Marine Chemistry 121 (2010) 145–156

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Marine Chemistry

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The distribution and speciation of dissolved trace metals in a shallowsubterranean estuary

Aaron J. Beck a,⁎, J. Kirk Cochran a, Sergio A. Sañudo-Wilhelmy b

a School of Marine and Atmospheric Sciences, Stony Brook University, Stony Brook, NY 11794-5000, USAb University of Southern California, Marine and Environmental Biology, Los Angeles, CA 90089-0371, USA

⁎ Corresponding author. Current address: VirginiaCollege of William & Mary, P.O. Box 1346, Gloucester Po

E-mail address: abeck@vims.edu (A.J. Beck).

0304-4203/$ – see front matter © 2010 Elsevier B.V. Aldoi:10.1016/j.marchem.2010.04.003

a b s t r a c t

a r t i c l e i n f o

Article history:Received 12 February 2010Received in revised form 9 April 2010Accepted 9 April 2010Available online 24 April 2010

Keywords:Subterranean estuaryTrace metalsSubmarine groundwater dischargeSGDGeochemistry

Geochemical cycles occurring at the interface between terrestrial and marine groundwaters, in the so-calledsubterranean estuary (STE), are not well understood for most elements. This is particularly true of thetransition metals, many of which have particular ecological relevance as micronutrients or toxicants. To gaina first approximation of trace metal geochemistry in the mixing zone, we examined the distribution of ninedissolved metals (Fe, Mn, Mo, V, Co, Ni, Cu, Pb, and Al) through a shallow STE in Great South Bay, New York,USA. We also performed a simple kinetic and chemical separation of labile and organic-complexed metalspecies in the STE. Dissolved Mn showed marked subsurface enrichment (up to 755 µM at 15 cm depth) thatwas suggestive of diagenetic remobilization. Dissolved Fe, however, was higher by more than three orders-of-magnitude in fresh groundwater (90 µM) as compared to marine groundwater (0.02 µM), and pH-mediated removal was evident as slightly acidic fresh groundwater (pH 6.8) mixed with marinegroundwater (pH ∼8.0). Dissolved Mo, Co, and Ni were primarily cycled with Mn, and highly elevatedconcentrations relative to bay surface waters (up to 300, 75, and 44 nM, respectively) were observed in theSTE. High levels of dissolved Pb (up to 4250 pM) observed in the fresh groundwater were nearlyquantitatively removed within the salinity mixing zone, in conjunction with marked reduction of dissolvedAl. Dissolved Cu exhibited non-conservative removal throughout the STE, and was correlated with the redoxpotential of the porewaters. Substantial percentages (N15%) of organic-metal species were only observed forCu and Ni, suggesting that these complexes were not generally very important for metal cycling in the STE.Kinetically labile species were observed for all metals examined except Cu and Pb, and represented anapproximately constant proportion (between 10% and 70%) of the total dissolved pool for each metal,indicating equilibrium between labile and non-labile species throughout the mixing zone. The non-conservative behavior observed for all metals examined in this study suggests that reactions occurring in theSTE are vastly important to the source/sink function of permeable sediments, and studies seeking to quantifySGD-derived trace metal fluxes must take into account biogeochemical processes occurring in thesubterranean estuary.

Institute of Marine Science,int, VA 23062, USA.

l rights reserved.

© 2010 Elsevier B.V. All rights reserved.

1. Introduction

The low organic content (b0.1 wt.%) of permeable sediments led tothe early presumption that such sedimentary environments arerelatively unreactive, and thus of limited importance for geochemicalcycling in the coastal ocean. A number of more recent studies haveshown that the low organic content is due to rapid remineralizationand flushing (e.g., Shum and Sundby, 1996; Huettel et al., 2003).Consequently, the role of permeable sediments in coastal biogeo-chemical cycling is being reevaluated, and all indications suggest thatthese sedimentary environments are vastly important to nearshore

chemical budgets (Shaw et al., 1998; Portnoy et al., 1998; Basu et al.,2001; Kim et al., 2005; Windom et al., 2006; Charette and Sholkovitz,2006).

At the same time, the importance of porewater advection throughsuch permeable sediments is also now recognized (Moore, 1996). Notonly does this submarine groundwater discharge (SGD; Burnett et al.,2003) represent a common and massive flux of water (although, notnecessarily fresh) to the coastal ocean, but there is also a very largeflux of associated chemical constituents (Moore, 1996, 1997; Shawet al., 1998; Basu et al., 2001; Charette et al., 2001; Jahnke et al., 2003;Charette and Buesseler, 2004). Although there is substantial literatureon nutrient fluxes due to SGD (Zimmermann et al., 1985; Lewis, 1987;Portnoy et al., 1998; Rutkowski et al., 1999; Charette et al., 2001;Jahnke et al., 2003; Ullman et al., 2003; Kim et al., 2005), less is knownabout the impact of SGD on trace metal fluxes and cycling in thecoastal ocean.

146 A.J. Beck et al. / Marine Chemistry 121 (2010) 145–156

The flux of metals across the sediment–water interface dependslargely on their geochemistry, and especially their behavior duringearly diagenesis (Shaw et al., 1990). Consequently, the subsurfacemixing zone of fresh groundwater and saline porewater, termed the“subterranean estuary” (Moore, 1999), is a highly dynamic zonewhere the composition of the advecting porewater is set beforedischarging into overlying marine waters (Charette and Sholkovitz,2006; Bone et al., 2006, 2007). Although some limited data areavailable for a handful of trace metals (Al, Cu, Co, Ni, Pb, and Ag; Becket al., 2007), detailed studies of metal cycling in the subterraneanestuary have been limited to Mn, Fe, Hg, and Mo (Charette andSholkovitz, 2002, 2006; Testa et al., 2002; Windom and Niencheski,2003; Duncan and Shaw, 2003; Snyder et al., 2004; Spiteri et al., 2006;

Fig. 1. Study site: Great South Bay, NY. The lowest panel shows an expanded view of the apptide water level.

Windom et al., 2006; Bone et al., 2007). Our limited knowledge oftrace metal behavior in this zone severely hinders our ability tounderstand and quantify the impact of SGD on metal cycling in thecoastal ocean.

Thus, the objective of the current study was to examine thedistribution and cycling of selected trace metals and their chemicalspecies in a subterranean estuary. A shallow (b1 m) subsurface zonehas been chosen as the region of interest, because the near sediment–water interface is the region where biogeochemical processes moststrongly affect the composition of the discharging porewater (Santoset al., 2009). For comparison with the profile through the STE mixingzone, we include a profile collected in a nearby fully saline subsurfacezone. The mechanisms regulating porewater flow differ between the

roximate locations of “STE” and “Saline” profiling locations. Both are well below the low

147A.J. Beck et al. / Marine Chemistry 121 (2010) 145–156

two sites, and the comparison is useful in examining geochemicalprocesses unique to the different environments.

1.1. Study site

The site chosen for this study is located in Great South Bay, NY, atthe end of Roe Avenue in the town of Patchogue (Fig. 1). Great SouthBay is a large, shallow lagoon on the south shore of Long Island, NY.Bottom sediments are predominantly permeable sandy glacialoutwash. These sediments are approximately 30 m in depth, andcomprise the Upper Glacial Aquifer. As this is the source of thegroundwater examined in this study, discussion of deeper aquifers isomitted here. The tidal range in the bay is small, less than 0.25 m. Thesalinity of the bay water is generally between 25 and 29, regulatedprimarily by freshwater input from groundwater and two majorrivers, the Connetquot and Carmans Rivers. These latter sources areapproximately equidistant from the Roe Avenue study site, 10 to15 km on either side.

The site was chosen because good documentation exists forgroundwater discharge patterns at this location for the past ∼25 years(Bokuniewicz, 1980; Bokuniewicz and Zeitlin, 1980; Bokuniewicz andPavilik, 1990; Seplow, 1991; Bokuniewicz et al., 2004), and a shallowdepth to fresh groundwater also has been reported (Seplow, 1991;Bokuniewicz et al., 2004). SGD flow rates at this site range betweenabout 1 and 11 cm d−1, with averages of about 2 to 4 cm d−1

(Bokuniewicz and Zeitlin, 1980; Bokuniewicz et al., 2004). Verticalhydraulic conductivities range between ∼2 and 25 m d−1, decliningbayward from shore, as do SGD flow rates (Bokuniewicz et al., 2004).Because Great South Bay is microtidal, no tidal modulation of SGD hasbeen observed at this site (Seplow, 1991; Bokuniewicz et al., 2004).Instead, discharge is driven solely by hydraulic head, and Darcy's Lawestimates of SGD match well with rates measured directly (Bokunie-wicz et al., 2004). Although prolonged rainy periods and intensestorms (e.g., tropical storms and hurricanes) can temporarily increaserates of discharge, return to normal conditions is rapid, on the order ofone week (Bokuniewicz and Zeitlin, 1980; Bokuniewicz et al., 2004).Thus, the SGD conditions at the time of sampling are expected to berepresentative of “normal” steady state conditions at this site.

The surficial subterranean estuary has been identified andreported often at this location (Seplow, 1991; Bokuniewicz et al.,2004), although not by that name. The mixing zone is unstable,essentially an inverse estuary, with fresh water underlying moresaline water (S=28). The shallow mixing zone is presumably

Fig. 2. Profiles of ancillary parameters in sediment porewaters. Solid and hollow symbols indby the gray area (after Charette and Sholkovitz, 2006). The missing point at 64 cm depth w

maintained by upward advection of fresh, terrestrial groundwaterand downward dispersion of salt water. The mechanism for the latterprocess is unknown (Rapaglia and Bokuniewicz, 2009), although saltfingering and wave stirring are the most likely possibilities (Boku-niewicz, 1992; Bokuniewicz et al., 2004).

The first sampling device (profile “STE”) was installed approxi-mately 30 m from the shore, in order to sample through thesubterranean estuary (Figs. 1 and 2). The second device (profile“Saline”) was placed farther offshore, about 100 m,wheremuch lowerflow rates have been measured (Bokuniewicz et al., 2004). Here, ratesof advection are low enough to allow dispersion or diffusion of salt todepths greater than 76 cm (the deepest sample in the profile).

2. Methods

2.1. Porewater collection and analysis

The sampling devices used in this study are described and tested indetail elsewhere (Beck, 2007). Briefly, each sampler was constructedof white PVC pipe, with a fiberglass tip. Porous plastic cylindersprotrude in semi-circles from the face of the PVC pipe and werespaced closely (∼2 cm between ports) in the top 30 cm of thesampler, and about three times farther apart below that depth. Thelength of each port was 2 cm, giving minimum sampling depthmidpoints of ∼4 cm intervals. Detailed discussion of sample overlap isprovided by Beck (2007), and at the porewater volumes extracted forthis study, it was probably minor. The porous ports were connected todedicated Teflon tubes by flexible C-flex tubing. The dead volume ofthe port with the longest tubing was less than 15 mL.

All components of the samplers were acid-washed separately forone week each in 3 N HNO3, 1 N HCl, and dilute (∼0.1 N) HCl. Aftereach acid treatment, samplers were rinsed 5 times with MilliQ water(MQ; 18.2 MΩ cm), including the ports and tubing. The entireapparatus was wrapped in several layers of heavy plastic sheetingfor transport. Both samplers were installed on 13 August 2006.Approximately 1 h after installing the devices at the field site, eachtube was purged of ∼30 mL of porewater using an acid-washedpolypropylene syringe. Both devices were then allowed to equilibratefor ∼36 h, with the free sample tube ends sealed within doublepolyethylene bags.

After the equilibration period, the sample tubes were eachattached to dedicated, acid-washed, all-polypropylene 60 mL syringeswith short sections of acid-cleaned C-flex tubing. Each depth was

icate STE and Saline profiles, respectively. The salinity transition zone (STZ) is indicatedas due to insufficient sample volume for probe measurement.

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purged of about 10–30 mL of porewater, and then samples werecollected. Approximately 45–60 mL was pulled into the syringes;nearly simultaneous sampling was achieved by cocking the syringesin “fill” position with sections of rigid Teflon tubing (Huettel, 1990;Beck, 2007). Sample collection rates were ∼50 mL min−1, well belowthe 100 mL min−1 maximum apparently allowable for obtaining goodtrace metal samples (Creasey and Flegal, 1999; Sañudo-Wilhelmyet al., 2002).

Each syringe was left attached to the sample tubing after beingfilled to prevent exposure to oxygen. Each syringe, in turn, wasdetached, and connected to an acid-washed 25 mm, 0.2 micron poresize, polypropylene syringe filter. Approximately 5 mL of sample waspurged through the filter to prevent oxidation artifacts (Bray et al.,1973; Lyons et al., 1979), and then 10–15 mL was collected in an acid-washed 30 mL LDPE bottle. The remaining sample was reserved forthe chemical separations described below.

2.2. Ancillary measurements

Salinity, pH, ORP (oxidation–reduction potential), and dissolvedO2 were determined in the field on the remaining porewater sampleusing a YSI 556 multi-probe. Where insufficient volume remained,additional sample was withdrawn with the syringe for probemeasurement. Chloride was analyzed as a check against theconductivity/salinity probe using a Copenhagen Radiometer chloridetitrator. Reproducibility for the chloride analyses was generally betterthan 2%.

2.3. Trace metals

Samples for total metals determination were returned to the cleanlab, acidified with 100 µL QHNO3, and allowed to sit for 2 months toensure that all metals were in solution. All metal samples, includingthe speciation fractions described below, were analyzed by ICPMS(ThermoFinnigan Element2) following 20-fold dilution. An indiumspike was used as an internal standard to correct for instrumentvariability during analysis. Standard addition techniques were used toanalyze samples that spanned the range of observed salinities andcorrect for matrix effects. Analysis of two standard referencematerials(NIST 1643d and SLRS-4) provided an accuracy check. Recoveries ofall elementswerewithin 10% of the reference value. Precision of metalanalyses were dependent on the analyte concentration, but weregenerally better than 5%.

2.4. Metal speciation

Metal speciation was operationally defined using a two-columnsolid-phase extraction technique (El Sayed and Aminot, 2000; Beckand Sañudo-Wilhelmy, 2007). Separations were carried out in thefield immediately following collection. Briefly, a first aliquot of 10–15 mL of sample was syringe-filtered as described above directlythrough a column of Chelex-100 resin (ammonium form). Thisprocedure isolates a fraction that is operationally defined for aparticular resin volume and sample flow rate (i.e., resin contact time);in the current study, a rapidly “exchangeable fraction” was targeted(Buckley et al., 1985; Gardner and van Veen, 2004; Donat et al.,1994; Shafer et al., 2004). Flow rate through the 1 mL resin bedwas maintained at a constant flow rate of ∼10 mL min−1. Theresin was immediately rinsed with ultraclean ammonium acetatebuffer (at pH ∼8), sealed, and double-bagged in polyethylene bags.

A second 10–15 mL aliquot of sample was then passed through asecond column of C-18 resin to extract hydrophobic organic-metalcomplexes. Although this operationally defined fraction certainly doesnot include all organic species (Dittmar et al., 2008), it allows goodcontamination control and a number of previous studies have found ituseful in characterizing surface water geochemistry (Hanson and

Quinn, 1983; El Sayed and Aminot, 2000; Zeri and Hatzianestis, 2005;Lemaire et al., 2006; Beck and Sañudo-Wilhelmy, 2007). Afterextraction, the resin was then rinsed with MilliQ water, capped, anddouble-bagged.

Columns were returned to the clean lab, where Chelex and C-18columns were eluted with several aliquots of 2 N QHNO3 and Optima-grade methanol, respectively. An elutant volume of at least 3 bedvolumes was used. To the methanol fraction, 100 µL of concentratedQHNO3 was added to oxidize the organic matter. Both eluate fractionswere then taken to dryness at room temperature under a HEPAfiltered flow hood. An additional 100 µL of QHNO3 was added to themethanol fraction to ensure complete destruction of the organics, andre-dried. Both eluate residues were then taken up in 2 mL of 1 NQHNO3, and allowed to sit for at least 1 month before analysis toensure that all metals were in solution.

3. Results and discussion

3.1. Ancillary parameters

Porewater depth profiles for water quality parameters (salinity,dissolved oxygen, pH, and oxidation–reduction potential) are shownin Fig. 2, geochemical carriers (Al, Fe, and Mn) in Fig. 3, and tracemetals (Co, Cu, Mo, Ni, Pb, and V) in Fig. 4. The subterranean estuaryconsists of the upper ∼30 cm of profile “STE,” with fully freshporewater salinities below. The “Saline” profile was completely saline(for Great South Bay, salinity 23–25) over the entire 75 cm depth.Dissolved oxygen showed penetration to approximately 10 cm inboth profiles, consistent with other sandy sediments (Falter andSansone, 2000). The surface oxygen peak observed in both profiles(Fig. 2) is probably due to production by benthic microalgae. Theapparent peaks at 37 cm depth in the STE profile and at 75 cm depthin both profiles may be a real feature due to horizontal advection ofoxygenated groundwater (Testa et al., 2002), although such focusedhorizontal advection is unlikely at these shallow depths. This is morelikely an artifact of briefly exposing the samples to air during transferto the probe measuring vessel. The porewater was consistently lessreducing in the STE profile, but both showed a marked decline inredox potential (ORP) to 15 cm depth. This decline coincided with apeak in pH, and may be reflecting consumption of acidity duringnitrate or metal oxide reduction by microbes. pH was consistentlyabout 1U lower in the STE compared to the saline profile.

3.2. Geochemical carrier elements

3.2.1. Mn, Fe, and AlTotal dissolved Mn concentrations were comparable between

profiles, and both showed marked peaks centered at ∼15 cm depth(Fig. 3). Little dissolved Mn was detected in the fresh groundwater,whereas concentrations remained elevated at depth in the salineprofile. The low concentrations of dissolved Mn at depth in the STEindicate a limited freshwater source of Mn to these shallowsediments. Instead, the coincident peaks of dissolved Mn at 15 cmdepth in both profiles indicate diagenetic remobilization by dissim-ilatory microbial reduction, likely driven by surface input of marineorganic matter (Santos et al., 2009; Kotwicki et al., 2005; D'Andreaet al., 2002).

Total dissolved Fe and Al showed similar patterns in the twoprofiles, being highly enriched at depth and decreasing surfaceward(Fig. 3). Both Fe and Al were higher in the STE profile, by 10-fold and4-fold, respectively. In the STE, a sharp decline in Fe and Alconcentration coincided exactly with the beginning of the mixingzone. Concentrations of these two elements were also very low in thesaline profile, indicating stronger control by solution chemistry thanmicrobial processes.

Fig. 3. Profiles of dissolved Mn, Fe, and Al (“geochemical carriers”). The upper row is from the STE profile, with the STZ indicated as in Fig. 2. The lower row is from the Saline profile.Total dissolved, Chelex-labile, and C-18 extractable metal fractions are indicated by solid circles, hollow circles, and solid triangles, respectively. Note different scales for metalconcentrations among STE and Saline profiles.

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The peak in dissolved Fe at 45 cm in the STE profile suggests thatmicrobial mobilization of Fe oxides may occur in the freshwateraquifer (Fig. 3). However, the marked removal of dissolved Fe inthe STE (Fig. 5) appears to be driven by Fe oxide precipitation aspH increased (Fig. 6). This is consistent with models showing thatincreasing pH in the STE greatly increases the rate of Fe oxidation andprecipitation, and that Fe removal in the STE is controlled by pHdynamics, not oxygen or redox potential (Spiteri et al., 2006).Although oxidation of dissolved Fe could be exploited by microbesfor energy gain, this does not appear to be the case for the STE (Rouxelet al., 2008). In the STE examined here, the threshold for Fe removaloccurs at a pH of approximately 7.5 (Fig. 6). This probably results information of a sediment layer that is greatly enriched in Fe oxides, asshown for other locations (Charette and Sholkovitz, 2002). However,Fe cycling does not appear to control the distributions of any of theother tracemetals examined in this study. It is likely that the relativelylow pH (7.4) in this layer maintains a positive surface charge on theprecipitated oxides and prevents sorption of other cations fromsolution. Given that the pH of the zero point of surface charge is muchlower for Mn oxides (b4) than Fe oxides (N7; Langmuir, 1997, andreferences therein), cycling of other metal cations in the STE isprobably regulated more strongly by the “manganese curtain” thanthe “iron curtain” (as noted for Ba, Charette et al., 2005; Charette andSholkovitz, 2002). The converse would be expected for dissolvedelements occurring primarily as oxyanion species, such as arsenic andphosphate (Charette et al., 2005; Bone et al., 2006).

The distribution of dissolved Al in the STE suggests a pH-mediatedremoval process similar to Fe (Fig. 5). Removal is consistent withprevious reports of Al in groundwater (Kjoller et al., 2004) and inanother subterranean estuary (Beck et al., 2007). However, Alconcentrations in the saline profile were comparable to those in thefresh groundwater, despite pH values that were at least 1Umore basic(Figs. 2 and 3). Thus, the observed Al removal probably results fromcoagulation and sorption to aquifer solids as the fresh groundwaterenters the STE mixing zone (Eckert and Sholkovitz, 1976; Hydes andLiss, 1977; Sholkovitz, 1978; Zhang et al., 1999).

3.2.2. SulfurAlthough sulfate–sulfide cycling is often a major control on

transition metals cycling in sediments, it does not seem especiallyimportant at this site. These porewater samples lacked any smell ofhydrogen sulfide, and other porewater samples we have collected inGSB (Beck et al., 2008) also lacked noticeable indication of sulfides.Hines and Buck (1982) showed that the distribution of sulfate-reducing bacteria in the subterranean estuary fluctuates with sulfateconcentrations. Consequently, we would not expect sulfate reductionto occur in the freshening zone of the STE. It also seems unlikely thatsignificant sulfate reduction would occur at this site, given the highredox potential of the bulk porewater (Fig. 2). In the current profiles,sulfide production in microsites of otherwise non-anoxic sediments ispossible, with subsequent diffusion into the bulk porewater (de Beeret al., 2005). This sulfide would be expected to react rapidly and

Fig. 4. Dissolved trace metal profiles. Symbols as in Fig. 3.

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Fig. 5. Salinity distribution of metals in the STE (solid symbols). Metal concentrations from the Saline profile are included for comparison (hollow symbols).

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precipitate as metal sulfides with Fe, Cu, or Mo. Because this wouldeffectively titrate these metals from solution, their relatively highconcentrations in the saline profile suggest limited production ofsulfides.

3.3. Dissolved trace metals

Trace metal concentrations in the porewaters were generallycomparable between the two profiles (Fig. 4), except for Mo (2-fold

Fig. 6. Covariation of dissolved Fe with pH. Solid symbols indicate the STE profile, andhollow symbols indicate the Saline profile.

higher in the saline profile), Co (2-fold higher in the STE profile), andPb (10-fold higher in the STE profile). The distribution of dissolvedmetals in the subterranean estuary showed three distinct trends:elevated concentrations in fresh groundwater (Fe, Al, and Pb), mid-salinity minimum (Cu and V), or mid-salinity maximum (Mo, Mn, Co,and Ni). The behavior of these metals in the saline profile provides acomparison for distinguishing between estuarine and diageneticprocesses.

3.3.1. Mo, Co, and NiDissolution and precipitation of Mn oxides in turn appears to drive

the trends observed for Mo, Co, and Ni. Mn oxides tend to have a netnegative charge at circumneutral solution pH, providing a highlyreactive surface for adsorption of othermetal cations. Co-precipitationof othermetals duringMn oxidative precipitationmay also cause theirremoval from solution. The observation that these metals co-variedstrongly but not exactly with Mn in both profiles (Fig. 7) suggestsdifferences in their sources or their rate and degree of mobilizationduring Mn cycling. This may reflect the relative dominance of surfaceadsorption versus co-precipitation, as well as stabilization bydissolved ligands (e.g., organic matter, see below). Dissolved Coshowed the strongest correlation with Mn (Fig. 7a), consistent withobservations elsewhere (Beck et al., 2007). There is a small freshwatersource of Co to the STE, but trapping by Mn oxides and release duringoxide dissolution causes the enrichment observed in the upper 20–30 cm of both STE and saline profiles (Fig. 4). These high peakconcentrations of dissolved Co appear to be nearly quantitativelyremoved during re-oxidation of dissolvedMn at the oxic near-surface.

Fig. 7. Covariation of dissolved Mo, Co, and Ni with Mn in the porewater. Symbols as inFig. 6.

Fig. 8. Covariation of dissolved Cu with redox potential. Symbols as in Fig. 6. Regressionlines, equations, and coefficients are shown.

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Dissolved Mo shows similar control by Mn, although with anapparent marine source. Mo is highly enriched in porewater at theshallow side of the Mn peak in the STE (Figs. 3 and 4), and there doesnot appear to be a major freshwater source of Mo. It appears thatseawater recirculation (i.e., during tidal or wave pumping) throughthese sands results in Mo sequestration on, and cycling with,sedimentary Mn oxides. However, as with Co, Mo removal fromsolution near the sediment–water interface indicates that trapping onthese oxides prevents discharge to the overlying water. Thus, incontrast to previous studies (c.f. Windom and Niencheski, 2003), SGDat this site does not seem to be a definitive sink for Mo (as for nearbyJamaica Bay; Beck et al., 2009) or a source (as for the Chao PhrayaRiver estuary; Dalai et al., 2005). It is likely that seasonal controls onMn cycling (Aller, 1994) result in non-steady state conditions onmonthly time scales for Mo flux via SGD. In locations where sulfidesare minimal, variations in Mn mineral precipitation and dissolutionmay regulate the SGD source–sink function for Mo.

In contrast to Mo, fresh groundwater appeared to be the primarysource of Ni to the subterranean estuary. Dissolved Ni concentrationsin fresh groundwater were nearly double those in the overlying water(24 versus 13 nM, respectively). Because of the deep Ni source, Niappears to be adsorbed onto and remobilized with Mn at the deepestside of the Mn peak (Figs. 3 and 4). Ni in solution at salinities higherthan ∼10 appears to be stable and resistant to subsequent removal in

the STE as was observed for Co and Mo. Instead, dissolved Ni mixesconservatively through the STE (r2=0.97 for correlation withsalinity), with an extrapolated zero-salinity “effective endmember”of 51.5 nM. Thus, trapping on Mn oxides at depth in the STE results inenrichment of the groundwater Ni endmember.

3.3.2. PbConsistent with previous reports, Pb loss in the STE coincided with

Al removal from solution (Figs. 3–5; Beck et al., 2007). The correlationof these two elements in the STE profile (not shown, r2=0.67,pb0.001) indicates that removal of the Al carrier phase drives Pbremoval. In locations such as rivers where fresh and saline ground-waters do not mix before discharging to the overlying water column,such Pb removal would not occur. This is consistent with the elevatedlevels of dissolved Pb measured in rivers in Great South Bay (Clarket al., 2006) where groundwater discharge also appears to be highest(Beck et al., 2008). Due to removal processes in the STE, Fe, Al, and Pbflux via SGD would all be expected to be strongly reduced relative tosites without the mixing zone.

3.3.3. Cu and VDissolved Cu and V both showed elevated levels in fresh

groundwater (Fig. 4; 17 and 149 nM, respectively), and were highin surface waters as well (30 and 65 nM, respectively). The mid-depthminima observed for both elements occurred at different depths, withdissolved Cu showing removal at lower salinities (b10) than dissolvedV (salinity 10–20; Fig. 5).

Cu variation in the STE profile is matched by similar pattern in thesaline profile, suggesting that Cu is largely unaffected by solutionchemistry changes associated with water mixing in the STE. Instead,dissolved Cu in both profiles co-varied with oxidation–reductionpotential in the porewater, showing that Cu was strongly removedunder reducing conditions (Fig. 8). It is unlikely that Cu was beingremoved from solution by formation of insoluble sulfides as most ofthe Cu removal was in the presumably sulfate-poor freshwater regionin the shallow aquifer. In fact, above salinity 10, Cu mixedconservatively in the STE (r2=0.99; Fig. 5). The data indicate aneffective zero-salinity endmember of −12.1 nM, suggesting that Cuwas removed from both water column and fresh groundwater andsequestered in these sandy sediments. The exact mechanism of Curemoval is not clear. While it is unlikely that metallic Cu would beproduced in the current setting, redox potentials in the porewater aresufficiently low that reduction of Cu(II) to Cu(0) is a possibility (Borchet al., 2010). Dissolved Cu removal at depth coincides with the disap-pearance of dissolved Fe (Fig. 5), and may be a result of Cu mineralprecipitation (e.g., cuprite) following Cu reduction by dissolved Fe(II)(Matocha et al., 2005).

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The mid-depth minimum of dissolved V (35 nM) in the STE isdifficult to explain; it coincided exactly with the peak of dissolvedMo,although it is not obvious why this might occur. Dissolved V did notco-vary in either profile with any of the other measured parameters.Levels of V were elevated in both fresh groundwater (149 nM) andsaline porewater (up to 137 nM), the former indicating that SGD maybe a major source of this element to overlying waters. Indeed,dissolved V in the Great South Bay water column (65 nM) is morethan double that predicted by conservative mixing with ocean water(∼30 nM). Dissolved V levels in the Connetquot and Carmans Riversare too low (∼10 nM; Beck and Sañudo-Wilhelmy, unpublished) toaccount for the excess. These levels in the fresh groundwater aresurprisingly high, given that dissolved V concentrations reported forrivers are generally far lower than seawater, and tend to be lowest inlow-pH water (Schiller and Boyle, 1987). Elevated V in thesegroundwaters may reflect contamination by fossil fuels (Hope,2008). This would also explain the co-enrichment of dissolved Ni infresh groundwater at this site, as fossil fuel materials are highlyenriched in both Ni and V relative to average crustal composition(Sholkovitz et al., 2009, and references therein). If the V source is notartificial, SGDmay be amajor component of the oceanic V budget, anddissolved V may even find application as an SGD tracer.

Fig. 9. Relationship between Chelex-labile metal fraction and total dissolved pool. Regressionare shownwhen differences between the two sample profiles are negligible. For Mo, the threan almost identical position). In the lower two panels, lines are shown for selected proport

3.4. Speciation of dissolved trace metals

3.4.1. Chelex-labile trace metalsA substantial fraction of Chelex-labile (hereafter, “labile”) metal

species was observed for all metals except Cu and Pb (Figs. 3 and 4).This is unsurprising given the tendency of these twometals for strongparticle interactions, and that complexation by strong ligands canreduce their sorption to surfaces (Schindler and Stumm, 1987). Thelabile fraction of other metals represented as much as 70% of the totaldissolved pool (Fig. 9). A surprisingly constant ratio was observedbetween labile and total dissolved metals, and tended to be the samefor both STE and saline profiles. As proposed by van den Berg (1993),the constant ratio suggests equilibrium between labile and non-labilespecies in the porewater, and further indicates that reactionsregulating the partitioning occur rapidly relative to advection andmixing in the interstices. At the current site, assuming a flow rate of∼10 cm d−1 (Bokuniewicz and Zeitlin, 1980; Bokuniewicz et al.,2004) and porosity of ∼0.5, and given the sharp chemical gradientsobserved, this suggests that chemical transformations occur on a scaleof several hours or less. The poor correlation between labile and totalNi is consistent with its slower reaction kinetics (Morel and Hering,1993). The rapid kinetics indicated by the high degree of correlation

lines, equations, and coefficients are shown for Mn, Fe, Mo, and Co. Single best-fit linese outlying points are separated from the other data (note that two of these points plot ations. Symbols as in Fig. 6.

154 A.J. Beck et al. / Marine Chemistry 121 (2010) 145–156

observed for the two Mn pools (Fig. 9) further confirms microbialcontrol on Mn cycling in these sediments, as abiotic reaction kineticstend to be very slow for Mn (e.g., Diem and Stumm, 1983; Balzer,1982).

No difference was found in labile species proportion between STEand saline profiles formost elements. Dissolved V,Mo, and possibly Coare exceptions to this (Fig. 9). Though there was substantial scatter inthe data, dissolved V clustered into low- and high-percent labilegroups (∼10% and ∼50% labile, respectively). Brackish and highsalinity samples tended to have a greater proportion of labile V, whilefresh groundwater contained the less-labile V (Fig. 9). Again we findno obvious explanation for this pattern, but the apparent non-labilenature of V in fresh groundwaters is consistent with the observed highconcentrations and potential for transport. Dissolved Co had a slightlyhigher percentage of labile species in the STE as compared to thesaline profile, though, given the scatter in the data, the difference maynot be particularly significant. Most of the labile:total Mo ratios matchbetween the two profiles, but three points stand out at ∼40% versus∼20% labile for the other samples (Fig. 9). These three samples mayindicate local production of sulfides, as Chelex-labile Mo is greater insulfidic porewaters, reaching up to 60% of the total dissolved pool(Malcom, 1985). Dissolved Fe was very low at these depths (compareFigs. 3 and 4), so presence of some sulfide is not an unreasonablepossibility.

3.4.2. C-18 extractable metalsMetal complexes extractible with C-18 resin were generally

undetectable for all metals except Ni (0.3–25 nM, or 1–72% of thetotal dissolved pool), Cu (1–14 nM, 10–50%), and V (1–17 nM, 2–13%)(Fig. 4). This is consistent with the common association of Ni and Cuwith organic matter in marine systems. The organic species for thesetwo metals in the current profiles tended to show highest concentra-tions at the sediment–water interface, and decrease with depth,suggesting that the organic ligands had a marine source or wereproduced by benthic microalgae in surface sands. Dissolved V was theonly metal for which a small but consistent organic fraction wasmeasured in deep porewaters of both STE and saline profiles.Stabilization by organic ligands may explain the lower lability of Vobserved for these samples.

4. Conclusions

It is evident that most of the tracemetal distributions in permeablesediment porewater at this site are controlled by cycling of thegeochemical carriers Mn and Al, but not Fe. Mn played a particularlysignificant role, presumably due to the greater sorptive capacity of itssolid oxide surfaces for transition metal cations. Mn cycling in thepermeable sediments and STE is likely controlled primarily bydissimilatory microbial metabolism. The role of organic matter inthis process is implicit, and dissolved Mn profiles in the current studyindicate greater input of organic matter reductant in the upper 25 cmthan at depth in the sands (i.e., not from fresh groundwater; Santoset al., 2009).

Cycling of dissolved Co, Ni, andMowithMn apparently dominatedthe effect of mixing on their distribution in the STE. The correlation ofCu with redox potential again indicated a secondary effect of changesin water chemistry most likely associated with microbial activity.In contrast, Fe, Al, and Pb all showed marked removal from freshgroundwater upon mixing in the STE.

It is noteworthy that all of the interpreted processes would act toaccumulate the metals in surface sands. Although we cannot identifyif there is temporal variability in the source–sink function of per-meable sediments at this location, the dissolved porewater concen-trations observed in this studywere similar or greater in magnitude tolevels reported for fine-grained sediments. Some metals, such as Mn,Mo, and Pb, were many orders-of-magnitude higher in these pore-

waters than in the overlying water column, and begin to approachlevels at which they are no longer “trace” (e.g., millimolar levels ofMn). This highlights the potential importance of permeable sedimentsand advective transport to mass budgets of trace metals in coastalmarine waters.

A major unknown remains regarding the reservoir of these metalspresent in the sands. How long can the concentrations we observed bemaintained in porewater if there is loss by SGD to overlying waters?What is the ultimate source of the metals; is it seawater, as suggestedfor Mo, or is there a fresh groundwater source, as for Fe, Pb, and V? InWaquoit Bay, MA, Charette et al. (2005) estimated that solid-phase Fecontent in the STE would require decadal to centennial time scales tobe formed by the fresh groundwater source. It will be important infuture endeavors to determine the time scale characteristic of othertrace metal mass balances in sands.

The importance of SGD and metals fluxes from sandy sediments ascomponents of regional and global material budgets dependsfundamentally on the metal source to the sands (i.e., groundwater,or seawater recirculation) and the material reservoir containedtherein. For instance, our data indicated Cu removal from the watercolumn, and SGD may play a primary role in Cu removal in GreatSouth Bay. Seawater recirculation apparently also resulted in Moenrichment in the sands, but cycling with Mn maintains highporewater concentrations near the sediment–water interface. If thetrapping efficiency for Mn decreases (such as during water columnhypoxia), then SGDmay represent a temporary, but major, Mo sourceto the water column. Conversely, V appears to have a fresh ground-water source to the STE, and with concentrations 5-fold higher thanthe ocean, SGDmay be a dominant source of V to Great South Bay andNY coastal waters. More work is necessary to better constrainbiogeochemical controls on metal cycling in the STE, and thequantitative role of the solid phase in regulating the source–sinkfunction of SGD. This knowledge will be critical to comprehensiveassessment of material budgets in coastal marine waters.

Acknowledgements

Funding for this project was provided by New York Sea Grant andthe Long Island Groundwater Research Institute. We greatly appre-ciate sampling assistance by J.P. Rapaglia, C. Panzeca, and S. Yang. Themanuscript benefitted from the constructive comments of twoanonymous reviewers and the associate editor. This is ContributionNumber 1393 from the School of Marine and Atmospheric Sciences,Stony Brook University, New York.

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