were measured in groundwater, pore water, and seawater in the Bay of Puck. Moreover, concentrations of mercury were measured in groundwater collected from wells located along the coast of the bay and in river water discharged to the bay.
The most important characteristic of the mercury concentrations profile in the groundwater seepage area is high concentrations of dissolved mercury in pore water of the uppermost sediment layer and the decrease of the concentration with depth below the sediment–water interface. This is attributed to seawater intrusion into the sediment.
Mercury behaves non-conservatively in the mixing of low-mercury groundwater and high-mercury seawater. This is caused by mercury removal in mixing seawater and groundwater due to redox reactions. Scaling up the results obtained for the study area indicated that the groundwater seeping to the bay seems not to be a significant source of mercury to the bay ecosystem.
The study reports the results obtained within the framework of the following projects: statutory activities of the Institute of Oceanology Polish Academy of Sciences theme 2.2 and research projects 3837/B/P01/201039 and 2012/05/N/ST10/02761 sponsored by the Polish Ministry of Science and Higher Education.
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