Trace metal behavior in sediments of Jiulong River Estuary and implication for benthic exchange fluxes

Environmental Pollution xxx (2017) 1e12

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Trace metal behavior in sediments of Jiulong River Estuary and implication for benthic exchange fluxes* Wenhao Wang a, Wen-Xiong Wang a, b, * a b

Environmental Science Program, The Hong Kong University of Science and Technology (HKUST), Clear Water Bay, Kowloon, Hong Kong Marine Environmental Laboratory, HKUST Shenzhen Research Institute, Shenzhen 518057, China

a r t i c l e i n f o

a b s t r a c t

Article history: Received 15 September 2016 Received in revised form 1 February 2017 Accepted 13 March 2017 Available online xxx

Severe metal pollution due to industrial effluents releases has been documented in Jiulong River estuary, Southern China. However, integrated understanding of trace metal behavior in the sediment is lacking. In the present study, DGT (diffusive gradients in thin films) technique was employed together with sediment cores to study the porewater dynamics of trace metals as well as the benthic exchange fluxes from four sampling sites over three different months. The sedimentary environment showed distinct spatial and temporal variations due to effluent discharge and biological activities. Metal behavior was controlled by early diagenetic reactions below the interface, in suboxic layer and in deeper sediment. Precipitation as sulfides and adsorption onto Mn/Fe (hydr)oxides were important in scavenging trace metals. Estimated exchange fluxes at sediment-water interface in this estuary indicated that the overlying water was a major source for trace metals, whereas sediments could also be the source if surface remobilization (Mn/Fe reduction) dominated. Our results highlighted the impacts of both natural and anthropogenic processes on the source, fate and transformation of trace metals in this dynamic system. © 2017 Elsevier Ltd. All rights reserved.

Keywords: Trace metal Porewater DGT profiles Estuarine sediments Industrial effluent

1. Introduction With progressive human activities and rapid industrial developments, estuaries in China are now facing increasing metal pollution pressures (Pan and Wang, 2012; Wang et al., 2014). Trace metals are of great concerns due to their toxic effects and potentials for substantial and long-term accumulation in sediments and organisms (Oursel et al., 2013). The sediments of the estuarine area have been considered as a sink for contaminants discharged to the environment (Kalnejais et al., 2015), however, trace metals may be remobilized allowing for diffusion back to the water column. Various processes as well as redox, pH, and sulfide influence the distribution of trace metals between solid to liquid phase (Gao et al., 2009), and determine whether a metal is sequestered in the sediment or remobilized to overlying water. The early diagenetic processes result in the first biogeochemical transformations of particles once they were deposited on the

* This paper has been recommended for acceptance by Dr. Harmon Sarah Michele. * Corresponding author. Environmental Science Program, The Hong Kong University of Science and Technology (HKUST), Clear Water Bay, Kowloon, Hong Kong. E-mail address: [email protected] (W.-X. Wang).

sediment (Berner, 1980). The mineralization of organic material is accompanied by the reduction of electron acceptors such as O2,
NO
3 , Mn(IV), Fe(III), and SO4 (Zhang et al., 1995). Oxidation of biodegradable organic matter in surface layer tends to supply metals to porewater, and dissolved organic matter could also play a role (e.g. stabilization of metals) even in deep layers (Charriau et al., 2011; Rigaud et al., 2013; Dang et al., 2015). Trace metals are scavenged from porewater by manganese and iron (hydr)oxides (Sundby, 2006), whereas reductive dissolution of these carrier phases causes the release of trace metals (Canavan et al., 2007). The production of sulfide by sulfate reduction bacteria can precipitate metals like Fe, Pb, Zn, Cu, or Cd as insoluble phases (Morse and Luther, 1999), and conversely the oxidation of sulfide releases dissolved metals. The relatively importance of these opposing reactions in surface sediment is crucial to understanding the source and fate of trace metals. Jiulong River Estuary is a shallow estuarine system on the southeast coast of Fujian Province, characterized by high fluctuation of hydrological and geochemical conditions due to natural mixing process and anthropogenic activities. Both intermittent effluent discharges and continuous drainage water releases are common (Wang et al., 2014), with input of industrial contaminants in an unprecedented manner. Previous studies in this estuary have

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revealed multi-metal pollution in surface sediments (Liu et al., 2006; Wang et al., 2011a) and in some bivalves (Wang et al., 2011a; Weng and Wang, 2014), but data on the vertical distribution and benthic flux of trace metals are limited. In coastal and estuarine area, it is challenging to predict metal cycling because of the strong variations in organic matter flux, temperature or biological activity, and porewater dynamics of trace metals have been studied in only a few locations worldwide (Gao et al., 2009; Lourino-Cabana et al., 2014; Kalnejais et al., 2015; Dang et al., 2015). The porewater profiles can be investigated by squeezing of sediment cores, however, it may be difficult to achieve satisfactory resolution and artificial reactions may occur (Zhang et al., 2002). DGT (diffusive gradients in thin films) is an in situ technique measuring directly the flux of metal from the sediment during the deployment time, which reflects the concentration in the porewater, its diffusional transport and the supply from the solid phase to solution (Davison and Zhang, 1994; Stockdale et al., 2010). The DGT probe consists of a filter membrane, a diffusive gel and a binding gel containing Chelex cation-exchange resin. Labile metal fraction small enough to diffuse through the gel and capable of binding to the resin layer is assessed. Furthermore, the exchange fluxes at the sediment-water interface (SWI) can be estimated based on the high resolution profiles of trace metals derived by DGT technique (Gao et al., 2009). In practice, the rate of release from the sediment can limit resupply, resulting in underestimation of porewater concentration. A steady-state condition will not be truly reached, and time-dependent models have been applied to quantify the contribution of diffusional supply and release from the solid phase to the accumulated mass of solute (Davison et al., 2007). The main purpose of the study was therefore to combine both conventional and novel techniques to reveal trace metals behaviors in porewater, the most sensitive fraction of sediment. Several questions were addressed in this study, including 1) how environment conditions differed in space and time, 2) how these changing conditions controlled metal remobilization or scavenging? 3) what influenced benthic exchange fluxes of these metals? 4) what influenced the partitioning of metals among labile, dissolved and particulate phases in sediment. DGT-Chelex probes were employed together with physico-chemical parameters measurement and sediment cores collected in three different months from four sites of Jiulong River Estuary. Vertical profiles of labile, dissolved and total metal concentrations were obtained, and metals in the colloidal fraction as well as partitioning for Cu, Zn Ni, Co, Pb, Cd, Mn were compared. We discussed the two main pathways for removing metals from porewaters in this estuarine area and highlighted both natural processes and anthropogenic activities impacting trace metal fluxes at sediment-water interface. Results from this study will provide information on long-term environmental risk assessment and pollution mitigation. 2. Materials and methods 2.1. Study sites The Jiulong River, with an annual net flow rate of 14 billion m3, discharges to the coastal sea of Xiamen City through the Jiulong River estuary. The estuary is a typical subtropical macro-tide system (Yan et al., 2012) with a catchment area of 14,741 km2. This study focused on four sites of north branch of Jiulong River Estuary (Fig. 1), close to newly established industrial zones that host factories (Wang et al., 2011a). Industrial effluents were often stored in holding ponds and frequently discharged into the estuary during low tide (Wang et al., 2014). Site 1 (S1) and Site 2 (S2) were directly receiving the visible intermittent effluent, whereas Site 3 (S3) and Site 4 (S4) were located at about 1 km downstream.

The mixing of effluent and estuarine water was characterized by high suspended particulate matter along with low dissolved oxygen. Our earlier study indicated different sources of trace metals, e.g., particulate Cu, Zn and dissolved Ni were from the intermittent effluent discharge, but dissolved Cu and Zn were from the potential effluent release downstream (Wang and Wang, 2016). Sediments in S1 were black and muddy, on the contrary, saltmarsh vegetation and macrobenthos were abundant around S4. The inter-site differences were complex, and we hypothesized that the contaminated S1 and S2 would show high total concentration of trace metals, while the diagenetic reactions might limit their mobility. However, the downstream sites would show contrastive situations. Given the fact that bacteria activity varied according to temperature, seasonal migration of porewater profiles was also expected, which implied benthic exchange and cycling of trace metals. 2.2. Sampling and processing of sediment cores During 2015, three sediment sampling campaigns in Jiulong River Estuary were carried out on June 1, July 18 and October 20, respectively. For these three periods the atmosphere temperature were 24e28, 25e34, 22e29  C. The four study sites were in shallow systems that allowed by hand collection or deployment. At each site, one sediment core was collected with 25 cm long PVC tube. The cores were placed on ice and brought back to the laboratory immediately after sampling. Measurements of pH, redox potential (Eh, referred to the Ag/AgCl electrode) and temperature in the porewater were directly conducted in the sediment core by introducing electrodes (PE-06HD and ORP-14, Lutron) and thermometer through the holes of an additional PVC tube. As logistically soon as possible, sediment cores were sectioned every centimeter under inert atmosphere (N2) in a glove bag and put partly in plastic bags and partly in centrifuge tubes (Superville et al., 2014). Sectioned sediments were centrifuged at 5000 rpm for 15 min. Porewater was recovered also under nitrogen by filtration (0.2 mm syringe filters, Millipore) and then preserved in 0.2% HNO3 (trace metal grade). The sediment samples were freeze-dried, followed by homogenization, passing through 0.18 mm sieve and digestion in nitric acid with microwave heating based on the modified method of USEPA 3051. The method here provided quasi-total fraction to consider. After centrifugation, the supernatant from the digested solutions was pipetted and diluted for further metal analysis. 2.3. Deployment and retrieval of DGT probes The DGT probes (DGT Research Ltd.) were 180 mm  40 mm in size, with a window of 150 mm  18 mm open to the aquatic system. Before deployment, DGT probes were de-oxygenated by immersing them in N2 saturated NaNO3 solution (trace metal cleaned, 0.01 M). Sulfides were measured by DGT-AgI probes, similar to DGT-Chelex probes, but AgI was used here as binding agent (Lourino-Cabana et al., 2014). In June, only DGT-Chelex probe was used. In July and October, we employed both DGT-Chelex and DGT-AgI probes. At each site, DGT probes of 0.78 mm diffusive gel thickness were vertically inserted in situ in the estuarine sediment during an exposition time of 25 h. The sediment-water interface was marked upon retrieval and the probes were rinsed quickly with Milli-Q water. In the laboratory, the Chelex-resin gel removed from the DGT assembly was cut into 5 mm intervals. Each gel slice was eluted in 1 M HNO3 solution in 2 ml centrifuge vials and then diluted for analysis by ICP-MS. DGT measured all species in labile equilibrium along with the species that can bind to binding agent, which are considered to be bioavailable fractions. Previous studies have demonstrated that the

Please cite this article in press as: Wang, W., Wang, W.-X., Trace metal behavior in sediments of Jiulong River Estuary and implication for benthic exchange fluxes, Environmental Pollution (2017), http://dx.doi.org/10.1016/j.envpol.2017.03.028

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Fig. 1. Locations of the four study sites in Jiulong River Estuary.

concentration gradient in the diffusive gel stayed approximately constant for a 24 h deployment (Gao et al., 2006; Leermakers et al., 2005), thus DGT derived results can be interpreted as the timeaveraged concentrations of solutes in bulk solution. Assuming that the capacity of the solid phase to resupply the porewater is large (Zhang et al., 2001), labile metal concentrations in porewaters measured by DGT (CDGT) are calculated from the mass accumulated in the resin gel (M) using Fick's Law:

CDGT ¼

M Dg ðDtAÞ

were quantified. The recoveries of IAEA-158 were 90%e105% for Cu, Zn, Ni, Co, Pb, Cd, Mn and Fe. The recoveries of SLEW-3 also showed good agreement with certified data. Besides, dry sediment was weighed into pre-combusted silver capsules and acidified with 10% high purity HCl to remove inorganic carbon, according to Hedges and Stern (1984). TOC content was then measured by high temperature combustion using CHN elemental analyzer (PE 2400 Series), with cysteine used as standard material.

(1)

where Dg is the thickness of diffusive gel, D is the diffusive coefficient of a metal in the gel, t is deployment time and A is the exposure area. Dissolved sulfide, defined as S(-II) (H2S þ HS
þS2
), diffused from porewater through the filter membrane as well as diffusive gel, and reacted with pale yellow AgI in the binding gel to form black Ag2S (Wu et al., 2014). The AgI-resin was dried and scanned by a flatbed scanner where color intensity was digitized and calibrated to obtain the concentrations initially present in porewater (Teasdale et al., 1999; Naylor et al., 2004).

2.4. Analytical procedures Trace metal (Cu, Zn, Ni, Co, Pb, Cd) concentrations in porewater, digested sediment and eluted DGT samples were determined by inductively coupled plasma mass spectrometry (ICP-MS, PE NexION 300X), and the metals Mn and Fe were determined using inductively coupled plasma optical emission spectrometry (ICP-OES, PE Optima 7000 DV). The porewater samples were diluted (10e15 times) with acidified Milli-Q and standard addition methods were selected to use for trace metals analysis to avoid the impact of high ‘salt’ concentrations (Simpson et al., 2014). Procedural blanks for sediment and DGT samples were also processed and examined. Daily performance was checked and optimization was performed before running the instruments, and quality control samples were repeatedly measured after every 10e20 samples. To evaluate the accuracy of analytical procedures, certified reference materials (CRMs) IAEA-158 (marine sediment) and SLEW-3 (estuarine water)

3. Results and discussion 3.1. Physico-chemical parameters The pH values ranged between 6.7 and 8.1 for all the in situ measurements (Fig. 2). In S1, S2 and S3, the highest pH was in June, and the lowest pH was found in July for the four sites, probably due to enhanced organic matter mineralization in the summer when bacterial activity was promoted. Generally, the pH values did not show significant variation with depth, unless a decreasing trend of pH in June in S3 that dropped below 7 at
10 cm. The redox potential (Eh) represents the oxidation or reduction state in porewater. Negative Eh values indicated depletion of oxygen and reducing environment in sediments of the study sites (Song et al., 1990), which was responsible for Mn/Fe reduction and even sulfate reduction. Within 8 cm depth of S2 and S3, the Eh was lower in July than in October, suggesting a strong anaerobic condition in summer. As a result of anthropogenic hypoxia and input of organic matter, S1 was the most reducing site with the most negative Eh value (
180 mV on average). Even lower Eh values were observed in October in S1, subject to effluent discharge as well as high demand for oxidants. Contrastively, S4 was the least reducing among these sites. Several high values of Eh were found in July in S4, likely due to enhanced bioturbation in the summer time that introduced oxygen to the sediment. Fig. 2 also shows the total organic carbon contents as well as estimated dissolved sulfide concentrations for the four sites from three different months. The average TOC content in S1 (1.22%) and S2 (1.20%) were higher than those in S3 (1.00%) and S4 (1.07%). These values were lower than the literature reported ones in

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Fig. 2. Vertical profiles of PH, redox potential, TOC, and dissolved sulfide for the four sites from three different months. Redox potential and dissolved sulfide were not measured in June.

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Jiulong River Estuary (Yin et al., 2015) and Massachusetts Bay (Kalnejais et al., 2015). The spatial variations of sedimentary TOC could be found, which was potentially related to terrestrial input at upstream sites. This corresponded to the observed differences in redox potential, and high organic matter flux usually accounted for more reducing environment (Zhu et al., 2011). There was no obvious variation for the three months sampling, except for sporadic fluctuations at S1 where effluent-induced deposition along with tide-induced erosion was often and at S4 where bioturbation was important. The average concentration of dissolved sulfide derived by DGT for the study sites suggested the following sequence: S1 (1.79 mM) > S2 and S3 (1.22 mM and 1.30 mM respectively) > S4 (0.62 mM). The highest sulfide production in S1 was related to effective sulfate reduction in anoxic sediment, and profile peaks located at around 2 cm of depth. The sulfide profile in S2 had typical peaks below the interface (
2 to
3 cm) and in deeper layer (
13 cm). S3 was characterized by a single peak in sub-surface sediment, and dissolved sulfide showed higher values in July. In July at S4 (the least reducing case), sulfide was hardly detected, corresponding to the >
150 mV Eh values. In fact, these sulfide levels were very low when compared to Fe, similar to situations (Naylor et al., 2004) where simultaneous remobilization of metals and sulfide could occur. Sulfide was easily exhausted for its relatively low inventory and variations of salinity in the water column might limit sulfide production in this estuarine area (Yin et al., 2015). These physico-chemical parameters implied spatial and temporal changes of the sedimentary environment, i.e., more reducing conditions in the upstream sites and in the summer (except S4), which would have impacts on the trace metal behaviors.

5

1064 mg/L in October), coincidental with the less contaminated sediments in Massachusetts Bay (Kalnejais et al., 2015). 3.2.2. Copper, zinc and nickel The DGT profiles of Cu were characterized by elevated concentrations below the interface, reflecting diffusive process from the overlying water. In June, the aerobic mineralization of organic material contributed to the peak values of Cu at S3 and S4 within 2 cm. Cu was also considered to be associated with Mn oxides (Shaw et al., 1990), but in this study similar patterns for Cu and Mn were only observed in October at S1 showing a rapid decrease followed by a stable trend with depth. Compared to the Cu profiles, higher labile concentrations were observed for Zn (ranging between 0.91 and 19.4 mg/L). Zn generally displayed surface maxima due to the releases from organic matter and Mn/Fe (hydr)oxides. Literature documented great supply of Zn from the ferruginous zone (Naylor et al., 2004), and our results demonstrated the potential correspondence of Zn concentration with Mn and Fe dissolution, such as the peaks locating at 1e2 cm of depth for the July and Oct. measurements. The peak locations of Zn were slightly shallower in July at S2 and S3, implying upward benthic exchanges. The vertical distribution of Ni resembled that of Cu and Zn, suggesting their similar releasing or scavenging pathways in porewaters, which was also documented elsewhere (Fones et al., 2004). However, there were two major differences. First, Ni supply from decomposing organic matter might be insufficient (Naylor et al., 2004), thus the elevated concentrations below the interface at S3 and S4 were not as obvious as Cu, Zn. Second, previous studies suggested undersaturation of NiS (Naylor et al., 2004), which led to a more stable pattern with gradual decline for Ni in deeper layer. Ni profiles did not show pronounced monthly trend.

3.2. Vertical profiles of labile metals derived by DGT 3.2.1. Manganese and iron Vertical profiles of labile metals (Cu, Zn, Ni, Co, Pb, Cd, Mn, and Fe) derived by DGT for the four study sites in Jiulong River Estuary from three different months are illustrated in Fig. 3. The metal concentrations above sediment-water interface (marked with dash line) actually corresponded to the DGT measurements in bottom estuarine water. First, we presented the major redox species to establish the variable sedimentary conditions. Generally, manganese and iron profiles displayed an increase below the SWI, attributing to the reductive dissolution of Mn and Fe (hydr)oxides mediated by bacterial activity (Lourino-Cabana et al., 2014). Mn was released to the porewater shallower than Fe, consistent with the ideal redox sequence in marine sediments (Jørgensen and Kasten, 2006). For all the four sites, Mn showed a typical peak at 1e2 cm of depth in the July measurement. Mn was not completely removed from porewater, given the control rather by carbonates (Gao et al., 2009). Very low concentrations of Fe were observed in overlying water and in the first 1 cm of depth, except for S1 in October where the most reducing condition contributed to sharp Fe gradient below the interface followed by rapid decrease in deeper layer. For S2 and S3, upward profiles as well as elevated release of dissolved Fe (reaching 6000 mg/L) in July was due to the enhanced reduction of iron (hydr)oxides. Similar migration of ferruginous zone in Hingham Bay was documented by Kalnejais et al. (2015), but in their case sulfide played a more important role in the long-term transition. There appeared to be simultaneous release of sulfide with Fe, e.g., in suboxic layer of S2 and S3, probably due to the coexistence of sulfate and iron reducing bacteria as described by Naylor et al. (2004). For Site 4 where oxygen might contribute, Fe had relatively low concentrations (depth average of 1349 mg/L in July and

3.2.3. Cobalt, lead and cadmium The Co profiles presented seasonal variations as well as vertical fluctuations, with labile concentrations ranging from 0.11 to 10.3 mg/L. Significant correlation (regression coefficient 0.567) between Co and Mn was found. Co was known to be coincidental with Mn and Fe mobilization (Gao et al., 2006; Tankere-Muller et al., 2007), and either Mn or Fe, or both, determined the peaks of Co below the interface (e.g. in October at S1) and in suboxic layer (e.g. in all 3 months at S3, S4). The Pb profiles displayed sporadic high concentrations at various depths. Pb did not show significant relationship to Mn or Fe, and previous thermodynamic calculations illustrated that the association of Pb to other fractions such as organic ligands and sulfide was strong (Lourino-Cabana et al., 2014; Dang et al., 2015). Similarly, no clear seasonal trend was reported from the high frequency monitoring by Lourino-Cabana et al. (2014). The maxima of Cd concentration below the interface were due to releases from organic matter when it was oxidized (TankereMuller et al., 2007). In suboxic sediments (even above the sulfidic zone), Cd precipitated as CdS and was removed from porewaters even if only trace levels of sulfide were present (Rosenthal et al., 1995). There were sharp decreases of Cd profile from depths of 0.5 cm at S1, 1 cm at S2 and S3, and 1e2 cm at S4. The average labile Cd concentrations were 0.006, 0.010, 0.014, 0.037 mg/L for the four sites, respectively, indicating the lowest Cd bioavailability in the most reducing site. 3.3. Seasonal variations and benthic exchange fluxes Previous studies explored the temporal variability and porewater dynamics of trace metals, and demonstrated the importance of temperature-dependent bacterial activity as well as organic

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Fig. 3. Vertical concentration profiles of Cu, Zn, Ni, Co, Pb, Cd, Mn, and Fe (in mg/L) derived by DGT for the four sites from three different months. Fe was not assessed in June.

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Fig. 3. (continued).

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matter flux (Gao et al., 2009; Lourino-Cabana et al., 2014; Kalnejais et al., 2015). In this study, the highest temperature might account for the enhanced Mn/Fe and sulfate reduction in the summer time. Such trend was pronounced from the July measurement at S2 and S3, as reflected by lower pH and Eh values, higher dissolved sulfide concentrations, along with advancing redox boundary. There could be other reasons for the observed variations, including intermittent flooding due to tidal hydrology and intermittent hypoxia caused by effluent discharge, both of which more or less had impacts on the reducing conditions in sediment, and ultimately influenced trace metal behaviors. Overall, with shallower release of Mn and Fe in July, the Zn, Ni and Co profiles moved upward, so did Cu, Pb and Cd but less obviously. Although Lourino-Cabana et al. (2014) documented low bioavailability of several metals in summer under high sulfide levels, we did not find strong sulfidation processes (unless sometimes at S1 and S2), which was related to the limited sulfide production in these sites. S1 and S4 displayed dramatic opposite trends, with even more negative Eh values in October, indicating factors other than those discussed above would alter the oxidation or reduction state. Frequent effluent discharge was recorded in October, generating low oxygen in overlying water that directly facilitated diagenetic reactions (reduction) in the S1 porewater. Another concern was that the demand for oxidants in autumn was still high, and similarly Kalnejais et al. (2015) documented advancing profile as well as sharp gradient of Fe in a sulfidic site. However for S4 with less anthropogenic activities, localized bioturbation/irrigation should be considered, as the saltmarsh macrobenthos were abundant. The highest secondary production and biomass of macrobenthos (e.g. Notomastus latericens and Potamocorbula laevis) have been found in summer in this region (Zhou et al., 2008). This caused the introduction of oxygen, formation of Mn/Fe (hydr)oxides, and thus sequestration of trace metals in the solid phase. Theoretical diffusive fluxes across sediment-water interface can be calculated from high resolution DGT profiles using Fick's Law (Berner, 1980):

F ¼ ∅Ds

  vC vz

(2)

where ø is porosity; Ds is the diffusion coefficient in porewater,

which is estimated by combining ø and D0 (diffusion coefficient in water), according to Ullman and Aller (1982). Concentration gradient was obtained by the difference between metal concentrations in the upper 1 cm porewater and in the overlying water. Fig. 4 illustrates the benthic exchange fluxes of these metals for the four sites from three different months. The fluxes of Mn and Fe at SWI were much higher than the other trace metals and were generally upward (ranging from
6  10
4 to
1
2
1 4.3  10 mmol m $d ), consistent with literature reported magnitudes and directions (Pakhomova et al., 2007; Kalnejais et al., 2015). In fact, their dissolved concentrations in the water column were low due to oxidation and precipitation (Gao et al., 2009). This demonstrated the effective remobilization mediated by Mn/Fe reduction in suboxic condition, and thus the benthic sources of Mn and Fe were important when budgeting their fluxes in the estuarine system. The negative values indicated downward fluxes and sediment served as the sink for Cu, Zn, Ni, Co, Pb, Cd at most times. In July, there were upward benthic fluxes especially for Cu (at S3, S4) as well as Zn and Ni (at S2 and S4), similar to the results from Gao et al. (2009) where typical fluxes of 5.4  10
5 to
4
2
1 1.6  10 mmol m $d for Cu and Ni were recorded. The seasonal advancing of redox boundary and metal profile peaks resulted in the diffusion from surface sediment back to the water column. The magnitude of Pb and Cd fluxes was even smaller (<1  10
5 mmol m
2$d
1), indicating limited diffusion across the interface. Because of high affinity for fractions other than Mn/Fe, no significant monthly trend for Pb and Cd was observed. The sediment would become the source of trace metals to the estuary if these metals were solubilized and remobilized (mainly due to corelease with Mn/Fe reduction) near the interface. Moreover, not only natural processes but also human activities had impact on metal fluxes at SWI. The Jiulong River Estuary suffered from intermittent and continuous effluent releases. The DGT derived concentration above the interface was coincidental with the previous study focusing on long-term fluctuations of trace metals (Cu: 2.84e12.54, Zn: 4.81e19.92, Ni: 4.62e26.10, Co: 0.15e0.95, Pb: 0.05e0.39, Cd: 0.04e0.12 mg/L, Weng and Wang, 2014) and suggested the contamination in the estuarine water. The overlying water became the dominated source, determining the downward fluxes, and our results in October were the example.

Fig. 4. Exchange fluxes of Cu, Zn, Ni, Co, Pb, Cd, Mn, and Fe (in mmol$m
2$d
1) at sediment-water interface.

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Some elevated labile metal concentrations such as for Cu, Ni and Cd within 1 cm of depth were suspected to be affected by such diffusion. These additional metals in surface sediment might further undergo adsorption, precipitation or remobilization with changing physico-chemical conditions and participate in the cycling over the year. Sediments were not the closed environments, and other processes were involved in the interaction across SWI. The role of bioirrigation in benthic exchange remained unclear since either enhanced flushing or scavenging could occur (Kalnejais et al., 2015). The recurrent resuspension led to mobilization of sediment along with entrainment of porewater (Prygiel et al., 2015; Superville et al., 2015), altering the distribution of trace metals. The long-term monitoring would be necessary for better quantifying the relative importance of the benthic source in this dynamic system, and profile modelling (e.g. Berg et al., 1998; Dang et al., 2014) would also be promising in interpreting these processes. 3.4. Dissolved metal concentrations in porewater and R value Vertical profiles of dissolved metals in porewater for the four study sites in Jiulong River Estuary from three different months are shown in Fig. S1. The concentrations of Cu, Zn, Ni, Co, Pb, Cd, Mn derived by DGT were generally 1e10 times lower than those obtained by centrifugation. Comparison between the conventional method and the DGT technique should be carried out with caution (Gao et al., 2006) as DGT actually measures the localized interfacial concentration in a small volume of sediment while micro-features may be missed with centrifugation method. Nevertheless, the coupling between metal distributions derived by DGT and by centrifugation demonstrated profile shape fidelity. The Mn peaks from the dissolved profiles resembled those from the DGT results. Mn showed relatively stable trend at S1, with the lowest value in June but highest in July, corresponding to the redox variation. For S4, the decrease in dissolved Mn concentration in July again suggested the removal from porewater. The dissolved profiles of Cu, Zn displayed some similarities, and the peaks below the sediment-water interface were documented. Zn appeared to be more mobilized (on average 10.2 mg/L for all the measurements), with elevated concentrations in surface layer (e.g. in October, resulted from overlying water diffusion and Mn/Fe reduction). In general, dissolved concentrations for Cu and Zn decreased in the order S3, S4 > S2 > S1. Otherwise, the dissolved concentration for Ni indicated the following order: S1, S2 > S3 > S4. This was consistent with previous study showing the intermittent effluent next to S1 was the source for dissolved Ni in Jiulong River Estuary (Wang and Wang, 2016). Interestingly at S1, high values of dissolved Ni was found from
2 to
3 cm, which was simultaneously recorded by DGT and was attributed to local niches in the disturbed sediment. The controlling of Mn on Co behaviors was also demonstrated from the dissolved profile, and surface or sub-surface maxima of Co were clear. In July, the peaks located within 1 cm for S1 and S2, and at
3 cm for S3 and S4. In October, the peak locations were around 2e3 cm of depth, with highest concentration reaching 17.5 mg/L at S4. There appeared to be seasonal advancing (upward profiles in summer) as well as spatial variation (S1 and S2 were more reducing than S3 and S4). For Pb, such variation was not pronounced. The factors determining solid-solution interaction of Pb were rather complex, although Dang et al. (2015) documented mobilization peaks in the surface layer. Besides, the Cd maxima below the interface due to organic matter decomposition was also observable, corresponding to the DGT measurements. Similarly, the Site 4 showed the highest concentration of Cd. Concentration fidelity can be estimated in terms of a ratio R,

9

calculated as DGT/centrifugation (Harper et al., 1998). The R value usually varies between ~0.1 and 1, interpreted as the capacity of the solid phase to resupply solutes to the porewater. The underestimation of DGT derived concentration was also due to existence of non-labile fraction (Lesven et al., 2008), such as complexing ligand, that competed with the Chelex resin. The depth average R values along with maximum and minimum values from one profile for the four sites from three different months are illustrated in Table 1. R values for these metals decreased in the sequence: Zn (0.64) ¼ Pb (0.64) > Co (0.42) > Mn (0.31) > Cd (0.29) > Ni (0.19) > Cu (0.16). The sequence and values were coincidental with the results of Helkijn sediments (Gao et al., 2006). Cu has been considered to have low lability (Huo et al., 2015), while Zn and Pb might depend on in situ conditions with various labile fractions documented (Lesven et al., 2008; Superville et al., 2014). Co, Cd, Ni and Mn were moderately labile. It would be important to investigate the trace metal speciation in porewaters. How sediment property influences concentration fidelity was further investigated, and there was reverse trends between R values and pH for trace metals except Ni and Cd. Low pH values facilitated resupply kinetics and lability of these metals. Certain spatial and temporal differences for the DGT/centrifugation ratio were also found. Cu and Zn displayed lower R values at the two downstream sites than upstream, and the formation of organic complexes influenced by newly supplied DOC/POC at S3 and S4 was possible. Gao et al. (2009) investigated the effect of phytoplankton blooms on low labile Fe concentration in surface sediment. In Jiulong River Estuary, the areas with high phytoplankton abundance were discriminated at seawaters and brackish waters (Wang et al., 2011b). Additionally, in July at S4 where there was increasing redox potential, elevated R value for Mn was observed, similar to what Lesven et al. (2008) described. This was because equilibrium of labile Mn between gel and porewater was shifted towards the former, when Mn presented in the gel was oxidized. 3.5. Total metal concentrations in sediment and partition coefficient Vertical profiles of total metal contents in sediment for the four study sites from three different months are shown in Fig. S2. The concentrations were generally within the ranges from the literature reported ones in north branch of Jiulong River Estuary and its vicinity (Wang et al., 2011a; Lin et al., 2009; Yu et al., 2013). The high concentration of Mn (on average 1058 mg/kg) was related to the native rock type, either basalt or granite (Liu et al., 2006). Mn remained stable with increasing depth, except for the variations in the first 1 cm or high sporadic levels in July at S4, both of which were contributed by potential mobilization or disturbance. The diagenesis-driven surface enrichment that could occur in suboxic sediment with redox boundary migration (Gobeil et al., 1997; Kalnejais et al., 2015) was not found in the studied system. For the other trace metals, the vertical distribution with less fluctuation at S3 and S4 indicated that the sedimentary environment was rather steady-state. At S2, the elevated concentrations in the first 3 cm were probably due to recent input of anthropogenic pollutants. Historical contamination may exist, however, it was difficult to conclude whether the total metal contents were increasing through the years since the trend was not obvious from the profiles obtained. S1 displayed differences among the sampling periods, with total metal concentrations ranging from 67 to 254, 155e467, and 30e126 mg/kg for Cu, Zn and Ni, respectively, which may be caused by the repeatedly effluent-induced deposition along with tide-induced erosion, and may reflect local heterogeneity. For S2, S3, S4 where natural processes dominated and hydrological conditions were moderate, the three months results were relatively constant.

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10

W. Wang, W.-X. Wang / Environmental Pollution xxx (2017) 1e12

Table 1 R values (DGT/centrifugation) for the four sites from three different months. Sampling dates

Cu

Zn

Ni

Co

Pb

Cd

Mn

06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20

S1

S2

S3

S4

Ave

Max

Min

Ave

Max

Min

Ave

Max

Min

Ave

Max

Min

0.09 0.31 0.15 0.75 0.90 0.45 0.15 0.12 0.20 0.13 0.12 0.47 1.25 0.80 0.54 0.08 0.31 0.07 0.36 0.23 0.26

0.19 0.90 0.23 1.47 1.99 0.70 0.30 0.19 0.37 0.22 0.18 1.27 3.72 1.73 0.71 0.18 0.52 0.13 0.57 0.29 0.46

0.06 0.02 0.09 0.43 0.32 0.23 0.09 0.08 0.10 0.06 0.03 0.14 0.47 0.18 0.31 0.04 0.12 0.02 0.15 0.13 0.17

0.20 0.26 0.12 0.55 1.21 0.52 0.36 0.14 0.17 0.35 0.18 0.28 0.15 0.72 0.44 0.45 0.32 0.18 0.39 0.20 0.20

0.42 0.53 0.18 1.10 1.84 0.97 0.64 0.24 0.23 0.61 0.34 0.52 0.27 1.16 1.19 1.54 0.77 0.30 0.46 0.25 0.56

0.08 0.05 0.09 0.22 0.80 0.27 0.20 0.06 0.11 0.24 0.05 0.12 0.05 0.24 0.21 0.10 0.11 0.12 0.32 0.14 0.07

0.07 0.13 0.13 0.42 0.52 0.18 0.23 0.13 0.10 0.33 0.65 0.25 0.24 0.42 0.35 0.34 0.23 0.15 0.18 0.31 0.22

0.23 0.21 0.22 0.86 0.71 0.30 0.35 0.16 0.16 0.67 1.40 0.38 0.70 0.85 0.82 1.14 0.33 0.55 0.30 0.81 0.48

0.02 0.07 0.08 0.22 0.30 0.10 0.11 0.05 0.06 0.10 0.27 0.20 0.11 0.18 0.09 0.04 0.06 0.04 0.08 0.15 0.11

0.14 0.17 0.13 0.69 1.22 0.27 0.27 0.22 0.26 0.33 1.43 0.55 0.65 1.91 0.18 0.53 0.32 0.56 0.11 0.95 0.24

0.29 0.60 0.16 1.00 5.47 0.34 0.39 0.48 0.36 0.47 3.65 1.14 1.51 5.31 0.25 0.98 1.33 2.11 0.15 1.50 0.37

0.07 0.03 0.11 0.41 0.39 0.19 0.19 0.08 0.18 0.19 0.47 0.21 0.17 0.22 0.12 0.23 0.05 0.10 0.07 0.56 0.17

concentration in particulate phase (mg/kg) and in dissolved phase (mg/L). The Kd value applies only to in situ conditions where it was assessed and depends on different factors. The depth average, maximum, and minimum Kd values for the four sites from three different months are illustrated in Table 2. The Log Kd values for trace metals showed the following sequence: Pb (5.6) > Zn (4.4) > Cu (4.1) > Cd (3.7) > Co (3.6) > Ni (3.3) > Mn (2.1), consistent with previous literature survey (Allison and Allison, 2005) or field investigation (Weng and Wang, 2014). Note that the contrasting particulate partition for Zn and Ni was attributed to anthropogenic discharge of particulate Zn and dissolved Ni in this area. Among the concerned metals, the highest Kd value for Pb but the lowest value for Mn were probably due to the large reducible or oxidizable fraction that trapped Pb and large exchangeable fraction that kept Mn mobilized, respectively (Charriau et al., 2011). The Kd values for Cu, Zn and Co from the four sites decreased in the sequence S1 > S2 > S3 > S4. For all these metals except Mn, highest partition coefficients were recorded at S1 or sometimes at

The total Mn, Co, and Pb contents did not exhibit significant spatial differences. Contrastively, Cu, Zn, Ni, Cd in sediments of the four sites decreased in the order S1 > S2 > S3 > S4, illustrating the accumulation of particulate metals resulted from the intermittent effluent discharge. The average concentrations of total Cu, Zn, Ni, Cd in S1 were 2.6, 1.6, 1.8, and 1.4 times higher than those in S4. At this stage, S1 was considered to undergo more severe pollution, and the downstream S4 was less contaminated which was quite similar to the adjacent Western Xiamen Bay where metal concentrations in sediments met the Chinese National Standard Criteria for Marine Sediment Quality (Zhang et al., 2007). However, the high total metal concentration did not guarantee high supply to porewaters, which were impacted on much shorter time scales (Jahnke et al., 1982). Various factors such as redox, oxygen and sulfide controlled the distribution of trace metals between solid to liquid phase. The solid-solution interactions can be described using the partition coefficient Kd (L/kg), calculated as the ratio between metal

Table 2 Log Kd values for the four sites from three different months. Sampling dates

Cu

Zn

Ni

Co

Pb

Cd

Mn

06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20 06e01 07e18 10e20

S1

S2

S3

S4

Ave

Max

Min

Ave

Max

Min

Ave

Max

Min

Ave

Max

Min

4.4 4.3 4.3 4.8 4.7 4.5 3.4 3.3 3.3 3.6 3.8 3.8 6.0 5.5 6.0 3.9 3.8 3.9 2.3 2.1 2.0

4.5 4.6 4.0 5.0 5.0 4.7 3.5 3.4 3.6 3.7 3.9 3.9 6.4 5.8 6.2 4.1 3.9 34.0 2.5 2.2 2.1

4.3 3.9 4.6 4.6 4.4 4.2 3.3 3.2 3.0 3.5 3.6 3.6 5.8 5.2 5.9 3.7 3.7 3.7 2.2 2.1 1.9

4.2 4.3 4.2 4.4 4.8 4.3 3.4 3.3 3.2 3.5 3.7 3.5 5.1 5.6 5.8 3.9 3.9 4.0 2.2 2.0 1.9

4.3 4.5 4.3 4.5 5.0 4.5 3.4 3.5 3.3 3.6 3.8 3.7 5.6 6.0 6.0 4.3 4.2 4.2 2.2 2.1 2.5

4.0 4.2 3.7 4.2 4.5 4.0 3.3 3.2 3.0 3.4 3.6 3.2 4.6 5.3 5.4 3.7 3.7 3.3 2.1 2.0 1.6

3.8 4.0 4.0 4.2 4.5 4.1 3.2 3.1 3.1 3.4 3.5 3.6 5.3 5.6 5.6 3.8 3.7 3.5 2.0 2.0 2.1

4.1 4.2 4.1 4.4 4.7 4.2 3.3 3.3 3.4 3.5 3.7 3.8 5.7 5.9 6.1 4.0 4.0 3.7 2.1 2.7 2.7

3.6 4.0 3.9 4.2 4.3 4.0 3.2 2.9 2.6 3.3 3.4 3.4 4.9 5.3 5.3 3.6 3.3 3.2 1.8 1.9 1.9

3.7 4.0 3.8 4.3 4.4 4.0 3.3 3.2 3.3 3.3 3.8 3.4 5.6 5.6 5.1 3.5 3.4 3.6 1.8 2.9 2.0

3.8 4.2 3.9 4.4 5.2 4.1 3.3 3.4 3.4 3.4 4.2 3.6 5.8 6.1 5.3 4.0 3.9 3.9 2.0 3.9 2.6

3.6 3.8 3.6 4.3 4.0 3.8 3.1 3.0 3.2 3.3 3.6 3.2 5.2 5.0 4.9 2.9 2.9 2.9 1.7 2.5 1.7

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W. Wang, W.-X. Wang / Environmental Pollution xxx (2017) 1e12

S2. The formation of Mn/Fe (hydr)oxides in July at S4 was the example for effectively removal in a more oxidizing state. Otherwise, early study documented enhanced adsorption onto particulates as well as limited role of colloids during mixing of effluent and estuarine water (Wang and Wang, 2016), and the particulates with high total metal contents were deposited near Site 1 and 2. The anthropogenic hypoxia and discharge of organic matter further generated anoxic environment, where sulfides would scavenge metals as pure minerals and/or by co-precipitation (Charriau et al., 2011). When all the data were combined together, interestingly R values showed positive relationship with Kd values, meaning that the high partition coefficients did not necessarily limit the resupply kinetics but might limit the non-labile fraction. These were the reasons for the significant spatial variation observed. 4. Conclusion Due to anthropogenic input of contaminants in Jiulong River Estuary, study sites near the intermittent effluent release showed higher total metal contents (except Co and Mn) than the downstream ones. However, dissolved Cu and Zn concentrations in porewaters with reversed sequence S1 < S2 < S3 and S4 suggested the enhanced particulate/dissolved phase partitioning in reducing environment. S1 was the most reducing site while S4 was the least, as reflected by redox potential and dissolved sulfide level in sediments. The coupling between metal distributions derived by different approaches demonstrated the profile fidelity, although complexing ligand and kinetic resupply might influence the DGT/ centrifugation ratio. Taking advantages of high resolution DGT results, trace metal behaviors controlled by early diagenetic reactions were revealed. The organic matter oxidation led to surface maxima of Cu and Cd. The reductive dissolution of Mn/Fe (hydr)oxides in suboxic layer was obvious, to which Zn and Co corresponded. Sulfide accounted for scavenging trace metals, but its role was not all that significant, such as in an oxidizing state occurred in July at Site 4. Estimated exchange fluxes indicated the generally downward diffusion of trace metals at sediment-water interface, subject to the pollution condition in estuarine water, except for July when bacterial activity was promoted and profile peaks moved upward. Long-term monitoring would be important for better understanding the source and fate of trace metals in this dynamic system. Acknowledgement The authors thank for DGT products provided by DGT Research Ltd. This work was supported by a Key Project from the National Natural Science Foundation of China (21237004). Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2017.03.028. References Allison, J.D., Allison, T.L., 2005. Partition Coefficients for Metals in Surface Water, Soil, and Waste. Rep. EPA/600/R-05, 74. Berg, P., Risgaard-Petersen, N., Rysgaard, S., 1998. Interpretation of measured concentration profiles in sediment pore water. Limnol. Oceanogr. 43 (7), 1500e1510. Berner, R.A., 1980. Early Diagenesis: a Theoretical Approach. Princeton University Press, Princeton. Canavan, R.W., Van Cappellen, P., Zwolsman, J.J.G., Van den Berg, G.A., Slomp, C.P., 2007. Geochemistry of trace metals in a fresh water sediment: field results and diagenetic modeling. Sci. Total Environ. 381, 263e279. Charriau, A., Lesven, L., Gao, Y., Leermakers, M., Baeyens, W., Ouddane, B., Billon, G., 2011. Trace metal behaviour in riverine sediments: role of organic matter and sulfides. Appl. Geochem. 26 (1), 80e90.

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