Science of the Total Environment 517 (2015) 178–194

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Porewater dynamics of silver, lead and copper in coastal sediments and implications for benthic metal fluxes Linda H. Kalnejais a,⁎, W.R. Martin b, Michael H. Bothner c a b c

Ocean Processes Analysis Laboratory, and Department of Earth Sciences, University of New Hampshire, 8 College Rd, Durham, NH 03824, USA Marine Chemistry and Geochemistry, Woods Hole Oceanographic Institution, Woods Hole, MA 02543, USA United States Geological Survey, Woods Hole, MA 02543, USA

H I G H L I G H T S • • • • •

Porewater profiles of silver, copper and lead for two coastal sites are presented. High trace metal fluxes only occur in winter at sulfidic sites. The Cu diffusive flux contributes to the elevated metals in Boston Harbor waters. Trace metal enrichment can develop at the surface of sites with no sulfide. Metal contamination will persist at the surface due to this enriched layer.

a r t i c l e

i n f o

Article history: Received 4 November 2014 Received in revised form 4 February 2015 Accepted 4 February 2015 Available online xxxx Editor: F. Riget Keywords: Trace metal Early diagenesis Silver Benthic flux Coastal marine sediments

a b s t r a c t To determine the conditions that lead to a diffusive release of dissolved metals from coastal sediments, porewater profiles of Ag, Cu, and Pb have been collected over seven years at two contrasting coastal sites in Massachusetts, USA. The Hingham Bay (HB) site is a contaminated location in Boston Harbor, while the Massachusetts Bay (MB) site is 11 km offshore and less impacted. At both sites, the biogeochemical cycles include scavenging by Feoxyhydroxides and release of dissolved metals when Fe-oxyhydroxides are reduced. Important differences in the metal cycles at the two sites, however, result from different redox conditions. Porewater sulfide and seasonal variation in redox zone depth is observed at HB, but not at MB. In summer, as the conditions become more reducing at HB, trace metals are precipitated as sulfides and are no longer associated with Fe-oxyhydroxides. Sulfide precipitation close to the sediment–water interface limits the trace metal flux in summer and autumn at HB, while in winter, oxidation of the sulfide phases drives high benthic fluxes of Cu and Ag, as oxic conditions return. The annual diffusive flux of Cu at HB is found to be significant and contributes to the higher than expected water column Cu concentrations observed in Boston Harbor. At MB, due to the lower sulfide concentrations, the association of trace metals with Fe-oxyhydroxides occurs throughout the year, leading to more stable fluxes. A surface enrichment of solid phase trace metals was found at MB and is attributed to the persistent scavenging by Feoxyhydroxides. This process is important, particularly at sites that are less reducing, because it maintains elevated metal concentrations at the surface despite the effects of bioturbation and sediment accumulation, and because it may increase the persistence of metal contamination in surface sediments. © 2015 Elsevier B.V. All rights reserved.

1. Introduction The sediments of the coastal zone have been a sink for contaminants discharged to the environment for centuries. With improved environmental legislation, the direct discharge of pollutants has been significantly reduced in recent decades (Bothner et al., 1998; Luoma and Rainbow, 2008; Birch et al., 2013). Nevertheless, the legacy of contaminated sediments remains. Sediments have been shown to supply ⁎ Corresponding author. E-mail address: [email protected] (L.H. Kalnejais).

http://dx.doi.org/10.1016/j.scitotenv.2015.02.011 0048-9697/© 2015 Elsevier B.V. All rights reserved.

dissolved trace metals to the water column in quantities greater than riverine fluxes in many regions: Pb and Ag in San Francisco Bay, USA. (Rivera-Duarte and Flegal, 1994, 1997a); Cu and Pb in the northeastern Irish Sea (Williams et al., 1998); and Cu, Ni and V in the Vigo Ria, Spain (Santos-Echeandia et al., 2009). Despite the potentially important impact on coastal water quality, the mechanisms that determine if a metal is permanently sequestered in the sediment or remobilized to the water column is not well established. Trace metal behavior is controlled by the early diagenetic reactions within the sediments. Only dissolved metals can be released from stable (not eroding) sediment. The dissolved metals can be transported to the

L.H. Kalnejais et al. / Science of the Total Environment 517 (2015) 178–194

overlying water by diffusion, flushing by benthic organisms irrigating their burrows (Burdige, 2006) or advective flow in permeable sediments driven by waves, tides, and pressure gradients from microtopography (Huettel and Webster, 2001). Reactions that increase the dissolved metal concentration by converting solid-phase trace metals to the porewater phase can increase the metal release to the overlying water. In coastal sediments, these reactions are coupled to the oxidation of organic carbon, especially when Fe and Mn-oxyhydroxides and sulfate are the terminal electron acceptors. Specifically, mineralization of organic carbon can supply metals to porewaters (Heggie et al., 1986; Shaw et al., 1990; Audry et al., 2006), whereas trace metals are scavenged and removed from porewaters by Fe and Mn-oxyhydroxides (Burdige, 1993; Sundby, 2006). The reductive dissolution of these carrier phases can in turn release trace metals to porewaters (Shaw et al., 1990; Canavan et al., 2007). Sulfide, formed during sulfate reduction, precipitates insoluble trace metal phases (Rosenthal et al., 1995; Gobeil et al., 1997; Morse and Rickard, 2004), and the oxidation of sulfides can release dissolved metals. The relative importance of these opposing processes is unknown and it is difficult to predict the net effect of these reactions on trace metal cycling (Sundby, 2006). In the coastal zone, trace metal cycles are particularly challenging to predict due to continually varying conditions. The sediments are subject to strong seasonal variations in organic and terrigenous matter flux, water column stratification, temperature and benthic macroinvertebrate activity. The early diagenetic processes and redox zones within the sediments vary considerably over the course of a year, and so do the reactions competing for trace metals. Despite the variability in processes that impact metal mobilization, seasonal changes in porewater concentrations and benthic metal fluxes have been investigated in only a limited number of locations worldwide (Hines et al., 1984; Westerlund et al., 1986; Warnken et al., 2001; Gao et al., 2009; Lourino-Cabana et al., 2014). The stability of the seasonal variation in these studies is unknown, due to limited periods of observation, and the dominant controls on metal release and burial remain elusive. Here we present a record of trace metal porewater concentrations collected over seven years from contaminated sediments in Boston Harbor, USA, with the objectives of (1) quantifying the diffusive flux from sediments in different seasons; (2) establishing how consistent the seasonal fluxes are between years; and (3) determining the diagenetic reactions which influence the magnitude of the diffusive flux and the ultimate burial of trace metals. We show stable seasonal cycles in metal remobilization and benthic flux that are consistent between years, even as the sediment conditions evolve due to reduced contaminant discharge. The cycle identified in Boston Harbor is contrasted with seasonal porewater profiles from a less impacted offshore site in nearby Massachusetts Bay to assess the general applicability of the cycle. The important contaminant metals Ag, Pb, and Cu are investigated. All previously published coastal Ag porewater profiles are from San Francisco Bay (Rivera-Duarte and Flegal, 1997a; Huerta-Diaz et al., 2007). The data presented here extend Ag results to additional locations, providing new information on the distribution of an element that remains poorly understood in marine environments (Tappin et al., 2010). This long term record of trace metals in porewaters provides new insight into the fate of metals in contaminated coastal sediments. Our findings suggest longer persistence of metal contamination in the surface sediments of less reducing sites, due to diagenetic processes that maintain the metals at the surface, despite the effects of bioturbation and sediment accumulation. 2. Study sites Boston Harbor is an urban estuary in coastal Massachusetts USA, partially surrounded by the city of Boston (Fig. 1). The sediment and water quality of the Harbor have been degraded by more than 130 years of sewage discharge. In the 1980s the nutrient loads to the Harbor were among some of the highest in the world (Kelly, 1997),

179

and based on metal and persistent organic pollutant concentrations in the sediment, it was considered the most contaminated harbor in the United States (Bothner et al., 1998). A comprehensive modernization of greater Boston's wastewater treatment facilities was initiated in the 1980s to address the deteriorating sediment and water quality of the Harbor. The upgrade proceeded in stages, with sewage sludge discharge to the Harbor ending in 1991 and effluent discharge from the Nut Island wastewater treatment plant (Fig. 1) ending in 1998. In 2000, secondary treatment of sewage was implemented and discharge of treated sewage effluent was diverted from the harbor mouth to the Massachusetts Bay outfall, 15 km offshore (Taylor et al., 2011) (Fig. 1). Due to the reductions in sewage discharge, as well as improvements in chemical techniques to recover metals from industrial wastes, and the phaseout of leaded gasoline, the concentrations of metals in the sediments of Boston Harbor have decreased (Bothner et al., 1998; Zago et al., 2001). The concentration of metals in the water column of the Harbor however, did not decrease significantly between 1991 and 2002. A combined hydrodynamic-water quality model of the Harbor predicted dissolved concentrations of Pb and Cu that were less than 20% of measured values, which is likely due to benthic fluxes from contaminated sediments (Li et al., 2010). We conducted the study described here to determine the mechanisms that drive the benthic metal flux. Two fine-grained sediment sites (Table 1) with contrasting contamination histories were investigated (Fig. 1). Both are located in depositional areas of the Bay (Knebel et al., 1991), where the bottom water oxygen concentration is close to saturation throughout the year, with maximum oxygen concentrations in February (380 μM) and minimum in October (250 μM). The Hingham Bay site (HB) in Boston Harbor is shallow (5 m water depth) with 94% fine-grained sediment. The average proportion of fines in the surface sediment did not change between 2002 and 2008 (at the 95% level of certainty). There are historically high trace metal concentrations due to past sewage discharge from the nearby Nut Island wastewater treatment plant (Bothner et al., 1998). The average organic carbon contents in the surface sediment over 2002–2004 were 3.0 ± 0.2%, and 2.6 ± 0.1% in 2008 (average ± standard deviation, Table 1). The 2008 average value is statistically different from the earlier value (95% level of certainty), consistent with the observations of Tucker et al. (2014) that the surface organic carbon content has decreased over time due to the reductions in sewage discharge. Morford et al. (2007) determined that the average organic carbon oxidation rate at HB in 2001–2002 was 880 μmol C/cm2/yr, the oxygen penetration depth was less than 4 mm, bioirrigation was active at the site with a maximum in early summer, and bioturbation was seasonal, with a bioturbation coefficient of 3 cm2/yr in winter and 20–25 cm2/yr in June and October. The most abundant organism was the tubebuilding amphipod Ampelisca spp. with Leptocheris pinguis also important in 2008 (Tucker et al., 2014). There were 0.3–1 × 105 organisms/m2 retained on a 1 mm sieve in the spring and summer of 2003 (Benoit et al., 2006). Measurements of burrow surface area show burrows present to a maximum depth of 6 cm (Benoit et al., 2006) and the depth of bioturbation is also unlikely to extend below 6–8 cm, the depth where dissolved sulfide is detected in the porewaters (Morford et al., 2009). The sedimentation rate was 1.8 cm/yr in 1978 (Bothner et al., 1998) and has since declined considerably due to cessation of sewage discharge. The annual average sedimentation rate between 1978 and 2000 (estimated from 137 Cs profiles) was 0.6 cm/yr (Morford et al., 2007), which is likely biased by high sedimentation rates before 1991 when sewage sludge was still discharged to the Bay, so the sedimentation rate after 2000 is even lower. The second study site is 11 km offshore in Massachusetts Bay. This site (MB) is in 30 m of water and is 2 km west of the nearest diffuser port in the offshore sewage outfall (Fig. 1). There is 82% fine-grained sediment, 2% organic carbon and the average rate of organic carbon oxidation (600 μmol C/cm2/yr (Martin and Kalnejais, 2007)) is about twothirds that at HB. The oxygen penetration depth is less than 6 mm (Sayles and Goudreau, 2007). Bioturbation at MB is complex and not well described by biodiffusive mixing models (Crusius et al., 2004).

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MB HB

Massachusetts Bay Sewage Outfall

Mass Bay Site

Boston Deer Island

Boston Harbor Massachusetts Bay

Nut Island

Qu

inc

Hingham Bay Site

yB

ay

2 km

Fig. 1. Map of coastal Massachusetts showing sampling locations for the Hingham Bay (HB) and Massachusetts Bay (MB) sites.

There are diverse and abundant benthic organisms at MB, with 5 × 105 organisms/m2 retained on a 500 μm sieve in summer, with the polychaetes Spio limicola and Polydora socialis as the most abundant organisms (Wheatcroft et al., 1994). In winter the number of organisms falls by 50%. Tracer studies suggest that there are two modes of bioturbation, one that transports surface particles to N15 cm deep and the other that transports particles at depth to the surface (Wheatcroft et al., 1994). The upper limit for the sedimentation rate is 0.3 cm/yr, and more accurate estimates are not possible due to the extensive sediment reworking (Crusius et al., 2004). Porewater analysis from the upper 12 mm of sediment collected in 2002 and 2003 showed that the average Ag and Cu concentrations at HB were higher than at MB, while the average Pb porewater concentrations at both sites were comparable (Kalnejais et al., 2010). Despite comparable surface Pb concentrations between the offshore and Harbor site, MB does not show evidence of change associated with the relocation of the outfall. Bothner et al. (2002) found that the concentrations of Ag and Clostridium perfringens spores (a biological tracer of sewage) in sediments in 2001 were statistically equal to data collected in 1997–2000, before the outfall was Table 1 Site location, water depth and surface sediment (0–0.6 cm) characteristics. Location

Water depth

Massachusetts Bay (MB) 42.38981 N 30 m 70.8305 W Hingham Bay (HB) 42.29083 N 5m 70.92778 W a

Year

% sand/silt/claya

2002–2003

18/58/24

2 ± 0.2

2002–2004 2008

6/57/37 10/48/42

3.0 ± 0.2 2.6 ± 0.1

% Corgb

Data for 0–0.5 cm depth range. Average ± standard deviation (MB 2002–2003 n = 19, HB 2002–2004 n = 11 and HB 2008 n = 3). b

operating. Nevertheless, MB is impacted by northeast storms that transport contaminated sediment out of Boston Harbor and into Massachusetts Bay (Bothner et al., 2002; Warner et al., 2008). Thus the historical contamination in Boston Harbor does contribute trace metals to the MB site. 3. Materials and methods 3.1. Sediment sampling Between 2002 and 2008, HB was sampled 7 times and MB was sampled 3 times to obtain trace metal porewater profiles (Tables 2 and 3). Sampling times were selected so that different seasons of the year were investigated over the course of the study. Samples were collected from each season in multiple years whenever possible. Sediment samples were collected in a manner designed to minimize any disturbance to the sediment–water interface. Divers using SCUBA collected cores at HB, while a hydraulically-damped gravity corer (Bothner et al., 1998) was used at MB. Core barrels were polycarbonate with an inner diameter of 10.8 cm. Collected cores were placed on ice immediately after collection and brought back to shore. The cores for geochemical analysis were sectioned under nitrogen as soon as logistically possible (all sectioning and porewater processing was completed within 8–10 h). For HB samples, the cores were driven to Woods Hole, MA on ice, and sectioned in a 4 °C walk-in refrigerator, while the MB samples were sectioned in a van parked on the dock in Boston at ambient air temperature. The cores were collected with at least 30 cm of overlying water to ensure that the oxygen concentration did not change significantly over the time it took to process the samples. While trace metal benthic fluxes are sometimes measured with benthic flux chambers (Westerlund et al., 1986; Warnken et al., 2001; Morford et al., 2007), here we have chosen to investigate the diffusive flux with high resolution porewater profiles. At sites with shallow

L.H. Kalnejais et al. / Science of the Total Environment 517 (2015) 178–194

181

Table 2 Sampling times, depth at which sulfide is detectable (zs) and depth scaling factor (%) for the normalized depth axis (Z⁎) for each Hingham Bay profile. Winter

zs (cm)

Depth scaling

Summer

zs (cm)

Depth scaling

Autumn

zs (cm)

Depth scaling

Jan 2002 core 1 Jan 2002 core 2 Jan 2008 core 1

7.0 6.1 10.0

9 21 −30

July 2002 core 1 July 2002 core 2 June 2004 core 1 June 2004 core 2

8.0 6.5 6.0 7.9 7.1

−13 9 15 12

Sept 2002 core 1 Sept 2002 core 2 Oct 2003 core 1 Oct 2006 core 1

5.6 6.3 5.2 5.3 5.6

0 −12 7 5

Average

7.7

oxygen penetration and rapid organic matter remineralization rates, such as those investigated here, benthic chambers can measure artificially high metal fluxes (Morford et al., 2007). This is because at these sites, oxygen is rapidly depleted in the benthic chamber and the oxygen penetration depth decreases. The result is a breakthrough of dissolved Fe and any species that are affected by Fe-oxyhydroxide reduction. The porewater approach selected here allows us to both investigate diffusive fluxes and determine the reactions occurring in the sediments. The sediments were sectioned under nitrogen at a resolution of 3 mm for the first 1.8 cm, 6 mm for the next 3.6 cm and then the resolution decreased with depth to a maximum of 2 cm by 20 cm deep. The high resolution close to the surface was achieved by completely filling scintillation vials (for 3 mm intervals), followed by 50 mL centrifuge tubes (for 6 mm intervals) with a spatula and calculating the sample interval by volume. Sectioned sediments were centrifuged at 5000 rpm for 15 min. The centrifuge vials were returned to a nitrogen atmosphere where the porewater was removed and filtered through 0.45 μm polycarbonate syringe filters. The filtered porewater was distributed into aliquots for analyses of trace metals, nutrients, and sulfide. The porewaters for trace metals were filtered directly into 4 mL vials preacidified with 40 μL of trace metal grade (Fisher Optima) nitric acid, and instantly preserved at a pH of less than 1. The samples for sulfide were transferred into vials containing a solution of zinc acetate in 6% sodium hydroxide to sequester the sulfide as ZnS. The solid phase samples were frozen at − 40 °C after sectioning and porewater removal was completed. All sampling equipment in contact with the samples for trace metal analysis was acid cleaned, rinsed in 18.2 MΩ Milli-Q water (MQ) and dried in a laminar flow hood prior to use. Syringe filters for

trace metal analysis were cleaned by filling each filter with 2 N HCl for 1–2 h, followed by copious rinsing with MQ water and a final air purge. 3.2. Porewater and solid phase analysis All metal analyses were performed under trace-metal clean conditions in a laminar flowhood with overhead HEPA filtered air and acidcleaned plasticware. Acidified porewater samples were analyzed for Fe and Mn with either a Hitachi graphite furnace atomic adsorption spectrophotometer (GFAAS), quantified with a matrix-matched 5-point external calibration, or with a Finnigan Element 2 inductively coupled plasma-mass spectrometer (ICP-MS), operated in a medium resolution with an indium internal standardization based on the method of Rodushkin and Ruth (1997). Isotope dilution mass spectrometry (IDMS) was used to quantify the Ag, Pb, and Cu in acidified porewater samples. Pre-concentration of the metals was necessary and the cobalt–ammonium pyrrolidine dithiocarbamate (Co–APDC) method developed by Boyle and Edmond (1977) was used. This method relies on the co-precipitation of the Co–APDC complex to extract the metals from the porewater matrix, and is described in detail in Kalnejais (2005). Briefly, the APDC was purified with repeated chloroform extractions, the cobalt solution was purified with an AG1-X8 resin and ammonium hydroxide was purified by headspace distillation. An isotope spike of 109Ag, 65Cu and 208Pb (Oak Ridge National Laboratory) was added to each sample, and allowed to equilibrate for 48 h. To co-precipitate the metals, 200 μL of 2.5 mM cobalt solution, 100 μL of distilled ammonium hydroxide and 100 μL of 2% APDC solution were added to 1400 μL of porewater. The resulting precipitate

Table 3 Sampling times, water temperature and calculated diffusive flux across the sediment–water interface for Hingham Bay and Massachusetts Bay sites. The modeled benthic flux for Boston Harbor (Li et al., 2010) is also given. Flux (μmol/m2/yr) Hingham Bay Winter January 02 core 1 January 02 core 2 January 08 core 1 Summer June 04 core 1 June 04 core 2 July 02 core 1 July 02 core 2 Autumn September 02 core 1 October 03 core 1 October 06 core 1 Annual average (μmol/m2/yr) b Modeled benthic flux (μmol/m2/yr) (Li et al., 2010) Massachusetts Bay Feb 02 core 1 May 02 core 1 Oct 02 core 1 Annual average (μmol/m2/yr)b,c a

Water temperaturea

Ag

Pb

Cu

Fe

Mn

5.5 5.5 2.0

11 20 13

0 0 69

440 460 420

0 0 0

na 52 0

18.9 18.9 19.7 19.7

0 0 3 2

−4 0 12 16

−46 42 40 36

240 320 220 260

56 97 na 170

20.9 13.4 13.6

2 4 0 5±2

0 0 0 8 ± 10 50–90

68 210 110 170 ± 30 700–1300

320 0 290 170 ± 80

na 140 290 130 ± 50

4.7 9.6 12.1

2 3 0 2±1

0 7 0 2±4

180 240 80 160 ± 50

0 0 0 0

na na 260 na

Daily average water temperature from nearby NOAA Buoys; Hingham Bay (Buoy BHBM3) and Massachusetts Bay (Buoy 44013). Annual Average = 1/12 × (3 × winter flux + 4 × summer flux + 5 × autumn flux) Error is calculated by propagating the standard deviation of the seasonal averages through the annual average estimate. c Due to similar flux magnitudes, standard deviation of HB autumn fluxes is used to estimate uncertainty in Cu and Ag fluxes, standard deviation of HB summer fluxes is used to estimate Pb flux uncertainty. b

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was digested with two additions of 25 μL of concentrated nitric acid (Fisher Optima), diluted with MQ water and analyzed with a Finnigan Element 2 ICP-MS. The water overlying the core (sampled at 10 cm above the interface) was also analyzed with the Co–APDC method and 5 to 10 mL of seawater was extracted. The accuracy of the analyses was tracked with CASS-4 standard seawater (Nearshore Seawater Reference Material, National Research Council Canada). The Cu value measured for the standard seawater over 9 days (9.4 ± 0.2 nM) agreed with the certified value of 9.3 ± 0.9 nM. There is no certified value for Ag, but the measured value (58 ± 1 nM) is consistent with the value of 56 ± 3 measured by Ndung'u et al. (2001). The value measured for Pb was 70 ± 9 pM, on the high end of the certified value of 47 ± 17 pM. The porewaters were at least a factor of 10 higher than the standard seawater values and the accuracy of the analysis at this concentration was confirmed with standard additions of a commercial standard (Alfa Aesar) to one sample per experiment. Detection limits for the porewater analysis were 6, 12 and 350 pM for Ag, Pb, and Cu, respectively, and were constrained by instrument and reagent blank values. Sulfide was determined by the colorimetric method of Cline (1969); the detection limit was 1 μM. Porosity and resistivity were determined for each cruise on an additional core collected at the same time. Resistivity was determined at a resolution of 0.062 cm with a probe with an electrode spacing of 1 mm (Andrews and Bennett, 1981). Two resistivity profiles per core were collected following which the undamaged half of the core was sectioned for porosity. Porosity was determined gravimetrically by drying the sediments to a constant weight in a 60 °C oven and assuming a grain density of 2.6 g mL−1 (Morford et al., 2009). The total concentration of metals in the solid phase of the sectioned sediments was analyzed at the United States Geological Survey in Denver with a total digestion method (Briggs and Meier, 1999). Briefly, 0.2 g of dried sediment is digested with additions of concentrated hydrochloric acid, nitric acid, perchloric acid and hydrofluoric acid. All solid phase data are corrected for salt content. Certified reference materials including marine sediment BCSS-1 and MESS-1 and harbor sediment PACS-2 from the National Research Council Canada were run with each batch of 50 samples. The data for Fe, Mn, Cu, Ag and Pb were all within the published values for these elements. Organic carbon was determined on ground sediment that had been leached with dilute sulfurous acid to remove calcium carbonate. The sample was then combusted in a Perkin Elmer 2400 carbon–hydrogen–nitrogen analyzer. Acid volatile sulfide was determined with the purge-and-trap method of Allen et al. (1993).

3.4. Calculation of diffusive fluxes The diffusive fluxes were quantified from the porewater profiles using Fick's First Law of diffusion for sediments (Berner, 1984), J ¼ −ϕDs

∂C ; ∂z

ð2Þ

where J is the diffusion flux per area of total sediments, ϕ is the sediment porosity, Ds is the whole sediment diffusion coefficient, C is the porewater concentration and z is the depth below the sediment– water interface. The diffusion coefficient in water at infinite dilution D was used to determine the whole sediment diffusion coefficient Ds by accounting for the tortuosity θ within the sediments with Ds ¼ θD2 . Following the method of McDuff and Ellis (1979), a relationship between the tortuosity and porosity of the sediment was determined from resistivity and porosity measurements to derive values for Ds for each sampling date as a function of depth. The values for D for each metal are taken from Li and Gregory (1974). Due to rapid changes close to the sediment–water interface, the concentration gradient was estimated from the porewater profiles, with the assumption that the gradient across the sediment–water interface is one dimensional and can be approximated as dC C pw1 −C olw ; ≈ Δz dz

ð3Þ

where Cpw1 is the porewater concentration in the uppermost porewater sample (0–3 mm). Colw is the concentration measured in the water overlying the core. Δz is taken as half the sampling resolution (1.5 mm). This assumes that the benthic boundary layer is thinner than Δz cm and can be neglected. A positive flux represents a release from the sediments to the overlying water. Variations in the porewater concentration within the upper 3 mm cannot be identified by this sectioning method and may introduce inaccuracies into the calculated fluxes. If an undetected thin oxic layer exists at the sediment–water interface, it could trap upwardly diffusing metals at the sediment–water interface (Sundby et al., 1986; Westerlund et al., 1986). Therefore the fluxes calculated from Eq. (3) represent an upper limit for the diffusive release from sediments. 4. Results

3.3. Analysis of porewater profiles

4.1. Porewater iron, manganese, and sulfide

The sediments are heterogeneous environments (Davison et al., 1997; Fones et al., 1998), so to extract the most important features from the large number of porewater profiles collected from HB, seasonally-averaged porewater profiles were calculated. The porewater profiles collected over seven years at HB are compiled into three seasonal groups (Table 2). To determine average profiles for each season from the cores with different sampling intervals, the data points were linearly interpolated at 0.5 mm depth intervals. The redox zone depth changes between cores, so to ensure that averages were only calculated between points in the same redox zone, the depth data were first transformed by

As the benthic flux of trace metals is strongly influenced by the redox conditions in the sediments, we first present the porewater data for the major redox species at each site to establish how the conditions change over the course of a year. The porewater profiles of Fe, Mn, and sulfide for HB are shown in Fig. 2. Dissolved Mn is measurable in the surface porewater sample (0–3 mm depth) in all seasons and reaches a maximum concentration at the surface in July, and in the upper 4.5 cm at all other sample times. The dissolved Fe maximum is typically deeper than the Mn maximum, with the maximum Fe concentration between 0.9 and 6 cm deep depending on the season. Both Fe and Mn are removed from the porewaters with depth, with Fe removed at shallower depths than Mn. In all profiles, except in January 2008, the Fe is removed from porewaters by 7 cm, while the Mn persists down to 10 cm. In the 2008 profile, Mn and Fe are both measurable to 10 cm. Once all the Fe is removed from the porewaters, sulfide builds up, typically reaching a maximum concentration between 10 and 15 cm deep. At MB, Mn is also released to the porewaters shallower than Fe (Fig. 3). The Mn concentration decreases sharply with depth, but Mn is never completely removed from the porewaters in the two cores measured. The Fe in the porewaters at MB persists to the maximum




Z ¼

z z ; zS S

ð1Þ

where Z⁎ is the redox zone normalized depth, z is the true sampling depth in a profile, zs is the depth in the profile where sulfide is first detected and zS is the average depth where sulfide first appears within the seasonal group. This normalization allows us to compare seasonal data over many years at HB, even as the redox conditions in the sediments change due to cessation of sewage discharge. The normalization was typically less than 15% but was 30% for the 2008 profile (Table 2).

L.H. Kalnejais et al. / Science of the Total Environment 517 (2015) 178–194

183

Mn Concentration (μM) 0

0

10

(a)

Depth (cm)

0

8

16

(c)

(b)

5

Winter

20

Iron (bottom axis) Sulfide (bottom axis)

10 Manganese (top axis) 15 20

Jan 2002 c2

Jan 2008

600

1200

0

250

500

0

250

500

0

10

20

0

15

30

0

10

20

(d)

(e)

(g)

(f)

5 Depth (cm)

Summer

0

Jan 2002 c1 0

10 15 20

Jun 2004 c1 0

450

0

450

(h)

Jul 2002 c1 900

Jul 2002 c2

0

500

1000

0

500

1000

0

10

20

0

20

40

(i)

(j)

5 Depth (cm)

Autumn

0

Jun 2004 c2 900

(k)

10 15 20

Sep 2002 c1 0

250

Sep 2002 c2 500

0

250

Oct 2003 500

0

250

Oct 2006 500

0

500

1000

Fe or S Concentration (μM) Fig. 2. Seasonal variation at Hingham Bay of porewater iron, manganese, and sulfide. Each panel presents data from a single core. Each row is data from the same season. Manganese was not measured on all cores. Data below the detection limit are plotted at zero concentration.

sampling depth of 35 cm. No dissolved Fe was detected in the uppermost porewater samples, indicating that Fe is internally recycled within the sediments at MB. No dissolved sulfide was detected in any MB porewaters. There is a distinct seasonal oscillation in the depths of the redox zones over the course of a year in the sediments of HB. The seasonal variability is best observed in the average seasonal profiles (Fig. 4), where averaging smooths the sharp maxima and minima observed in Fig. 2. The zs values used with Eq. (1) to normalize each profile (Table 2) indicate that the depth to the sulfidic zone varies by up to 2 cm within each seasonal group. Much of this variation is due to heterogeneity at the sampling site, as observed by up to 2 cm of variation in zs between cores collected at the same time (e.g. the June 2004 cores). The individual profiles that comprise each seasonal average are also presented in Fig. 4(c)–(e) to allow comparison with Fig. 2, and show that normalizing the data with Eq. (1) has a minimal impact on the individual profiles, except for the 2008 data (Fig. 4c), where the normalization compresses the profile by 30% (Table 2) relative to the measured data. The variation between seasons in HB is most evident in the upper few cm of the sediment (Fig. 4a). In winter, dissolved Fe is not detected in the top 1 cm. The Fe concentration increases below 1 cm, reaching a maximum around 2 cm in the cores from 2002 to 6 cm in the 2008 winter core. The cores collected in summer show the profile moving upwards, with measurable Fe in the surface porewaters and a steepening of the Fe gradient in the upper 1 cm of sediment (Fig. 4a). By autumn, the Fe profile has moved further upwards and the steepest surface gradient in Fe is observed. There is no distinct seasonal variation in

the Fe profiles at MB (Fig. 3). The depth that Fe is first released to the porewaters and the upper Fe gradient are relatively constant. 4.2. Solid phase data The HB data show constant concentrations of solid phase trace metals through the top 2 cm (Fig. 5), with average concentrations of 25 ± 2, 1100 ± 50, and 380 ± 10 μmol/kg for Ag, Cu and Pb respectively (average ± standard deviation). Below 2 cm the concentrations increase with depth due to the historical contamination at the site (Bothner et al., 1998) to an average of 70 ± 7, 1800 ± 120, and 600 ± 70 μmol/kg between 15 and 25 cm for Ag, Cu, and Pb, respectively. The Fe, Mn and organic carbon concentrations are more constant, with concentrations of 730 ± 20, 11 ± 0.5 mmol/kg and 3.1 ± 0.2% respectively, over the entire core length. In MB, all metals show a surface maximum extending down to 3 cm, with lower concentrations over 5–15 cm (Fig. 5). The concentrations in the surface are 20–40% greater than in the sediments below. The surface concentrations of Pb, Fe, and Mn are within 10% of those at HB, while the Ag and Cu concentrations are lower, at 24% and 66% of the HB surface values, respectively. The concentrations of all metals except Ag increase below 15 cm to levels comparable to the surface concentrations. There is an organic carbon surface maximum of 2.3 ± 0.1% over the top 3 cm at MB, that declines to low values of 1.6 ± 0.2% below 7 cm. The MB surface metal data in Fig. 5 are statistically equal to data collected previously at this site (Bothner et al., 2002, 2007), providing no evidence at the time of sampling that the outfall relocation altered the metal content. The acid volatile

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Fe Concentration (μM) Fig. 3. Seasonal variation at Massachusetts Bay of porewater iron and manganese. Sulfide was not detected in any core. Manganese was only measured on September 2003 cores. Each panel presents data from a single core. Each row is data from the same season.

sulfide (AVS) concentration is measurable in the surface sample at HB (Fig. 5). The AVS levels increase to a maximum at 5 cm depth, with one sharp minimum at 4.2 cm, and then have an average value of 5.9 ± 0.8 mmol/kg between 6 and 19 cm. There is no AVS in the surface samples at MB, but below 1 cm there is detectable AVS that reaches a maximum of 3 mmol/kg at 2.7 cm and then declines to an average value of 0.8 ± 0.7 mmol/kg. 4.3. Porewater silver, copper, and lead The average seasonal porewater profiles show that the concentrations of Ag ranged from 0.04–1.2 and 0.01–0.6 nM for HB (Fig. 6) and MB (Fig. 7) respectively. The individual profiles for each trace metal and core are provided in the Supplementary information (SI) in SI — Figs. 2 to 14. The MB average porewater Ag concentration is lower than the HB Ag average (at the 95% level of certainty). Ag data from both sites fall within the range of values found by Rivera-Duarte and Flegal (1997a) in San Francisco Bay and Morford et al. (2008) on the Washington Margin in water depths of 400–3800 m. The Cu concentrations range from 4–45 and 8–43, and the Pb concentrations range from 0.3–24 and 0.6–9 nM in HB (Fig. 6) and MB (Fig. 7) respectively. The average porewater concentrations of Cu and Pb at each site are statistically equivalent (at the 95% certainty level). These values fall within the range reported for a variety of aquatic environments and human impacts including the Tyrrhenian Sea (Ciceri et al., 1992), Laurentian Trough (Gobeil and Silverberg, 1989), San Francisco Bay (Rivera-Duarte and Flegal, 1994, 1997b), the coastal North Sea (Gao et al., 2009), the Gironde Estuary (Audry et al., 2006, 2007), Bolivian Altiplano lakes (Tapia and Audry, 2013) and an Iberian Ria (Santos-Echeandia et al., 2009). The trace metal porewater concentrations show distinct variations with depth and season. The behaviors in three distinct zones of the sediment are discussed below. The first zone is the surface porewaters, which include the upper 1–2 porewater samples (0–6 mm) that determine the gradient across the sediment–water interface. The second zone is the ferruginous chemical zone (Canfield and Thamdrup, 2009)

in which Fe is present in the porewaters. This zone is generated due to the biotic and abiotic reductions of Fe oxides, with the lower bound of the zone controlled by Fe saturation with Fe minerals, including Fe sulfides (Canfield and Thamdrup, 2009). The third zone is the sulfidic chemical zone, with sulfide in the porewaters. 4.3.1. Surface porewaters The surface porewater concentrations change considerably over the seasons at HB. In winter, the sediments show a distinct release of Ag and Cu at the surface and the concentration is higher than in the overlying water (Fig. 6), indicating a winter Ag and Cu diffusive flux to the water column. In summer and autumn, there is typically a diffusive Cu flux, but the winter flux is 2–10 times higher than in the other seasons (Table 3). A diffusive Cu flux into the sediments is calculated for a single core in summer. There is either minor or no Ag flux in summer and autumn (Table 3). The Pb profiles indicate limited diffusion across the sediment–water interface in any season, except in January 2008 (Table 3, SI — Fig. 4), with a surface Pb concentration significantly higher than the overlying water. This core dominates the average Pb profile but the two earlier winter cores show no evidence of Pb release. At MB (Fig. 7), the behavior of each metal in the surface porewaters is similar to HB. Cu shows a strong release at the sediment–water interface throughout the year. A minor release of Ag is observed in the winter and late spring, while a significant release of Pb at the interface is not observed. 4.3.2. Ferruginous zone At HB in summer there is an accumulation of trace metals in the porewaters of the ferruginous zone that is coincident with the accumulation of Fe (Fig. 6). The trace metal profiles track the Fe profile with a linear relationship, where the correlation coefficients for all summer profiles pooled together are r = 0.56 (Ag), r = 0.57 (Pb) and r = 0.67 (Cu), with N = 47, and all are significant at p b 0.05. There is no significant relationship between the trace metal and Mn concentrations (for all summer profiles pooled together r = − 0.15 (Ag), r = − 0.03 (Pb) and r = 0.22 (Cu), N = 69). By contrast, both autumn and winter

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Concentration (μM) Fig. 4. Seasonal variation of porewater iron and sulfide at Hingham Bay. Depth axis is the redox zone normalized depth, Z⁎, described in Section 5.2. Normalization typically alters the depth by less than 15% (but is 30% for 2008 data), scaling factors for each core are given in Table 2. (a) Average seasonal concentration of iron and sulfide. (b) Inset of (a) showing detail in the top 8 cm. Shading is one standard deviation on either side of the seasonal average. (c–e) Data from each core and average displayed by season. The average profile and standard deviation for autumn do not include the 2006 data as the Fe concentrations are clearly different from the earlier data.

profiles at HB show no significant relationship between Fe and trace metal concentrations. The cores from September 2002 and January 2008 (SI — Figs. 4 and 9) have relatively constant, low metal concentrations throughout the ferruginous zone. The January 2002 and October 2006 profiles show a release of trace metals coincident with the first release of Fe to the porewaters (e.g. January 2002 core 1 at 1 cm, core 2 at 1.2 cm, October 2006 at 1.4 cm (SI — Figs. 2, 3 and 11). Unlike the summer profiles, however, the relationship does not continue with depth. At MB (Fig. 7), there is a significant linear relationship between porewater Ag and Pb (but not Cu) with Fe concentrations in the upper 5 cm throughout the year (all cores pooled, r = 0.44 (Ag), r = 0.43 (Pb), N = 31, significant at p b 0.05). Below 5 cm, there is a secondary maximum in Fe concentration in each core, but there is no significant relationship between the trace metals and Fe. Instead the concentrations

of the three trace metals largely stabilize below 7 cm, except for the winter core, which shows more variability with depth.

4.3.3. Sulfidic zone A sulfidic zone is only found within our sampling range of 0–30 cm at HB, and it begins once the dissolved Fe is almost completely removed from the porewaters and is at or approaching the Fe detection limit. Between the depths where Fe is first below the detection limit and the sulfide inflection point, there are sharp maxima in the trace metal concentrations (e.g. January 2002 core 2 at 7 cm (SI — Fig. 3) and July 2002 core 1 at 8 cm (SI — Fig. 6)). Deeper in the sulfidic zone, the trace metal profiles are smooth and stable with depth in most cores. The fluctuation seen in the average winter Pb profile is driven by three high values in

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Fig. 5. Solid phase metal, organic carbon and acid volatile sulfide (AVS) concentrations for both sites. The Hingham Bay metal and organic carbon data are from July 2002 core 2 and the Massachusetts Bay data are from February 2002 core 3. The acid volatile sulfide (AVS) data are from Hingham Bay June 2004 core 1 and Massachusetts Bay February 2002 core 1. Note the silver and organic carbon data from MB are plotted against the top axis to show detail (the arrows indicate which axis to read from), all other data are plotted against the lower axis.

the January 2002 core 1 (SI — Fig. 2), while the other winter cores show stable concentrations with depth. 5. Discussion 5.1. Early diagenesis of iron and manganese The appearance of dissolved species in the sediments of HB and MB follows the same sequence as previously observed in pelagic sediments, due to the remineralization of organic matter (Froelich et al., 1979; Canfield and Thamdrup, 2009), but with a compression of redox zones, characteristic of organic rich environments (Sorensen and Jorgensen, 1987). Anaerobic remineralization dominates, as evident by the increasing concentrations of porewater Fe and Mn below the sediment–water interface due to the reductive dissolution of Fe and Mn–oxyhydroxides. Porewater sulfide is only found at HB, but sulfate reduction occurs at both sites, as indicated from the presence of AVS in the sediments (Fig. 5). The persistence of dissolved Fe at greater depths, and absence of porewater sulfide at MB reflects the lower rate of organic carbon oxidation at MB compared with HB, resulting in lower rates of sulfate reduction and sulfide formation. In addition, the activity of benthic organisms at MB forming burrows deeper than 15 cm (Wheatcroft et al., 1994) supplies oxidants at depth and keeps sulfide concentrations low (Emerson et al., 1984). Many of the Fe porewater profiles from MB show wide excursions that are not found in replicate cores collected at the same time, indicating limited spatial extent (e.g. Fig. 3(c) shows a minimum at 10 cm). These rapid changes in dissolved Fe are likely due to the activity of benthic organisms (Gobeil and Silverberg, 1989). Two annelid worms were found at 10 cm in May 2002 core 1 (Fig. 3c), a depth that coincides with the sharp reduction in Fe concentration. The worms irrigate their burrows and maintain a more oxygenated environment in the vicinity. Every core sectioned at MB contained at least three worms, found down to 35 cm. Fewer burrowing benthic organisms were encountered while sectioning the HB cores, but the sharp minima found in some profiles (e.g. Fig. 2(f) at 4 cm and Fig. 2(h) at 1.6 cm) are also attributed to the activity of benthic organisms. The net effect of bioturbation and

irrigation is complex and likely contributes to the variability observed between replicate cores. Dissolved Mn and Fe are removed from deep porewaters by different mechanisms. At HB, the impacted Harbor site, due to high rates of dissolved inorganic carbon production, a Mn carbonate phase of the form MnxCa(1 − x)CO3 removes Mn (SI Section 1.2), similar to the Mn phase forming in Narragansett Bay (Elderfield et al., 1981). At MB, the offshore site, some phase other than carbonate must be controlling the Mn removal as there is no relationship with carbonate and all the samples are undersaturated with respect to rhodochrosite (SI Section 1.2). At HB, there is a tight coupling between the Fe and sulfide profiles, with sulfide detected only once the Fe concentration is close to or below the detection limit, suggesting that FeS is forming. These conditions are the same as those found in Long Island Sound where, even with active sulfate reduction, the concentration of sulfide in porewaters is maintained at very low levels until all reactive Fe is consumed (Canfield, 1989). The presence of AVS in the surface sediments at HB (Fig. 5) suggests that sulfate reduction does occur close to the sediment–water interface even though porewater sulfide is only detected below 7 cm. At MB, benthic organism activity introducing oxygen to sediments is likely responsible for the removal of some Fe from the deep porewaters as Fe-oxyhydroxide precipitates. In addition the porewater data are consistent with the removal of Fe from the porewaters by the precipitation of vivianite, Fe3(PO4)2·8H2O. The ion activity product is 1–10 times greater than the Ksp for this mineral phase below 5 cm deep (SI Section 1.3). Equilibrium with vivianite is also found in Long Island Sound (Martens et al., 1978). The removal of Fe as a phosphate phase at MB, compared to a sulfide at HB, is consistent with the higher organic carbon oxidation rates and lower benthic organism activity at HB generating more sulfide. 5.2. Seasonal variation in redox conditions and long-term changes The seasonal cycle in porewater chemistry at HB is driven by the temperature and the flux of organic matter. In January, when the water temperature is between 2 and 5 °C (Table 3), the sediments are at their most oxidized, with a distinct zone in which there is no measurable dissolved

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Lead Concentration (nM) Fig. 6. Seasonal variation in silver, copper and lead at Hingham Bay from all cores collected between 2002 and 2008. Depth axis is the redox zone normalized depth, Z⁎, described in Section 5.2. Normalization typically alters the depth by less than 15% (but is 30% for 2008 data), scaling factors for each core are given in Table 2. Solid line is the seasonal average and shaded area is one standard deviation on either side of the average. The average iron profile from Fig. 4 is shown as the dashed line in each graph.

Fe and no production of dissolved Mn. Low reaction rates and a limited supply of fresh organic matter in winter reduce the demand for oxidants, so a surface zone in which Fe and Mn-oxyhydroxides are not reduced becomes established. Although dissolved oxygen was not measured, the oxic zone lies above the zone of Mn production (Canfield and Thamdrup, 2009), so we infer that it was less than 5 mm in 2002 and less than 20 mm in 2008. By summer, the water temperature rises to almost 20 °C and productivity increases in Boston Harbor (Keller et al., 2001). Oxidants in the sediments are consumed at a faster rate and are depleted closer to the sediment–water interface. Dissolved Mn is produced in the top 3 mm (the sampling resolution), thus the oxic zone must be less than this thickness. By autumn, the average temperature has cooled to 13 °C and the primary productivity has fallen from the summer high (Keller et al., 2001). Nevertheless, the steepest Fe gradient and the shallowest sulfidic zone (Table 2), indicate that the demand for oxidants is still high.

There is no distinct seasonal signal in the Fe porewater profiles at MB, despite an annual temperature variation of 6 °C in the bottom waters (Butman et al., 2007) and annual variation in phytoplankton productivity (Keller et al., 2001). Sayles and Goudreau (2007) suggest that there is a small, but detectable seasonal variation in both the oxygen flux and oxygen penetration depth, from 400 ± 130 to 650 ± 210 μmol/cm2/yr and from 4.9 ± 0.8 to 3.6 ± 1.1 mm in winter and autumn, respectively. The absence of a seasonal signal in porewater Fe is probably due to the abundant benthic organisms obscuring the small changes. The porewater data from HB in 2008 are different from earlier data. The maximum depth of dissolved Fe is 3–4 cm deeper than the January cores collected six years earlier (Table 2). This deepening of the ferruginous zone is likely due to geochemical changes at the site associated with the cessation of sewage discharge to the Harbor in 2000. This is consistent with a decrease in surface organic carbon content over 2002–2008 (Table 1), and the observations of Diaz et al. (2008) and

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Lead Concentration (nM) Fig. 7. Seasonal variation in silver, copper and lead at Massachusetts Bay from all cores. Note, unlike Fig. 6 this data are not averaged as only single profiles are available.

Tucker et al. (2014) that since 1992 the benthic habitats in the Harbor have shifted to more aerobic states, due to the outfall relocation and a decrease in organic carbon supply. The October 2006 data also show signs of changing sediment conditions. Although the redox zone depths are comparable to those collected in the same season in earlier years, the porewater Fe concentration in 2006 is more than twice that measured previously (Fig. 2). The higher Fe concentrations may reflect the transition to a more oxidizing sediment state, with lower rates of sulfate reduction, and less sulfide available to precipitate Fe. The continued evolution in sediment geochemistry associated with the cessation of wastewater discharge to the Harbor is likely and sediment cores should continue to be collected at HB to document the rate of sediment recovery and its effect on metal fluxes. 5.3. Trace metal porewater dynamics 5.3.1. Surface porewaters Release of metals at the sediment–water interface has been attributed to the degradation of biogenic material or particles that scavenged

metals from the water column (Shaw et al., 1990; Tapia and Audry, 2013). Cu is a bio-active trace metal so release from organic matter is a likely source of Cu in the surface porewaters at both sites. Ag is considered to associate with diatoms (Flegal et al., 1995), but diatoms are unlikely to be the most abundant settling material at HB in winter due to the proximity to terrestrial sources of sediments. Resuspension of the sediments occurs recurrently in Boston Harbor (Kalnejais et al., 2007) so the metal release may be recycled from benthic sources. Incubations of suspended HB sediment showed that up to 6% of the total Ag and Cu in particles can be released to the dissolved phase in oxygenated water (Kalnejais et al., 2010), thus settling of resuspended sediment and oxidation in surface sediment may also contribute to the high surface porewater concentrations of Ag and Cu. The strongest trace metal release at the interface occurs in winter suggesting that an in-situ source of the dissolved metals may also be important. The high porewater concentrations could result from the oxidation of reduced metal species (such as those associated with AVS phases) that were formed close to the surface, or mixed from below

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during the summer, and are oxidized as the redox horizons move downwards in winter. There is at least a 5 mm deepening of the dissolved Fe front between winter and summer (Fig. 4), which suggests that the oxygen penetration depth also deepened over winter. Any sulfide or other reduced species in the surface sediments in summer would become unstable in winter as oxygen penetrates downwards. Similar seasonal releases of dissolved Fe (Aller, 1980), Cu (Hines et al., 1984) and Cd (Metzger et al., 2007) have been observed in winter due to the oxidation of sulfides. The relative importance of recently settled material versus in-situ sources in driving the release of trace metals at the interface cannot be firmly established from this data. Nevertheless, the primary productivity in Boston Harbor is lowest in winter, when total suspended solids are at an annual low (Taylor et al., 2011). 5.3.2. Ferruginous zone Fe and Mn oxides are known to adsorb a large range of trace metals (Jenne, 1968) and the linear relationship observed here suggests that the trace metals were associated with Fe-oxyhydroxides and released to the porewaters during reductive dissolution of the Fe. Similarities in the porewater profiles between Fe, Ni, and V (Shaw et al., 1990) and Pb (Gobeil and Silverberg, 1989) have likewise been used to infer scavenging by Fe-oxyhydroxides in sediments. In marine systems, Cu has been observed to associate with Mn oxides rather than with Feoxyhydroxides (Klinkhammer, 1980; Shaw et al., 1990), but in this study there is no significant relationship between trace metal and Mn concentrations at either site. The adsorption of Ag by Fe-oxyhydroxides observed here is not observed in all marine environments. In sediments from the Washington Margin, Ag was released to the porewaters close to the surface, and there was no relationship with either porewater Fe or Mn concentrations (Morford et al., 2008). Wen et al. (1997) investigated a relationship between Ag and Fe in estuarine particulates by conducting a radiotracer experiment with Ag and hematite in seawater. They concluded that the association of Ag with Fe-oxides was minimal, attributing the relationship to Ag associating with Fe-sulfide phases. However, the results of Wen et al. (1997) do not discount Fe-oxyhydroxides adsorbing Ag in these sediments as hematite is a crystalline phase and the amorphous phase is the most reactive fraction of the Fe-oxide pool in coastal sediments (Canfield, 1989). The higher Fe content in coastal sediments, likely leads to the observation of Ag and Cu adsorption by Fe-oxyhydroxides in this study compared with those of open ocean sites (Klinkhammer, 1980; Shaw et al., 1990; Morford et al., 2008). The relationship between porewater trace metal and Fe at HB is only observed in summer, indicating that at other times of the year the trace metals associate with different phases. Precipitation as a sulfide above the sulfidic zone is likely an important removal mechanism. Due to the low solubility of Ag, Cu, and Pb sulfides (Morel and Hering, 1993), even if only trace levels of sulfide are present, the metals will precipitate out as a sulfide phase, which was proposed by Rosenthal et al. (1995) as the mechanism responsible for the removal of Cd in suboxic sediments. These authors suggest that even though there is no detectable sulfide in the porewaters, from thermodynamic calculations CdS is sufficiently insoluble that trace levels of sulfide (N0.01 μM) can lead to a supersaturation of the porewaters with respect to CdS. The Ksp values for Ag, Cu, and Pb sulfides are comparable to the Ksp for CdS (Morel and Hering, 1993), so this mechanism is just as applicable. The relative importance of Feoxyhydroxides versus sulfides as the dominant removal phase for the trace metals varies seasonally and will be discussed in Section 5.4. 5.3.3. Sulfidic zone The concentrations of porewater trace metals in the HB sulfidic zone below 10 cm are mostly stable between seasons (Fig. 6). Chemical equilibrium calculations were carried out with Visual Minteq (Gustafsson, 2005) to predict the speciation and ion activity product of the dissolved trace metals in the sulfidic porewaters. Additional aqueous sulfide complexes from the compilation of Saito et al. (2003) were added to the

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Minteq database. The concentration of the major ions in seawater, pH (calculated from alkalinity and dissolved inorganic carbon), sulfide, phosphate, and dissolved inorganic carbon were included in the speciation calculation. The measured total dissolved metal concentration compared to that calculated to be in equilibrium with a sulfide mineral (Ag2S — acanthite, PbS — galena and CuS — covellite) was 0.1–0.8 for Ag, and 2–100 for Cu and Pb. Thus Ag is slightly undersaturated, while Pb and Cu are oversaturated with respect to the sulfide minerals. However, this thermodynamic calculation has a high degree of uncertainty in sedimentary environments due to complications such as non-equilibrium conditions, co-precipitation with FeS (Morse and Arakaki, 1993), the formation of non-stoichiometric sulfides (especially Cu (Boulegue, 1983)) and the presence of strong organic and polysulfide ligands. Solubility products determined by different research groups can vary by a few orders of magnitude (Stumm and Morgan, 1981). An additional uncertainty arises for Cu because there are two possible oxidation states (Skrabal et al., 2000). Considering the uncertainties, the measured and calculated values are relatively close, indicating that these metals are near saturation with respect to sulfide minerals at depth in the HB cores.

5.4. The impact of seasonal oscillations on trace metal mobilization Due to seasonal migrations in redox boundaries, the concentrations of trace metals in the surface porewaters vary seasonally at HB, and therefore so too do the benthic fluxes. The seasonal patterns are remarkably consistent over the seven years of sampling. The fluxes calculated for 2008 fall within the range reported in earlier years. The only exception is Pb, which had a high flux in the winter of 2008, but a negligible flux for the other winter samples. A conceptual model of the mechanisms responsible for the seasonal variation is shown in Fig. 8, including three trace metal phases (dissolved; metal sulfide; and adsorbed to Feoxyhydroxides), reactions that transfer metals between each phase as a function of season, and processes that transfer material between redox zones (diffusion and bioturbation of particles). The seasonal migration of redox boundaries also transfers material between redox zones; to reduce complexity in the figure this is pictured together with bioturbation. In summer, the redox horizons shoal and Fe-oxyhydroxides that were stable over winter are reduced. Dissolved Fe and adsorbed trace metals are released into the porewaters so the trace metal profiles and Fe profiles are coupled. The upper trace metal maximum in summer has roughly equal gradients up and down. This suggests that the trace metals diffuse both upwards and are re-scavenged by Fe-oxyhydroxides, and downwards, where they are precipitated as sulfides (Fig. 8a). As summer progresses, the ferruginous zone continues to move upwards. Conditions are more reducing than earlier in the summer, and due to greater sulfide abundance, a larger fraction of dissolved trace metals precipitate as sulfides than re-adsorb to Fe-oxyhydroxides (Fig. 8b). Over the course of the summer there is a net transfer of trace metals from the oxide phase to the sulfide phase. The impact of this transfer can be seen in the autumn profiles, which shows no coincident release of trace metals and Fe, because trace metals are no longer associated with Fe-oxyhydroxides. The association of porewater trace metal and Fe breaks down as summer progresses, not only because there is more sulfide available, but also because Fe is 3–4 orders of magnitude more abundant than the trace metals. When Fe-oxyhydroxides are reduced, the dissolved Fe is removed from solution by oxidation or by sulfide precipitation. Fe is in excess relative to sulfide at shallow depths and FeS is more soluble than trace metal sulfides (Morel and Hering, 1993). A greater proportion of the Fe, compared to the trace metals, is thus recycled back into the oxyhydroxide phase. Over the course of the summer, Fe-oxyhydroxides may undergo multiple reduction–oxidation cycles (Canfield et al., 1993), and each time they will have fewer trace metals associated with them. By autumn the Fe-oxyhydroxides that are reduced appear to have completely lost all their associated trace metals.

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(a) Summer

Porewater Concentration Dissovled Ag Ag-Fe Oxides

Ag2S

oxidation

precipitation

O2

adsorption

Depth (cm)

Sulfidic Ferruginous

Depth

Oxic

Settling Flux

reduction of FeOOH

Fe Ag

S(-II) precipitation Reaction Solid phase transport/ redox zone movement

burial

burof ialsolute phase Diffusion Dominant seasonal pathway

(b) Autumn Ag2S

Dissolved Ag Ag-Fe Oxides

oxidation

precipitation

adsorption

Depth (cm)

Sulfidic Ferruginous

Depth

Oxic

Settling Flux

Ag

Fe

reduction of FeOOH

S(-II)

precipitation

burial

burial

(c) Winter Ag2S

Dissolved Ag Ag-Fe Oxides

adsorption

Ferruginous

precipitation

reduction of FeOOH

precipitation

burial

Porewater Concetration

O2 Depth (cm)

Oxic

oxidation

Sulfidic

Depth

Settling Flux

Ag

Fe

S(-II)

burial

Fig. 8. Conceptual model to account for the seasonal variation in trace metal mobilization and benthic flux at Hingham Bay (HB). The figure is drawn for silver, but also applies to copper and lead. Each box represents a metal phase within the sediments for each redox zone. The arrows represent the mechanisms by which material can be transferred between boxes. Solid phase transport includes sedimentation and bioturbation. An idealized porewater profile is also provided. Oxygen was not measured in this study and the oxygen profile is inferred from the iron porewater profiles.

In winter, reduced species formed close to the surface, or mixed upwards over summer, are oxidized as oxygen penetrates downward. The surface porewater concentrations of Ag and Cu become high enough for

the metals to diffuse across the sediment–water interface (Fig. 8c). This was also observed for Pb in January 2008. The surface dissolved trace metal maximum also drives diffusion of metals downwards into the

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sediments, where the metals can be scavenged by freshly precipitated Fe-oxyhydroxides (Fig. 8c), resetting the seasonal cycle. The low Pb concentration typically found in the surface porewaters suggests that Pb has the strongest affinity, or adsorbs rapidly, to Fe-oxyhydroxides, suppressing its diffusive flux throughout the year. The high Pb flux in the winter of 2008 may be another consequence of the sediments becoming more oxidizing. Further sampling is required to determine if there is a persistent increase in the benthic fluxes of Pb and other trace metals due to sediment redox changes associated with the cessation of sewage discharge.

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enrichment is from the oxidation of upwardly diffusing dissolved Fe. The trace metal enrichment is coincident with the Fe enrichment, and is consistent with the porewater observations that the trace metals associate with Fe-oxyhydroxides. Organic carbon (Fig. 5) is also at a maximum in the surface sediments of MB, however the organic carbon enrichment extends deeper than the surface trace metal enrichment. Organic carbon likely complexes with the trace metals and influences the rates at which they are scavenged from porewater, but the differences in profile shape suggest that Fe is the most important phase controlling the trace metal distribution. There is no significant relationship between Cu and Fe in the MB porewaters, however the solid phase data suggest that Cu also associates with Fe-oxyhydroxides at MB. The absence of a porewater relationship at MB may be because Cu is delivered to the sediment as biogenic particles and the release of Cu from the particles obscures the relationship. To generate the narrow surface enrichment of trace metals, a mechanism must maintain Fe and trace metals at the surface despite bioturbation and sedimentation. When Fe-oxyhydroxides are reduced, adsorbed trace metals are released to the porewaters. At MB, readsorption to Fe-oxyhydroxides is more likely than at HB, due to the lower chance of encountering a sulfide ion (Fig. 9). The increased chance of re-associating with Fe-oxyhydroxides rather than sulfides, acts to progressively increase the concentration of metals associated with the Feoxyhydroxide phase. Successive reductions of Fe-oxyhydroxides due to burial or seasonal redox zone migration, will continue to focus trace metals upwards thus increasing the enrichment over time. If, on the other hand, the metals are precipitated as a sulfide, they will remain immobilized if conditions become more reducing, so there is no repeating process that can generate an enrichment. If this metal cycle were a steady process, and unaffected by benthic organism activity, the MB trace metal porewater profiles would have maxima with a steeper gradient towards the surface. Instead profiles are complex (Fig. 7) with sharp maxima and minima that sometimes make it difficult to establish an average flux direction. The activity of benthic organisms that complicate the porewater profiles also mix solid phase metals and may contribute to the enriched layer by supplying reduced metals from depth to the surface. This model of long-term metal behavior is related to that suggested by (Gobeil et al., 1997) to account for the offset of solid-phase Mn and Cd in continental margin sediments. These authors suggest that repeated migrations of the redox boundary on continental margins drive a ‘Mn pump’ upwards, and a ‘Cd pump’ downwards. In the case of MB however, the trace metals and Fe are pumped upward. A similar upward

5.5. Trace metal mobilization in the absence of detectable porewater sulfide Calculated trace metal fluxes at the offshore MB site were measured in three seasons within a single year and were more stable than the seasonal variation observed at HB, and comparable to non-winter HB fluxes (Table 3). The winter flux at HB is significantly higher than the MB fluxes due to the downward penetration of oxic conditions. However, the similar fluxes at MB and HB for the rest of the year are surprising considering the long history of contamination at HB, and result from different sediment chemistry at the two sites. HB is more sulfidic, so for most of the year dissolved trace metals in the surface sediments will be immobilized by either adsorption to Fe-oxyhydroxides or sulfide precipitation, depending on the season. At MB, there is significantly less sulfide in the surface sediments, so there are fewer sinks for dissolved metals in the sediments allowing porewater concentrations to become elevated and drive a diffusive flux (Fig. 9). Another surprising comparison is that the solid phase Pb concentration in the surface sediments of MB and HB is similar, due to the presence of the trace metal enriched surface layer at MB (Fig. 5). The surface enrichment and comparable concentrations are also clear when the Pb concentrations are normalized by Al to correct for variations in sediment type (Daskalakis and O'Connor, 1995; Bothner et al., 1998). The trace metal enrichment must be a diagenetic feature, rather than due to a contaminated layer introduced by relocation of the outfall, as the solid phase concentrations are the same as those measured before the outfall started operation. The thickness of the enriched solid phase Fe layer is comparable to (but deeper than) the ‘concave-up’ segment of the porewater Fe profiles. Coupling porewater data with the solid phase data is difficult because the porewater data are impacted on much shorter timescales than the solid phase data (Jahnke et al., 1982), however the porewater data support that the surface Fe

Dissolved Ag Ag-Fe Oxides

Ag2S

oxidation

precipitation

Porewater Concentration O2

adsorption

reduction of FeOOH

burial

Depth (cm)

Ferruginous

Depth

Oxic

Settling Flux

Fe Ag

burial Reaction Solid phase transport/ redox zone movement Diffusion of solute phase Dominant pathway

Fig. 9. Conceptual model of metal cycling in Massachusetts Bay (MB). The figure focuses on silver, but also applies to copper and lead. Each box represents a metal phase within the sediments for each redox zone. The arrows represent the mechanisms by which material can be transferred between boxes. Solid phase transport includes sedimentation and bioturbation. An idealized porewater profile is also provided. Oxygen was not measured in this study and the oxygen profile is inferred from the iron porewater profiles.

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pumping is observed for phosphate due to its strong association with Fe-oxyhydroxides in the sediments of Saguenay Fjord (Mucci et al., 2000). The downward Cd pumping observed by Gobeil et al. (1997) is due to the preferential downward diffusion of Cd and immobilization as a sulfide. The contrasting behavior at MB may be because the free sulfide concentration is lower, due to benthic organism activity, or because there is a greater abundance of Fe-oxyhydroxides at MB compared with the continental margin sediments. Further work is needed to more fully understand the balance of factors that generate a surface enrichment. This is an important consideration for management of contaminated sediment, as some locations may be particularly susceptible to enhanced metal contamination in the surface sediment due to diagenetic focusing of metals, rather than diminished metal contamination through sedimentation. 5.6. Benthic fluxes and Boston Harbor water quality Annual average diffusive fluxes calculated here (Table 3) can be compared with the benthic fluxes estimated from the coupled hydrodynamicwater quality model of Boston Harbor (Li et al., 2010). The HB data are considered appropriate to represent other fine-grained sediments in Boston Harbor as the metal concentrations are comparable at other sites (Zago et al., 2001; Lefkovitz et al., 2006) and the porewater profiles from the northern Harbor (Zago et al., 2001) are similar to those at HB. The fluxes estimated in this study are within an order of magnitude of the modeled estimates, but are consistently lower (Table 3). This suggests that diffusive flux from sediments significantly contributes to the persistent, high metal concentrations in the waters of Boston Harbor, but may not be the sole benthic source. Sediment resuspension can also contribute dissolved metals to the water column due to entrainment of porewater, and oxidation and desorption reactions of particles resuspended in the water column (Simpson et al., 2002; Kalnejais et al., 2010). Kalnejais et al. (2010) investigated the impact of resuspension at the HB site with erosion chamber experiments coupled to shear stress data from a hydrodynamic model, and found the loads from resuspension are variable between years and about the same magnitude as the diffusive release. The loads from resuspension when included, bring the measured benthic release within a factor of two of the lower end of the modeled values. A potentially significant benthic source that has not been investigated is irrigation of the sediments by benthic organisms. The impact of irrigation on metal fluxes is unclear (Morford et al., 2007) as the exchange of metals may be increased due to enhanced flushing, or may be reduced due to increased aeration in the sediments producing more Fe-oxyhydroxides that remove trace metals from solution. Further work is needed to include the effect of irrigation which may be an additional important source of benthic metals. Additionally, it is unknown how long the high winter fluxes persist at HB and therefore the calculated annual average is approximate and requires higher temporal resolution to improve the accuracy. 6. Conclusions The porewater dynamics of Ag, Pb and Cu at two contrasting coastal sites have been investigated to determine the mechanisms that drive the diffusive release of metals from sediments. The HB site is an embayment of Boston Harbor with a record of contamination. This site is characterized by high carbon oxidation rates, seasonally migrating redox zones and is strongly reducing with porewater sulfide present by 7 cm. The MB site is a less impacted, offshore site that is bioturbated with no dissolved sulfide in the upper 30 cm. The calculated diffusive flux of Cu and Ag at HB is high in winter and low in other seasons. In contrast, the MB fluxes are low and show no significant seasonal variation. The diffusive flux of Pb at both sites is usually very low throughout the year. Annual average diffusive fluxes of Cu from HB sediments make up a significant fraction of the benthic flux calculated by Li et al. (2010) to account for the persistent high metal concentrations in the waters of Boston Harbor.

The two sites illustrate the different pathways of metal cycling that are possible within the coastal zone. In strongly reducing sediments with measurable sulfide close to the sediment–water interface, trace metals are immobilized as sulfides, keeping surface porewater concentrations low and limiting diffusive fluxes. Conversely, at sites like HB that are subject to seasonal migration of the redox boundary, significant diffusive fluxes can develop when sulfides in the surface sediments are oxidized during cooler, more oxic conditions. At less reducing sites like MB where sulfide is not detected, diffusive fluxes are restricted primarily by adsorption of trace metals to Fe-oxyhydroxides. Because adsorption to Fe-oxyhydroxides does not compete with sulfide precipitation to remove trace metals from porewaters, the metals can become focused in the surface Fe-oxyhydroxide layer. This difference in cycling is particularly important to understand for the long-term management of contaminated sites. The consequences of contamination may be greater at less reducing sites as surface contamination will persist because the metals will be maintained close to the sediment–water interface by diagenetic processes rather than being buried. Acknowledgments We express our gratitude to J. Goudreau and the USGS Woods Hole sampling team for the expert core collection, M. Reuer for guidance on the Co–APDC method, and D. Schneider for assistance with the ICPMS. Insightful comments from Kevin Kroeger and anonymous reviewers greatly improved the manuscript. This work was funded by the Woods Hole Oceanographic Institution (WHOI) Sea Grant Program, under a grant from the National Oceanic and Atmospheric Administration, U.S. Department of Commerce, Grant No. NA16RG2273, project no. R/G28; the WHOI Coastal Ocean Institute; the National Science Foundation grant OCE-0220892 and the U.S. Geological Survey. LHK was funded by a Hackett Scholarship from the University of Western Australia, the WHOI Academic Programs Office and the University of New Hampshire. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2015.02.011. References Allen, H., Fu, G., Deng, B., 1993. Analysis of acid-volatile sulfide (AVS) and simultaneously extracted metals (SEM) for the estimation of potential toxicity in aquatic sediments. Environ. Toxicol. Chem. 12, 1441–1453. Aller, R., 1980. Diagenetic processes near the sediment–water interface of Long Island Sound. II Fe and Mn. Adv. Geophys. 22, 351–415. Andrews, D., Bennett, A., 1981. Measurements of diffusivity near the sediment–water interface with a fine-scale resistivity probe. Geochim. Cosmochim. Acta 45, 2169–2175. Audry, S., Blanc, G., Schäfer, J., Chaillou, G., Robert, S., 2006. Early diagenesis of trace metals (Cd, Cu, Co, Ni, U, Mo, and V) in the freshwater reaches of a macrotidal estuary. Geochim. Cosmochim. Acta 70 (9), 2264–2282. Audry, S., Blanc, G., Schäfer, J., Guérin, F., Masson, M., Robert, S., 2007. Budgets of Mn, Cd and Cu in the macrotidal Gironde estuary (SW France). Mar. Chem. 107 (4), 433–448. Benoit, J., Shull, D., Robinson, P., Ucran, L., 2006. Infaunal burrow densities and sediment monomethyl mercury distributions in Boston Harbor, Massachusetts. Mar. Chem. 102 (1), 124–133. Berner, R., 1984. Sedimentary pyrite formation: an update. Geochim. Cosmochim. Acta 48, 605–615. Birch, G., Chang, C.-H., Lee, J.-H., Churchill, L., 2013. The use of vintage surficial sediment data and sedimentary cores to determine past and future trends in estuarine metal contamination (Sydney estuary, Australia). Sci. Total Environ. 454, 542–561. Bothner, M., ten Brink, M.B., Manheim, F., 1998. Metal concentrations in surface sediments of Boston Harbor — changes with time. Mar. Environ. Res. 45 (2), 127–155. Bothner, M., Casso, M., Rendigs, R., Lamothe, P., 2002. The effect of the new Massachusetts Bay sewage outfall on the concentrations of metals and bacterial spores in nearby bottom and suspended sediments. Mar. Pollut. Bull. 44 (10), 1063–1070. Bothner, M., Casso, M., Rendigs, R., Lamothe, P., Baldwin, S., 2007. Using sediments to monitor environmental change in Massachusetts Bay and Boston Harbor. In: Bothner, M., Bradford, B. (Eds.), Processes Influencing the Transport and Fate of Contaminated Sediments in the Coastal Ocean—Boston Harbor and Massachusetts Bay. U.S. Geological Survey Circular 1302, pp. 48–56. Boulegue, J., 1983. Trace metal (Fe, Cu, Zn, Cd) in anoxic environments. In: Wong, C., Boyle, E., Bruland, K., Burton, J., Goldberg, E. (Eds.), Trace Metals in Seawater. Marine Science Series vol. 9. NATO, pp. 563–577.

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