Chemical Geology 351 (2013) 76–94

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Chemical Geology

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Controls on tungsten concentrations in groundwater flow systems:The role of adsorption, aquifer sediment Fe(III) oxide/oxyhydroxidecontent, and thiotungstate formation

Karen H. Johannesson a,⁎, Heeral B. Dave b, T. Jade Mohajerin a, Saugata Datta c

a Department of Earth and Environmental Sciences, Tulane University, New Orleans, LA 70118-5698, USAb Department of Earth and Environmental Sciences, The University of Texas at Arlington, Arlington, TX 76019-0049, USAc Department of Geology, Kansas State University, Manhattan, KS 66506-3201, USA

⁎ Corresponding author. Tel.: +1 504 862 3193.E-mail address: kjohanne@tulane.edu (K.H. Johanne

0009-2541/$ – see front matter © 2013 Published by Elhttp://dx.doi.org/10.1016/j.chemgeo.2013.05.002

a b s t r a c t

a r t i c l e i n f o

Article history:Received 23 January 2013Received in revised form 30 April 2013Accepted 4 May 2013Available online 15 May 2013

Editor: Carla M. Koretsky

Keywords:TungstenGroundwaterSurface complexationThiotungstateSolution complexationGeochemical modeling

Groundwater samples were collected along flow paths within the Carrizo Sand aquifer (southeastern Texas)and the Aquia aquifer (coastal Maryland) for analysis of dissolved tungsten (W) concentrations, alongwith the major solutes, pH, and measures of in situ redox conditions [e.g., dissolved Fe(II), Fe(III), andS(-II) concentrations]. In addition, sediment samples were collected from both aquifers to evaluate thesolid-phase speciation of W. Tungsten concentrations in the Carrizo Sand aquifer range from 3.64 to1297 pmol kg−1 (mean ± SD = 248 ± 440 pmol kg−1), with the lowest concentrations reported from therecharge area. Tungsten concentrations progressively increase down gradient along the flow path withinCarrizo Sand aquifer groundwaters, reaching the highest levels in sulfidic groundwater roughly 50–60 kmfrom the recharge zone. Tungsten is strongly correlated with pH and dissolved sulfide [i.e., S(-II)] concentra-tions in Carrizo groundwaters (r = 0.78 and 0.95, respectively). In Aquia aquifer groundwaters W concentra-tions range between 14.3 and 184.4 pmol kg−1 (mean ± SD = 63.9 ± 59 pmol kg−1), and exhibit no, or atbest, weak relationships with other geochemical parameters measured along the flow path (e.g., r = 0.4 and0.08 for W vs. pH and S(-II), respectively). The data suggest that the increase in W concentrations in Carrizogroundwaters reflects, in part, pH-related adsorption/desorption, which has been shown to be substantialfor groundwaters and lake waters with pH > 8. Owing to the similar chemical properties of W and Mo,which form thiomolybdates in sulfidic waters, the formation of thiotungstate complexes may also be impor-tant in the sulfidic waters of the Carrizo aquifer. The substantially lower W concentrations in Aquia aquifergroundwaters reflect the fact that the maximum groundwater pH never exceeds 8.4 (mean ± SD pH =7.97 ± 0.23), dissolved low S(-II) concentrations remain low (mean ± SD S(-II) =0.29 ± 0.16 μmol kg−1),and that Aquia aquifer sediments have substantially higher Fe(III) oxide/oxyhydroxide contents compared tothe Carrizo Sand aquifer. Our data indicate that pH-related adsorption/desorption reactions and the Fe(III)oxide/oxyhydroxide content of aquifer sediments are key controls affecting W concentrations in oxic, suboxic,and anoxic groundwaters, and, further, that the formation of thiotungstate species may also be important insome anoxic (i.e., sulfidic) waters.

© 2013 Published by Elsevier B.V.

1. Introduction

A number of recently recognized clusters of childhood leukemia,including clusters in Fallon, Nevada, Sierra Vista, Arizona, and ElkGrove, California, are tentatively linked to the proximity of thesesites to tungsten (W)-bearing ore deposits, associated active or inac-tive smelting operations, and/or hard-metal processing facilities

sson).

sevier B.V.

(Seiler et al., 2005; Koutsospyros et al., 2006; Sheppard et al., 2006,2007; Bednar et al., 2007, 2008). Indeed, the childhood leukemia clus-ter afflicting Fallon, Nevada, spawned an investigation by the UnitedStates Centers for Disease Control and Prevention (CDC), which re-vealed through urine analysis that residents were exposed to elevatedlevels of W (Seiler et al., 2005; Koutsospyros et al., 2006). Althoughit is inherently difficult to directly link these childhood leukemia clus-ters to environmental exposure to elevated W concentrations, eithervia inhalation or through consumption of drinking waters with highW concentrations, many studies have shown that W can be toxic andmay indeed be carcinogenic (Peão et al., 1993; Marquet et al., 1996;ATSDR, 2005; Kalinich et al., 2005; U.S. EPA, 2008; Kelly et al., 2013).

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Moreover, inhalation of W dust in the hard metal manufacturing in-dustry is linked to pulmonary fibrosis and hardmetal pneumoconiosis(Edel et al., 1990; Sprice et al., 1994; Sahle et al., 1996). Furthermore,polytungstates like sodium metatungstate (3Na2WO4·9WO3) appearto be more toxic than monomeric tungstate (e.g., Na2WO4; Strigulet al., 2009, 2010), and may thus pose problems for people consumingdrinkingwaterwith elevatedW concentrations. The CDC investigationled to a study by the U.S. Geological Survey of W in groundwatersof the Carson Desert region of northwest Nevada, which showedelevated W concentrations ranging from 1.47 to 4036 nmol kg−1

(0.27–742 μg kg−1; Seiler et al., 2005). These concentrations arehigh and similar to W concentrations previously reported for ground-waters of the Carson Desert (1029 nmol kg−1; Johannesson et al.,2000), and may in part be responsible for the high levels of W foundin local residents.

In addition to natural sources of W to the environment, interestin the environmental geochemistry of W is also on the rise owingto its increasing use as a replacement for lead (Pb) ammunitionfor hunting and recreational shooting, for fishing weights, and formilitary ammunition (i.e., high kinetic energy penetrators and smallarms ammunition; Strigul et al., 2005; Koutsospyros et al., 2006;Bednar et al., 2007; Clausen et al., 2007; Bednar et al., 2008; Clausenand Korte, 2009). Here, the use of W was originally conceived as away to limit the addition of toxic Pb to the environment by whatwas thought to be a non-toxic, inert metal of low environmental mo-bility (Koutsospyros et al., 2006; Bednar et al., 2007). In fact, the U.S.Army's “GreenArmament Technology” (GAT) program initially soughtto limit environmental pollution by recommending the replacementof Pb-based ammunition with W-based rounds (Koutsospyros et al.,2006). Nevertheless, despite the initial beliefs that W was a non-toxic,chemically inert heavy metal, laboratory studies clearly reveal thatW can be toxic and carcinogenic (e.g., Marquet et al., 1996; Sahleet al., 1996; Kalinich et al., 2005; Steinburg et al., 2007; Strigul et al.,2009; Strigul, 2010; Strigul et al., 2010; Kelly et al., 2013), andfield-based investigations indicate that it is mobile in the environ-ment (e.g., Johannesson et al., 2000; Dermatas et al., 2004; Petrunicand Al, 2005; Seiler et al., 2005; Dermatas et al., 2006; Arnórssonand Óskarsson, 2007; Clausen et al., 2007; Dave and Johannesson,2008a,b; Bednar et al., 2009; Clausen and Korte, 2009; Johannessonand Tang, 2009).

In general, W exhibits similar geochemical behavior to otheroxyanion-forming trace elements such as molybdenum (Mo) and ar-senic (As) in the environment in that it is strongly sorbed to Fe(III)oxides/oxyhydroxides and other mineral surfaces in oxic watersof low to circumneutral pH, and desorbs from these mineral surfacesites at higher pH (Johannesson et al., 2000; Gustafsson, 2003;Koschinsky and Hein, 2003; Seiler et al., 2005; Arnórsson andÓskarsson, 2007; Bednar et al., 2007, 2008, 2009; Johannesson andTang, 2009; Kashiwabara et al., 2013). At low pH (pH b 5) and highdissolved W concentrations (W ≥ 10−2 mol kg−1), W tends toform polymerized species in solution, whereas monomers such asthe tungstate oxyanion, i.e., WO4

2−, predominate at lowW concentra-tions, neutral to alkaline pH, high temperatures, and high ionicstrength (Wesolowski et al., 1984; Wood, 1992; Bednar et al., 2007;Rodríguez-Fortea et al., 2008). Despite these observations, little is ac-tually known about the biogeochemistry of W in low-temperature,groundwater flow systems, which are important drinking watersources for much of the world's population (Solley et al., 1993;Clarke and King, 2004). To the best of our knowledge, there areno published studies that have investigated how biogeochemicalreactions occurring along groundwater flow paths, including geo-chemical reactions that alter the solution composition and redoxconditions, affect W concentrations and speciation within aquifersystems. For example, although WO4

2− is thought to predominatein natural waters, the similarity between W and Mo suggeststhat thiotungstate species may form in some anoxic, sulfidic waters

(e.g., Erickson and Helz, 2000; Dave and Johannesson, 2008a,b;Mohajerin and Johannesson, 2012).

In this contribution we present W concentrations for groundwaterscollected along flow paths in two well-characterized aquifers, theCarrizo Sand (Texas) and the Aquia (Maryland) aquifers. GroundwaterW concentrations are evaluated along the studied flow paths in eachaquifer along with major solute concentrations, pH, and a numberof redox sensitive indicators (i.e., Eh, dissolved oxygen, iron species,and sulfide) to develop a better understanding of the biogeochemicalprocesses that control W concentrations in groundwater flow systemsthat have not been contaminated by anthropogenic activities.

2. Study sites

2.1. Carrizo Sand aquifer

The portion of the Carrizo Sand aquifer studied is located inAtascosa and McMullin Counties in southeastern Texas (Fig. 1).Water abstracted from the aquifer is used for both irrigation andwater supply (i.e., drinking water) purposes in the study region(e.g., Hamlin, 1988). Within the study area, the Carrizo Sand aquiferis exposed and recharged by precipitation in northwest AtascosaCounty, where it defines the Carrizo Sand ridge (Fig. 1). From the re-charge area groundwater flows southeast towards the Gulf of Mexicodown the structural dip of the Carrizo Sand (Pearson andWhite, 1967;Castro and Goblet, 2003; Dutton et al., 2006). In southern Atascosa andnorthern McMullen Counties, wells finished within the Carrizo Sandaquifer are on the order of 1300 to 1400 m deep. These wells producerelatively warm, artesian groundwaters (i.e., 35°–55 °C; Table 1),consistent with heating due to a typical, nonorogenic, regional geo-thermal gradient (i.e., 15–23 °C km−1; Philpotts, 1990; Haque andJohannesson, 2006). In down-gradient portions of the Carrizo Sandaquifer, and beyond our study area, high fluid pressure within the“geopressured” zone causes groundwater to discharge from the aqui-fer by upward, cross-formational leakage (Castro and Goblet, 2003;Castro, 2004; Dutton et al., 2006). Discharge is also thought to occurin southern McMullen County by upward infiltration along fracturesassociated with the Wilcox geothermal corridor (Castro et al., 2000;Castro and Goblet, 2003; Castro, 2004).

The Carrizo Sand is composed of medium-grained, unconsolidatedEocene quartz sand with minor amounts of clay minerals, calcite,lignite, and pyrite (Pearson and White, 1967). Ferric iron oxides/oxyhydroxides are ubiquitous within the Carrizo Sand aquifer occur-ring principally as coatings on the aquifer mineral (i.e., sand) grains(Hamlin, 1988; Basu et al., 2007). The porosity of the Carrizo Sandaquifer is estimated to range between ~30 and 40%, and the meanhydraulic conductivity is ~8.2 m day−1 (Pearson and White, 1967).The thickness of the Carrizo Sand varies substantially throughout thestudy region. In the outcrop area in northwestern Atascosa Countythe thickness is estimated at 100 m, whereas in southern McMullenCounty the estimated thickness is ~360 m (Castro et al., 2000; Castroand Goblet, 2003).

The 14C age of groundwater in the Carrizo Sand aquifer increasesfrom ~5000 14C years near the recharge area in northern AtascosaCounty to roughly 30,000 14C years in southern McMullen County(Fig. 1; Pearson and White, 1967; Stute et al., 1992). Based on the14C age of the Carrizo Sand groundwater, Pearson and White (1967)estimated the flow velocity of Carrizo Sand groundwaters to be2.4 m a−1 at approximately 16 km down gradient along the flowpath, and 1.6 m a−1 at roughly 50 km down gradient. Stute et al.(1992) determined a similar average groundwater flow velocity of~2 m a−1 along the same flow path. Carrizo Sand groundwaters arepredominantly fresh and potable, except towards the southeast insouthern McMullen County where salinity increases reaching valuesof 10 g kg−1 (Castro et al., 2000; Castro and Goblet, 2003, 2005).The current study is limited to the portion of the aquifer in Atascosa

Fig. 1.Map of the Carrizo Sand aquifer study site within southeastern Texas, USA. The aquifer is recharged where it outcrops in northern Atascosa County (dashed orange lines), andgroundwater flows towards the southeast (i.e., large white arrow). All subsequent figures showing geochemical parameters plotted as a function of distance along the flow path areprojected onto this general flow path line. Wells F-1, BCW-5, and KS-1 are located in the unconfined, recharge zone of the aquifer. All other wells sampled in the study (i.e., wellsN-3, G-1, Poteet, PCW-1, Peeler-1, AC74R-1, and SMA74R-2) are from the confined portion of the aquifer. (For interpretation of the references to color, the reader is referred to theweb version of the article.) Map is modified from Johannesson and Neumann (2013), and is based on a Google Earth image.

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and northern McMullen Counties containing potable groundwater(Fig. 1).

2.2. Aquia aquifer

The Aquia aquifer is an unconsolidated, principally confined CoastalPlain aquifer that outcrops along a southwest to northeast trend from

near Washington, DC, through Annapolis, Maryland, and across theChesapeake Bay into Kent County, Maryland (Fig. 2; Chapelle, 1983;Chapelle and Knobel, 1983, 1985; Aeschbach-Hertig et al., 2002). TheAquia aquifers dip to the southeast, beneath Chesapeake Bay, endingin a facies change to progressivelymore clay-rich, and thus, less perme-able sediments roughly 90 km from the outcrop area (Chapelle, 1983;Drummond, 2001; Aeschbach-Hertig et al., 2002; Drummond, 2007).

Table 1Location and depth data for wells sampled from the Carrizo Sand and Aquia aquifers.

Carrizo Sand aquifer Latitude Longitude Well depth(m)

F-1 29°6′57″ N 98°40′58.6″ W NRBCW-5 29°4′41.5″ N 98°41′23.2″ W NRKS-1 29°5′57.7″ N 98°36′9.6″ W 122N-3 29°2′48.1″ N 98°38′44.4″ W 189G-1 29°4′3.8″ N 98°33′20.1″ W 145Poteet-1 29°2′4.2″ N 98°34′25.9″ W 289PCW-1 28°57′34.8″ N 98°29′19.8″ W 488Peeler 28°43′44.2″ N 98°36′11.3″ W 655AC74R-1 28°40′14.2″ N 98°22′17.2″ W 1317SMA74R-2 28°38′41.6″ N 98°20′42.4″ W 1420

Aquia aquifer Latitude Longitude Screened interval(mbgs)a

KE 66 0125 39°16′2.1″ N 76°2′8.42″ W 24.2–35.2QA 88 1277 39°11′39.1″ N 75°57′36.3″ W 46–49QA 94 1322 39°8′39.9″ N 75°51′54.1″ W 82.3–88.3AA DE 218 38°55′52.98″ N 76°34′18.96″ W 24–26.5QA 88 1268 39°1′29.58″ N 76°0′39.6″ W 110–113SHA 39°1′50.7″ N 76°1′16.7″ W NRQA 94 0111 39°1′37.8″ N 76°1′17.04″ W 105–110QA 95 0611 39°1′28.74″ N 76°2′22.26″ W 104b

QA 73 3626 39°1′33.12″ N 76°8′48.7″ W 55–61QA 81 2298 38°56′33.9″ N 76°9′45.2″ W 98–101QA 92 0451 38°57′36.6″ N 76°12′51.2″ W NRQA EB 144 38°58′46.98″ N 76°18′46.98″ W NR

NR = not reported.a Meters below ground surface (see Haque et al., 2008).b Well depth, screened interval is not reported.

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Water from the Aquia aquifer is an important domestic drinking waterresource in the Chesapeake Bay region (Drummond, 2001; Drummondand Bolton, 2010). The Aquia is overlain by the confiningMarlboro clayand is underlain by silts and clays of the Cretaceous Severn Formation(Chapelle, 1983). The Aquia aquifer consists of Paleocene and Eocenemarine deposits composed of quartz sand (~55%), glauconite (~30%),and shell fragments (~8%), in addition to clay minerals, iron oxides/oxyhydroxides (i.e., magnetite, hematite, and minor goethite), garnet,hornblende, pyrite, and lignite (Page, 1957; Hansen, 1974; Chapelle,1983; Haque et al., 2008). The aquifer ranges in thickness from 23 to30 mnear the Chesapeake Bay (Drummond, 2001, 2007). Groundwaterflow patterns within the Aquia are dominated by flow from the re-charge areas towards the south and east, although the predevelopmentpotentiometric surface indicates that flowwithin portions of the aquiferon the Eastern Shore of the Chesapeake Bay historically flowed southand west towards the Bay (Fig. 2; Chapelle, 1983; Drummond, 2001;Aeschbach-Hertig et al., 2002; Drummond, 2007; Haque et al., 2008;Drummond and Bolton, 2010). In this study, groundwaters werecollected from a flow path located on the Eastern Shore that waspreviously investigated owing to the occurrence of a local As hot-spot(Haque et al., 2008; Drummond and Bolton, 2010; Willis et al., 2011).We note that the majority of other investigations of the Aquia aquiferchiefly focused on flow paths on the Western Shore of ChesapeakeBay (e.g, Chapelle, 1983; Chapelle and Knobel, 1983, 1985; Purdy etal., 1996; Aeschbach-Hertig et al., 2002;Willis and Johannesson, 2011).

3. Methods

3.1. Field sample collection

All sample bottles (HDPE), Teflon® tubing used during samplecollection and filtration, and HDPE and Teflon® labware were cleanedprior to use via standard trace element cleaning procedures (e.g., seeJohannesson et al., 2004). Groundwater samples for W analysis werecollected (March 2006) from 10 different wells finished within theCarrizo Sand aquifer and from 12 wells completed within the Aquia

(August 2006) aquifer (Table 1; Figs. 1, 2). All of the wells sampledfrom the Carrizo Sand aquifer are production wells, and as such,were purged with dedicated high capacity pumps for at least 30 minprior to sample collection. The wells sampledwithin the Aquia aquiferwere generally private, domestic wells. These wells were purged of atleast three well volumes before sample collection commenced. Addi-tionally, to ensure that water samples collected from both aquiferswere representative of groundwater from the aquifer and not waterfrom the well bore, water temperature, specific conductance, and pHwere continuously monitored during the purging process and samplecollection did not commence until all three parameters stabilized.

Upon completion of the well purging, dissolved oxygen (DO),oxidation–reduction potential (i.e., Eh inmV), pH, specific conductance,total dissolved solids (TDS) and temperature (°C) were determined forgroundwaters from each well using a Hydrolab MiniSonde5 equippedwith a flow through cell. The MiniSonde5 was attached to each wellby connecting the flow through cell to the well outflowwith a polyeth-ylene tube, which allowed dissolved oxygen and oxidation–reductionpotential for each groundwater sample to be measured without expos-ing the groundwater to the atmosphere. A number of more quantitativeredox sensitive parameters were also measured on-site via UV/Visspectrophotometry (Hach DR/2800) using separate, filtered (0.45 μmpore-size, polyether sulfone membrane, see below) aliquots of ground-water samples collected from each well. Dissolved Fe speciationwas measured in the field using the 1,10 Phenanthroline method forFe(II) and the Ferrozine method after reduction with hydroxylamineHCl for total Fe (FeT; Stookey, 1970; Eaton et al., 1995; Hach, 2007).Ferric iron was then determined by difference. The method detectionlimits for FeT and Fe(II) are 0.16 μmol kg−1 and 0.36 μmol kg−1,respectively (Haque et al., 2008). Total dissolved Mn concentrationswere also determined spectrophotometrically using an adaptation ofthe PAN method (Goto et al., 1977). The method detection limit fordissolved Mn by the PAN method is ~0.13 μmol kg−1. Dissolved sul-fide [hereafter S(-II)] was measured on separate aliquots using thesame portable spectrophotometer and the Methylene Blue method(Cline, 1969; Eaton et al., 1995). The method detection limit forS(-II) is 0.16 μmol kg−1. Dissolved NO3

− and NH4+ concentrations

were quantified in Aquia aquifer groundwaters via UV/Vis spectro-photometry using the cadmium reduction method (Gal et al., 2004)and the salicylate method (Krom, 1980), respectively. The spectropho-tometer was calibrated prior to analysis using standard curves thatwere generated for each of the measured parameters using standardsolutions purchased from Hach Corporation. Alkalinity was titratedin the field on separate sample aliquots of known volume (25 mL)using a “digital” titrator (Hach, Model 16900) and either 1.6 N or0.16 N H2SO4.

Groundwater samples forW analysis were collected and processedin the field using trace element cleanmethods that followed the “cleanhands, dirty hands” protocol (i.e., Fitzgerald, 1999) to minimize po-tential sample contamination. Briefly, previously cleaned (i.e., tracemetal cleaned; Johannesson et al., 2004) low-density polyethylene(LDPE) cubitainers® (Hedwin Corporation, Baltimore, MD, USA) wererinsed three times with groundwater from each wellhead, before fillingthe cubitainer® with groundwater from the well. All personnel woreclean polyethylene gloves during sample collection and processing atthe field sites. The collected groundwater samples were then filteredthrough 0.45 μm pore-size, in-line filter capsules (Gelman Sciences,polyether sulfone membrane, Pall Corporation, Port Washington, NY)by drawing water from the cubitainer® through pre-cleaned Teflon®tubing with a field portable peristaltic pump. The filtered groundwaterwas pumped through the filter capsule and directly into pre-cleaned1 L high-density polyethylene (HDPE) sample bottles. These samplesbottles were rinsed three times with the filtered groundwater sampleto precondition the sample bottle before filling the sample bottle withthe filtered groundwater sample. Groundwater samples for tungstenanalysis were then immediately acidified to pH b 2 using ultrapure

Fig. 2.Map of the Aquia aquifer study site in eastern central Maryland, USA. The aquifer is recharged within the outcrop/subcrop region (dashed green lines; Drummond and Bolton,2010) and groundwater generally flows to the south and east beneath the Chesapeake Bay. However, the flow path chosen here for study (i.e., large, curving white arrow) is basedon the predevelopment (pre-pumping) potentiometric surface (Drummond, 2001, 2007). All subsequent figures showing geochemical parameters plotted as a function of distancealong the flow path are projected onto this general flow path line. Base satellite image is from Google Earth. (For interpretation of the references to color, the reader is referred to theweb version of the article.)

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nitric acid (Seastar Chemicals Inc., Baseline®), the caps tightly sealed,the bottles doubled bagged in pre-cleaned polyethylene, Zip-lock®-style bags, and then stored at ~4 °C until analysis.

In addition to water samples, we also collected sediment samplesfrom the Carrizo Sand and Aquia aquifers to investigate the solid-phase speciation of W in both aquifers. Four samples of the CarrizoSand aquifer were obtained from depths ranging between 900and 1030 m below ground surface (i.e., 900–930 m, 930–960 m,960–990 m, and 990–1030 m) from a recently drilled well in AtascosaCountry, Texas. Samples of the Aquia aquifer were acquired from apre-existing sediment drill-core stored in the core library at the

Maryland Geological Survey (see Haque et al., 2008, for details).Five separate samples of the Aquia aquifer sediment core were col-lected from different depth intervals from the archived sedimentcore (i.e., 61 m, 72 m, 86.5 m, 97 m, and 105 m) using clean techniques(Haque et al., 2008).

3.2. Sequential extraction procedure

To characterize the distribution of W in Carrizo Sand and Aquiaaquifer sediments, a sequential extraction procedure based on themethod of Tessier et al. (1979) and described in detail by Tang et al.

Table 2Geochemical reactions and associated equilibrium constants used tomodel the aqueousgeochemistry of tungsten in Carrizo Sand and Aquia aquifer groundwaters.

Association (protonation) of tungstic acid Association constantsa

H+ + WO42− ↔ HWO4

− log Ka,1 = 3.62H+ + WO4

2− ↔ H2WO40 log Ka,2 = 5.8

Thiotungstate complexation Stability constantsb

WO42− + H2S ↔ WO3S2− + H2O log β01 = 5.19

WO3S2− + H2S ↔ WO2S22− + H2O log β12 = 4.8WO2S22− + H2S ↔ WOS32− + H2O log β23 = 5WOS32− + H2S ↔ WS42− + H2O log β34 = 4.88

a Infinite dilution (I = 0 mol kg−1); Smith and Martell (2004).b Stability constants for formation of analogous thiomolybdate complexes (Erickson

and Helz, 2000).

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(2004) and Willis and Johannesson (2011) was employed. The se-quential extraction procedure was used to investigate the speciationof W between five operationally defined phases within the aquifersediments that include the exchangeable, acidic (i.e., carbonateminerals), reducible, oxidizable, and the residual phases of W in thesediments. In each case ~0.59 mg of sediment sample was employedfor the sequential extractions. The reader is referred to Willis andJohannesson (2011) for complete description of the procedures andreagents used in the sequential extractions.

3.3. Tungsten analysis

Tungsten concentrations were quantified in the Carrizo Sandand Aquia aquifer groundwaters and sediment extractions usinghigh resolution (magnetic sector) inductively coupled plasma massspectrometry (HR-ICP-MS, Finnigan MAT Element 2) within twoweeks of sample collection, or in the case of the sequential extrac-tions, within one week of completion of the extractions. Two W iso-topes (182W and 183W) were monitored during the analysis becauseneither has any direct isobaric interferences. The principal molecular(polyatomic) interferences that can occur during analysis of theseW isotopes by ICP-MS involve oxides of Er (i.e., 166Er16O+ on 182Wand 167Er16O+ on 183W). However, because Er concentrations inthese groundwaters are low (e.g., 3.99 ± 6 pmol kg−1 for CarrizoSand aquifer groundwater and 4.3 ± 0.7 pmol kg−1 for groundwaterfrom the Aquia aquifer; Tang and Johannesson, 2006; Willis andJohannesson, 2011) and formation within the plasma was generallyless than 0.3%, polyatomic interferences on W by Er oxides werenegligible. Moreover, the HR-ICP-MS allows for the clear delineationand separation of the W peaks from those of the polyatomic Er peaks.

All sample aliquots were spiked with 4 μg kg−1 of 115In as aninternal standard to monitor for instrument drift during analysis.The instrument was calibrated and the sample concentrations ver-ified using a series of W calibration standards of known concen-trations (e.g., 1 ng kg−1, 2 ng kg−1, 10 ng kg−1, 100 ng kg−1,500 ng kg−1, 1000 ng kg−1) prepared from an NIST traceable HighPurity (Charleston, SC, USA) single element standard. In addition,check standards prepared from Perkin-Elmer multielement solutionswere analyzed during the sample runs to certify accuracy. Both theNIST Standard Reference Material (SRM) “Trace Elements in Water”No. 1643e and the National Research Council Canada (Ottawa, Ontario,Canada) SRM for trace metals in river waters (SLRS-4) were routinelyanalyzed for other trace elements to ensure accuracy. Analytical preci-sionwas typically better than 3.24% RSD (relatively standard deviation)for W using the HR-ICP-MS.

3.4. Geochemical modeling

Aqueous and surface complexation modeling of tungsten ingroundwaters of the Carrizo Sand and Aquia aquifers was accom-plished using the React program of the Geochemist's Workbench®(release 8.0; Bethke, 2008; Bethke and Yeakel, 2010). The LawrenceLivermore National Laboratory database (i.e., thermo.dat; Delanyand Lundeen, 1989) provided with the software was modified toinclude protonation of the tungstate oxyanion (i.e., association oftungstic acid) using the equations and infinite dilution associationconstants from Smith and Martel (2004) that are listed in Table 2.These values are similar to those reported by Wesolowski et al (1984)and Yatsimirskii and Romanov (1965), and recently employed byPetrunic and Al (2005) to model the W geochemistry of mine tailingsin New Brunswick, Canada.

In addition, because W and molybdenum (Mo) exhibit broadlysimilar chemical properties (e.g., Cotton et al., 1999), and owing tothe fact that the molybdate oxyanion (MoO4

2−) forms thiomolybdatecomplexes (i.e., MoOxS4 − x

2− , x = 0–4) in sulfidic waters (Helz et al.,1996; Erickson and Helz, 2000; Helz et al., 2011), it is reasonable to

expect that the tungstate oxyanion also reacts with dissolved S(-II)in sulfidic waters, forming analogous thiotungstate complexes (Daveand Johannesson, 2008a,b; Haug et al., 2011). Consequently, the forma-tion of thiotungstate complexes in Carrizo Sand and Aquia aquifergroundwaters was modeled by adding the four reactions listed inTable 2 that describe the conversion of tungstate, WO4

2−, intotetrathiotungstate, WS42−, to the thermo.dat database. However, be-cause reliable stability constants for the formation of the fourthiotungstate complexes do not exist in the literature, we assumehere that the stability constants for the thiotungstates are similarto those of analogous thiomolybdates, which were determined previ-ously (Erickson and Helz, 2000), and use the thiomolybdate constantsas proxies in our computations. It is important to note, however,that our preliminary analyses suggest that thiotungstates are lessstable than thiomolybdates, and thus the actual stability constantsfor the thiotungstates are lower than those describing the formationof thiomolybdates (Mohajerin and Johannesson, 2012).

Surface complexation ofWO42−wasmodeled using the generalized

two-layer model of Dzombak and Morel (1990) and assuming thatferrihydrite (i.e., hydrous ferric oxide or HFO) was the adsorbingFe(III) oxide/oxyhydroxide in each aquifer (e.g., Gustafsson, 2003;Seiler et al., 2005). The surface complexation reactions used tomodel the adsorption of WO4

2− onto ferrihydrite are presented inTable 3 (also see Results section). Ferrihydrite concentrations wereestimated by extracting aliquots of Carrizo Sand and Aquia aquifersediments with hydroxylamine hydrochloride (i.e., fraction 3 of thesequential extractions). The extracted Fe concentrations were subse-quently used along with the aquifer porosity and the average densityfor sediment mineral grains (~2.6 g cm−3) to compute the amountof free ferrihydrite in the aquifer sediments following the approachoutlined in Welch and Lico (1998). We assumed an average porosityof 35% for the Carrizo Sand aquifer, which falls in the middle of thereported range (Pearson and White, 1967), and 30% for the Aquiaaquifer (Chapelle and Drummond, 1983), and used a gram-formulaweight of 89 g mol−1 for ferrihydrite (Dzombak and Morel, 1990).Based on the mean Fe concentrations extracted from the CarrizoSand and Aquia aquifer (i.e., 1579 μmol kg−1 and 14,593 μmol kg−1,respectively; Haque, 2007; Mohajerin and Johannesson, unpubl.),solid ferrihydrite contents of 0.852 g L−1 and 7.88 g L−1 are calculatedfor the Carrizo Sand and Aquia aquifers, respectively.

Because other anions can compete with WO42− for surface sites

(e.g., Gustafsson, 2003; Seiler et al., 2005) we also included surfacecomplexation reactions for both dissolved silica (Si) species and arsenic(As) species in the model (Table 3). Inclusion of Si and As species re-flects the fact that these species have been determined in groundwatersfrom both aquifers (Haque and Johannesson, 2006; Haque et al., 2008;Willis and Johannesson, 2011; Willis et al., 2011). Unfortunately, phos-phate (PO4

3−) was generally below detection (0.3–0.4 μmol kg−1) ingroundwaters from both aquifers (e.g., Tang and Johannesson, 2005;Drummond and Bolton, 2010), thus preventing a direct examination

Table 3Parameters used to model surface complexation of tungstate, silicic acid, arsenate,and arsenite onto ferrihydrite (i.e., hydrous ferric oxide — HFO), including surfacecomplexation reactions and associated intrinsic surface complexation constants. Weassumed that HFO and ferrihydrite are synonymous and that the solid ferrihydrite/HFO concentrations are 0.852 g L−1 and 7.88 g L−1 for the Carrizo Sand and Aquiaaquifers, respectively (see text).

Mineral Surface aream2 g−1

Weak sites(mol/mol mineral)

Strong sitesb

(mol/mol mineral)

Ferrihydritea 600 0.2 0.005

Surface acidity reactionsa Acidity constantsa

≡FeOH0 + H+ ↔ ≡ FeOH2+ log Ka1 = 7.92

≡FeOH0 ↔ ≡ FeO− + H+ log Ka2 = −8.93

Surface complexation reactions Surface complexationconstants

Tungstatec

≡FeOH0 + WO42− ↔ ≡ FeWO4

− + H2O log K2int = 10.28

≡FeOH0 + WO42− ↔ ≡ FeOHWO4

2 log K3int = 5.89

Silicic acidd

≡FeOH0 + H4SiO40 ↔ ≡ FeH3SiO4

0 + H2O log K1int = 4.28

≡FeOH0 + H4SiO40 ↔ ≡ FeH2SiO4

− + H2O + H+ log K2int = −3.22

≡FeOH0 + H4SiO40 ↔ ≡ FeHSiO4

2− + H2O + 2H+ log K3int = −11.69

Arsenatea

≡FeOH0 + AsO43− + 3H+ ↔ ≡ FeH2AsO4° + H2O log K1

int = 29.31≡FeOH0 + AsO4

3− + 2H+ ↔ ≡ FeHAsO2− + H2O log K2

int = 23.51≡FeOH0 + AsO4

3− ↔ ≡ FeOHAsO43− log K4

int = 10.58Arsenitea,e

≡FeOH0 + H3AsO30 ↔ ≡ FeH2AsO3

0 + H2O log K1int = 4.02

≡FeOH0 + HAsO32− + H+ ↔ ≡ FeHAsO3

− + H2O log K2int = 18.46

a Dzombak and Morel (1990).b Not significant for anion adsorption (Dzombak and Morel, 1990).c This study (see below).d Swedlund and Webster (1999).e Dixit and Hering (2003).

Table 4Concentrations of field parameters, major solutes, and redox sensitive indicator species iNote that distances are in kilometers along the flow path as defined in Tang and Johanneand Johannesson (2006), Haque and Johannesson (2006), Basu et al. (2007), and Johanness

F-1 BCW-5 KS-1 N-3 G-1

Distance 0.5 4.29 7.4 9.37 12.8pH 6.51 ± 0.07 6.69 ± 0.15 6.41 ± 0.2 6.14 ± 0.07 6.21 ±Temp °C 24.7 24.6 23.8 26.2 25.8Cond (μS cm−1) 559 302 335 166 274TDS (mg kg−1) 358 194 214 108 175Eh (mV) 112 ± 215 130 ± 63 297 ± 19 174 ± 10 135 ±

mmol kg−1

Ca ND 1.35 0.63 0.38 NDMg ND 0.34 0.2 0.12 NDNa ND 1.42 1.12 0.71 NDK ND 0.19 0.13 0.13 NDCl ND 2.11 1.43 1.06 NDAlk 1.64 1.40 1.02 0.46 0.78SO4 ND 0.72 0.27 0.22 NDSiO2 0.52 0.16 0.38 0.29 ND

μmol kg−1

FeT 20.8 ± 9.9 15.8 1.08 ± 0.49 19.7 ± 11.5 35.7 ±Fe(II) 19.8 ± 10.5 7.52 0.72 16.2 ± 6.9 32.7 ±Fe(III) 1 ± 14.4 8.23 0.36 ± 0.49 3.5 ± 13.4 3 ± 6Mn 0.8 1.02 0.2 0.75 0.59S(-II) 0.98 ± 1.26 1.28 ± 0.72 BD* 0.45 ± 0.22 0.84 ±DO 8.75 8.13 11.3 123 4.69

nmol kg−1

AsT 2.19 ± 0.12 2.08 ± 0.11 0.97 ± 0.05 2.44 ± 0.13 2.47 ±As(III) 0.6 ± 0.03 0.56 ± 0.03 0.41 ± 0.02 0.45 ± 0.02 0.87 ±As(V) 1.59 ± 0.09 1.52 ± 0.08 0.56 ± 0.03 1.99 ± 0.11 1.6 ±

*BD = below detection. Detection limits are: S(-II) = 0.16 μmol kg−1; FeT = 0.16 μmol kgND = not determined, or could not be determined.

82 K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

of the competitive adsorption effects of PO43− in the model simulations

(e.g., see Gustafsson, 2003; Xu et al., 2009).

4. Results

4.1. General geochemistry

Field parameters, major solute concentrations, and redox indica-tors measured in groundwaters from the Carrizo Sand and Aquiaaquifers are presented in Tables 4 and 5, respectively. Variations inthemeasured field parameters (i.e., pH, conductivity, and temperatureare plotted as functions of distance along the flow paths in the CarrizoSand 2003, 2004, and 2006 field campaigns) and Aquia (2006 fieldcampaign) aquifers in Figs. 3 and 4, respectively. The data for theCarrizo Sand aquifer demonstrate that during the period of study,the geochemistry of the groundwater was relatively stable alongthe flow path. Additional sampling of the Aquia aquifer is needed,however, to establish possible temporal variations in groundwatercomposition. Nonetheless, because the Aquia aquifer is confined andcontains groundwater as old as 100,000 years in age (Purdy et al.,1996), it is reasonable to assume steady-state conditions with regardto the groundwater chemical composition as a first approximation.

Groundwaters from the studied portions of the Carrizo Sand andAquia aquifers are dilute and of circumneutral pH (Figs. 3, 4). Withinthe Carrizo Sand aquifer pH ranges from 6.14 at the BCW-5 well nearthe recharge zone, to >8.5 at the end of the studied flow path (Fig. 3).In comparison, the pH of groundwaters from the Aquia aquiferexhibits a narrower range (i.e., 7.6 ≤ pH ≤ 8.4; Fig. 4). Alkalinity(as HCO3

−) ranges from 0.46 to 7.84 mmol kg−1 in groundwatersfrom the Carrizo Sand aquifer, and from 2.48 to 4.16 mmol kg−1 inAquia aquifer groundwaters (Tables 4, 5). For both aquifers, pHand alkalinity generally increase along the direction of groundwater

n groundwaters from the Carrizo Sand aquifer. Alkalinity (Alk) is reported as HCO3−.

sson (2005, 2006) and Haque and Johannesson (2006). Data summarized from Tangon and Neumann (2013).

Poteet PCW-1 Peeler-1 AC74R-1 SMA74R-2

14.5 26.1 41.1 59 65.80.07 6.36 ± 0.18 7.27 ± 0.1 8.5 ± 0.18 8.6 ± 0.09 8.3 ± 0.22

28.2 34.3 27.3 55 31.8387 590 495 974 1050247 378 317 623 673

3 165 ± 9 48 −113 ± 2 −136 ± 1 −118 ± 5

0.76 2.01 0.45 0.06 0.080.23 0.33 0.36 0.01 0.011.16 1.15 3.96 8.74 10.240.2 0.14 0.15 0.05 0.061.44 0.87 0.37 1.23 1.780.94 4.1 4.63 6.56 7.840.36 0.37 0.35 0.79 0.710.27 0.23 0.26 0.43 0.48

5.59 20.4 ± 5.44 15.1 ± 6.22 14.3 1.67 2.18 ± 1.732.97 15.9 ± 1.91 9.13 ± 1.51 3.22 1.43 1.26 ± 1.27

.33 4.5 ± 5.8 5.97 ± 6.4 11.1 0.24 0.92 ± 2.151.7 3.83 2.9 0.45 1

0.14 0.44 ± 0.18 0.91 ± 0.18 2.73 ± 0.1 6.04 ± 0.26 39.9 ± 21.28.75 16.4 6.23 5.63 4.06

0.13 2.33 ± 0.13 0.37 ± 0.02 0.91 ± 0.05 0.41 ± 0.02 1.92 ± 0.10.05 0.6 ± 0.03 0.43 ± 0.02 0.69 ± 0.04 0.29 ± 0.02 0.55 ± 0.03

0.08 1.73 ± 0.1 BD 0.22 ± 0.01 0.12 1.37 ± 0.07

−1; Fe(II) = 0.36 μmol kg−1; and Mn = 0.13 μmol kg−1.

Table 5Concentrations of field parameters, major solutes, and redox sensitive indicator species in groundwaters from the Aquia aquifer. Alkalinity (Alk) is reported as HCO3

−. Note that distances are in kilometers along the flow path as defined inHaque et al. (2008). Data summarized from Haque et al. (2008), Willis and Johannesson (2011), and Willis et al. (2011).

KE 66 QA 88 QA 94 AA DE QA 88 SHA QA 94 QA 95 QA 73 QA 81 QA 92 QA EB

0125 1277 1322 218 1268 0111 0611 3626 2298 0451 144

Distance 22.5 33.5 36.2 40 51.2 52 52.5 55 64.5 70.7 75.1 86.3pH 7.57 7.69 8.01 7.81 8.08 8.38 7.91 8.02 7.84 8.3 7.93 8.12Temp °C 16.6 15.6 18.8 14.5 16.7 19.7 16.2 16.1 16.9 17.5 17.8 15.9Cond (μS cm−1) 222 296 330 300 303 305 366 302 351 296 422 366TDS (mg kg−1) 143 189 211 192 194 196 234 194 213 189 267 271Eh (mV) 25.5 44 −11 60.5 −10.5 −46 −3 −18 27.5 −11 11.5 −53

mmol kg−1

Ca 0.925 1.42 0.943 1.6 0.715 0.317 0.985 0.81 1.24 0.669 0.693 1.23Mg 0.083 0.137 0.647 0.054 0.601 0.23 0.752 0.532 0.525 0.548 0.426 0.313Na 0.084 0.179 0.382 0.037 0.484 2.39 0.348 0.355 0.358 0.607 0.522 0.102K 0.056 0.072 0.313 0.087 0.362 0.152 0.393 0.323 0.239 0.363 1.27 1.08Cl 0.205 0.062 0.026 0.07 0.022 0.043 0.053 0.039 0.057 0.038 0.042Alk 2.99 2.79 3.08 2.48 3.19 2.91 3.79 2.98 3.46 3.79 4.16 3.24SO4 0127 0.077 0.028 0.52 0.039 0.042 0.058 0.071 0.041 0.054 0.05 0.069SiO2 1.26 ± 0.05 0.41 0.32 0.37 0.44 0.24 0.23 0.57 0.44 0.2 0.38 0.22

μmol kg−1

FeT 12.7 ± 0.21 17.7 ± 0.01 15.4 ± 0.05 14 ± 0.01 14.7 ± 0.08 7.01 ± 0.06 11.4 ± 0.10 3.35 ± 0.02 BD 9.94 ± 0.03 9.01 ± 0.01 8.29 ± 0.01Fe(II) 0.90 ± 0.01 7.70 2.87 ± 0.10 5.32 ± 0.1 0.18 2.38 ± 0.10 6.80 ± 0.10 3.35 ± 0.09 BD 9.94 ± 0.27 9.01 ± 0.24 4.71 ± 0.13Fe(III) 11.8 ± 2.58 10.0 ± 0.27 12.5 ± 0.34 8.7 ± 0.23 14.5 ± 0.39 4.63 ± 0.12 4.60 ± 0.12 ND ND ND ND 3.58Mn 1.02 0.15 ± 0.01 BD 0.76 ± 0.02 0.42 ± 0.01 BD BD BD 0.44 ± 0.01 0.04 ± 0.01 BD 0.11 ± 0.01S(-II) BD* 0.23 ± 0.18 BD 0.52 ± 0.18 0.10 ± 0.18 0.40 ± 0.02 BD BD BD 0.34 ± 0.03 BD 0.16DO 51 ± 1.8 13.5 ± 0.21 19.6 ± 0.1 124 ± 1.8 16.8 ± 0.07 21.4 ± 0.11 BD 50 23.3 ± 0.12 20.3 ± 0.05 12 ± 0.1 9.38 ± 0.08NO3 BD 0.48 1.13 0.32 0.32 3.23 11.1 ± 0.05 0.76 ± 0.09 1.13 0.65 0.48 0.48NH4 0.42 ± 0.19 2.03 ± 0.19 14.8 BD 9.12 ± 0.93 6.77 3.22 4.83 7.73 15.1 9.02 1.39 ± 0.19

nmol kg−1

AsT 4.92 ± 0.17 0.75 ± 0.03 291 ± 9.95 1.66 ± 0.06 512 ± 17.5 564 ± 19.3 1072 ± 36.7 411 ± 15.1 52.7 ± 1.8 241 ± 8.24 76.8 ± 2.63 27.3 ± 0.93As(III) 2.71 ± 0.07 0.59 ± 0.02 173 ± 4.64 1.5 ± 0.04 505 ± 13.5 703 ± 18.8 980 ± 26.3 446 ± 11.9 39.4 ± 1.06 198 ± 5.31 70.7 ± 1.89 24.8 ± 0.66As(V) 2.21 ± 0.07 0.15 118 ± 3.6 0.17 ± 0.01 7.25 ± 0.22 ND 91.9 ± 2.8 ND 13.2 ± 0.4 44 ± 1.34 6.05 ± 0.08 2.51 ± 0.08

*BD = below detection. Detection limits are as in Table 4.ND = not determined, or could not be determined.

83K.H.Johannesson

etal./

ChemicalG

eology351

(2013)76

–94

pH

6.0

6.5

7.0

7.5

8.0

8.5

9.0

200320042006

Co

nd

uct

ivit

y (µ

S/c

m)

0

200

400

600

800

1000

200320042006

Kilometers

0 10 20 30 40 50 60 70

Tem

per

atu

re (

oC

)

20

25

30

35

40

45

50

55

200320042006

Flow Direction

Fig. 3. Field parameter data for pH, conductivity (μS cm−1), and temperature (°C) forCarrizo Sand aquifer groundwaters plotted as a function of distance along the flowpath shown in Fig. 1 (i.e., white arrow). Data shown are for samples collected duringthe 2003, 2004, and 2006 field sampling campaigns (Haque and Johannesson, 2006;Tang and Johannesson, 2006; Basu et al., 2007; Willis et al., 2011).

pH

7.4

7.6

7.8

8.0

8.2

8.4

8.6

200

250

300

350

400

Kilometers

10 20 30 40 50 60 70 80 90

Tem

per

atu

re (

o C)

14

16

18

20

Flow Direction

Co

nd

uct

ivit

y (µ

S/c

m)

Fig. 4. Field parameter data for pH, conductivity (μS cm−1), and temperature (°C) forAquia aquifer groundwaters plotted as a function of distance along the flow pathshown in Fig. 2 (i.e., white arrow).

84 K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

flow (Figs. 3, 4). Conductivity and the concentrations of total dissolvedsolids (TDS) are low in groundwaters from both aquifers, but alsoshow general increases along the direction of groundwater flow. Amajor difference between the Carrizo Sand and Aquia aquifers is evidentin the groundwater temperatures, which are relatively constant in theAquia aquifer (i.e., mean ± SD = 16.9 ± 1.4°), but range from ~24 °Cup to 55 °C in the Carrizo Sand aquifer (Figs. 3, 4). The difference inthe distribution of groundwater temperatures between the two aquifersreflects the relatively shallow depths of the Aquia aquifer comparedto the substantially greater depths of down-gradient portions of theCarrizo Sand aquifer (e.g., Table 1).

Carrizo Sand groundwaters are Ca–Na–HCO3–Cl type waters nearthe recharge area, where well depths are generally less than 150 m.With flow down gradient along the flow path, Carrizo Sand ground-waters become Ca–HCO3 and Ca–Na–HCO3 type waters along themid-reaches of the studied flow path (well depths between 150 mand 760 m), and evolve to Na–HCO3 waters further down gradient(well depths > 760 m; Table 4; Tang and Johannesson, 2005).The evolution of the major solute composition of Carrizo Sand aquifergroundwater is consistent with a number of processes occurringwithin the aquifer including dissolution of calcite, oxidation of ironsulfides and precipitation of iron oxides/oxyhydroxides in shallowportions of the aquifer, dissolution of chloride-bearing minerals, andchemical weathering of alumino-silicate minerals further down gra-dient along the flow path (Hamlin, 1988).

The changes in composition of Aquia aquifer groundwaters areconsistent with freshening of sediments and associated pore watersinitially deposited in a marine environment with infiltration of dilutemeteoric waters (Valocchi et al., 1981; Chapelle and Knobel, 1983;Appelo, 1994). More specifically, as fresh, meteoric water infiltratesthe Aquia aquifer, it displaces relict seawater leading to chromato-graphic ion-exchange between Ca2+ in the fresh, meteoric water andNa+, K+, and Mg2+ that sorbed onto mineral surface sites when theAquia aquifer was saturated with seawater (Appelo, 1994). Fresheningof the Aquia aquifer can explain the progressive changes in groundwa-ter composition along the flow path, from Ca(Mg)–HCO3 near the re-charge area to Na–HCO3 type waters further down gradient (Chapelleand Knobel, 1983; Appelo, 1994). The substantial increase in HCO3 con-centrations beyond ~40 km along the flow path studied by Chapelleand Knobel (1983) was attributed to dissolution of carbonate shellswithin the aquifer sediments.

Petrographic analysis reveals that Fe(III) oxides/oxyhydroxides areubiquitous in both aquifers, occurring principally as coatings on quartzand other mineral grains (Basu et al., 2007; Haque et al, 2008).Furthermore, X-ray diffraction (XRD) and diffuse reflectance spectros-copy indicates that the Fe(III) oxides/oxyhydroxides chiefly consistof goethite with minor amounts of hematite, although the presenceof ferrihydrite cannot be ruled out. Petrographic analysis also suggeststhat magnetite occurs within the Aquia aquifer sediments (Haqueet al., 2008). In addition to Fe(III) oxides/oxyhydroxides, kaoliniteand illite are abundant within the surface coatings on mineral grains

W (

pm

ol/k

g)

0

40

80

120

160

200

µ m

ol/k

g

0

5

10

15

Fe(II)Fe(III)

µ m

ol/k

g

0

5

10

15

NO3-

NH4+

Mn

kg

120S(-II)Dissolved Oxygen

a)

b)

c)

d)

85K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

within the Carrizo Sand, and smectites occur in the mineral coatingswithin the Aquia aquifer sediments (Basu et al., 2007; Haque et al.,2008). Nevertheless, the most abundant clay mineral within Aquiaaquifer sediments is glauconite.

4.2. Redox sensitive parameters

Dissolved Fe species, sulfide, and oxygen concentrations are listedalong with measured Eh values for Carrizo Sand and Aquia aquifergroundwaters in Tables 4 and 5, respectively. In general, Eh decreaseswith flow down gradient along the flow paths in both aquifers,although the decrease in Eh is substantially greater in the Carrizo Sandaquifer (i.e., from 297 to −136 mV) than in the Aquia aquifer (60.5 to−53 mV; Figs. 5, 6). Except for well N-3 from the Carrizo Sand aquifer,dissolved oxygen concentrations are low, exhibiting a mean ± SDof 8.2 ± 3.8 μmol kg−1. Dissolved oxygen concentrations in Aquiaaquifer groundwaters (mean ± SD = 31.1 ± 32.6 μmol kg−1) are,on average, higher than in Carrizo groundwaters, with groundwa-ters from three wells in the Aquia exhibiting dissolved oxygenconcentrations ≥50 μmol kg−1 (Table 5). If these three wellswith oxic groundwaters are excluded, the mean ± SD dissolved

W (

pm

ol/k

g)

0

400

800

1200

1600

µm

ol/k

g

0

10

20

30 Fe(II)Fe(III)

S (

-II)

(µ

mo

l/kg

)

0

10

20

30

40

µm

ol/k

g

0

40

80

120Dissolved OxygenMn

Kilometers

0 10 20 30 40 50 60 70

Eh

(m

V)

-200

-100

0

100

200

300

a)

b)

c)

d)

e)Groundwater Flow

Fig. 5. Geochemical data for Carrizo Sand aquifer groundwaters plotted as a function ofdistance along the studied flow path. Shown are: (a) dissolved W concentrations(pmol kg−1); (b) dissolved Fe(II) and Fe(III) concentrations (μmol kg−1); (c) dissolvedS(-II) concentrations (μmol kg−1); (d) dissolved oxygen and Mn concentrations(μmol kg−1); and (e) Eh (mV).

µ m

ol/

0

40

80

Kilometers

10 20 30 40 50 60 70 80 90

Eh

(m

V)

-60

-40

-20

0

20

40

60

Groundwater Flow

e)

Fig. 6. Geochemical data for Aquia aquifer groundwaters plotted as a function ofdistance along the studied flow path. Shown are: (a) dissolved W concentrations(pmol kg−1); (b) dissolved Fe(II) and Fe(III) concentrations (μmol kg−1); (c) dissolvedNO3

−, NH4+, and Mn concentrations (μmol kg−1); (d) dissolved S(-II) and oxygen con-

centrations (μmol kg−1); and (e) Eh (mV).

oxygen concentration for groundwaters from the Aquia aquifer is16.4 ± 5 μmol kg−1.

Iron concentrations vary substantially along the flow path withineach aquifer (Tables 4, 5). In general, total dissolved Fe (i.e., FeT)decreases with flow down gradient along the flow path within eachaquifer, ranging from a high of ~36 μmol kg−1 to ~1 μmol kg−1 inthe Carrizo Sand aquifer, and from 18 μmol kg−1 to below detection(i.e., b0.16 μmol kg−1) in Aquia groundwater (Tables 4, 5). Ferrousiron predominates in the majority of Carrizo Sand aquifer groundwa-ter samples (Fig. 5). A large spike in dissolved Fe(II) is observed ingroundwaters of the Carrizo aquifer just down gradient from wherethe aquifer becomes confined (Fig. 5; Basu et al., 2007; Johannessonand Neumann, 2013). Within the Aquia aquifer, Fe(III) is the chiefform of iron in groundwaters along the first ~50 km of the flowpath, beyond which Fe(II) predominates (Fig. 6; Haque et al., 2008;Willis and Johannesson, 2011).

A major difference between the Carrizo Sand and Aquia aquifers isevident in the dissolved S(-II) concentrations (Tables 4, 5). DissolvedS(-II) concentrations in the Carrizo Sand aquifer groundwaters aregenerally b1 μmol kg−1 along the first 15 km of the studied flowpath, but subsequently increase in concentration with flow downgradient beyond the PCW-1 well (~26 km from the rechargezone), attaining a concentration of ~40 μmol kg−1 at the end of theflow path (Fig. 5; Haque and Johannesson, 2006; Basu et al., 2007;Johannesson and Neumann, 2013). Within the Aquia aquifer, however,

86 K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

S(-II) concentrations remain near or below the detection limit along theentire flow path (Fig. 6; Haque et al., 2008; Willis and Johannesson,2011).

4.3. Tungsten concentrations

Tungsten concentrations in groundwaters from the Carrizo Sandaquifer range from 1.9 pmol kg−1 up to 1297 pmol kg−1, increasingwith flow down gradient along the flow path (Table 6, Fig. 5).Tungsten concentrations in groundwaters from the Aquia aquifer aregenerally lower, and range between 6.5 and 184 pmol kg−1 (Table 6,Fig. 6). The mean W concentration of the Carrizo groundwaters(i.e., 248 pmol kg−1) is almost a factor of 4 larger than the mean Wconcentration of Aquia groundwaters (64 pmol kg−1). The differencein themeanW concentrations between the two aquifers chiefly reflectsthe substantially higher W concentrations measured in the down-gradient, sulfidic groundwaters of the Carrizo aquifer (Figs. 5, 6).

An outstanding feature of the W data is the monotonic increase inW concentration that occurs in the Carrizo Sand aquifer with flowdown gradient, which sharply contrasts with the highly variable Wconcentrations that characterize the Aquia aquifer groundwaters(Figs. 5, 6). Inspection of the data reveal strong and statistically signif-icant, positive correlations (p ≤ 0.001) between W and pH, alkalinity(as HCO3

−), and S(-II) concentrations (r = 0.78, 0.89, and 0.95,respectively) in the Carrizo Sand aquifer groundwaters, but onlyweak, inverse or none existent relationships between W and theseparameters in the Aquia aquifer (r = −0.4, −0.43, −0.08 for Wvs. pH, alkalinity, and S(-II), respectively). The relationship betweenW and these parameters in Aquia aquifer groundwater are eitherof low statistical significance (p ≤ 0.02) or not significant [i.e., W vs.S(-II); Dave and Johannesson, 2008a,b]. Moreover, correlations be-tween W and Fe(II) are low and statistically insignificant in ground-waters from both aquifers, although W is inversely related to thetotal groundwater iron concentration [FeT = Fe(II) + Fe(III)] in theCarrizo Sand aquifer (r = −0.61, p b 0.01), and positively relatedto FeT in Aquia aquifer groundwaters (r = 0.44, p b 0.02; Dave andJohannesson, 2008a,b). Previous studies have noted positive correla-tions between groundwater As and W concentrations (Seiler et al.,

Table 6Tungsten concentrations (in pmol kg−1) for groundwaters fromthe Carrizo Sand and Aquia aquifers.

Carrizo Sand aquifer W(pmol kg−1)

F-1 78.7BCW-5 9.46KS-1 3.64N-3 16.3G-1 18.1Poteet-1 1.9PCW-1 27.4Peller 246AC74R-2 780SMA74R-1 1297

Aquia aquifer W(pmol kg−1)

KE 66 0125 30.7QA 88 1277 184QA 94 1322 90.9AA DE 218 93.3QA 88 1268 6.5SHA 22.4QA 94 0111 14.3QA 95 0611 41.5QA 73 3626 132QA 81 2298 10.9QA 92 0451 16.6QA EB 144 124

2005). However, in both the Carrizo Sand and Aquia aquifer, W isinversely related to As concentrations (r = −0.17 and −0.56,respectively), and the relationship is not statistically significant inthe Carrizo Sand groundwaters. Tungsten exhibits a statistically sig-nificant, inverse relationship with measured Eh in the Carrizo Sandaquifer (r = −0.74, p ≤ 0.002), and a weak, positive relationshipwith Eh in Aquia aquifer groundwaters (r = 0.35, p b 0.02). Theinverse relationship between W and Eh observed for the CarrizoSand aquifer is broadly similar to observations of W in groundwatersfrom the Carson Desert of northwestern Nevada (e.g., Cutler, 2011).

Tungsten concentrations in the sequential extractions of CarrizoSand and Aquia aquifer sediments (Table 7) are presented in Figs. 7and 8, respectively. The sequential extraction experiments revealthat the bulk of the W in both the Carrizo Sand and Aquia aquifersediments occurs within the residual (Fraction 5) and the oxidizable(Fraction 4) phases. Consequently, because the residual phase repre-sents W structurally held within primary and secondary minerals inthese aquifer sediments, the sequential extraction data indicate thatthe chief reservoir of labile W within Carrizo Sand and Aquia aquifersediments are oxidizable phases. These oxidizable phases can includemetal sulfides like pyrite, which has been identified in sedimentsfrom both aquifers, and/or sedimentary organic matter (SOM).

4.4. Geochemical modeling

Results of the aqueous speciation modeling of W in Carrizo Sandand Aquia aquifer groundwaters are presented in Fig. 9 as percenttotal dissolved W vs. distance along each flow path. In both aquifers,the model predicts that tungsten chiefly occurs as WO4

2−, accountingfor between 84.5% and >99% of all W in Carrizo Sand groundwaters,and >99% of W in Aquia groundwaters. The model also predictsthat thiotungstate complexes, primarily the WO3S2− species, maybe important in Carrizo Sand aquifer groundwaters, accounting forbetween ~1% and 14% of W in solution (Fig. 9). Nevertheless, it is im-portant to stress that more accurate predictions of the distribution ofthiotungstate species in Carrizo Sand and Aquia aquifer groundwatersstill awaits determination of the actual formation constants for thesespecies (e.g., Mohajerin and Johannesson, 2012).

Surface complexation modeling of WO42− adsorption onto Carrizo

Sand and Aquia aquifer sediments required additional considerationsto achieve reasonable simulations. Previously, Dzombak and Morel(1990) estimated surface complexation constants for WO4

2− onto hy-drous ferric oxide using linear free-energy relationships between thefirst association constant for WO4

2− (i.e., log Ka,1; Table 2) and intrin-sic surface complexation constants for other anions. More recently,Gustafsson (2003) measured WO4

2− adsorption onto ferrihydrite,

Table 7Concentrations of W in sequential extractions of sediment samples from the CarrizoSand and Aquia aquifers, expressed as pg g−1 of sediment (dry weight)a. Samplesrepresent different depths below ground surface within each aquifer.

Fraction 1 Fraction 2 Fraction 3 Fraction 4 Fraction 5

Carrizo Sand900–930 m 3 0 36 796 580930–960 m 18 1 59 502 951960–990 m 9 0 14 560 776990–1030 m 44 0 57 230 3042

Aquia61 m 49 38 18 1470 165072 m 26 65 24 846 217078 m 35 25 28 216 220086.5 m 39 40 16 1620 228097 m 16 8 0 229 1240105 m 28 2 44 1810 287

a Corresponding values for Fe and Mn can be found in Haque (2007) and Haque et al.(2008).

Fig. 7. Tungsten contents (pg g−1) of the sequential extractions for sediment samplesfrom the Carrizo Sand aquifer. Fractions F1, F2, F3, F4, and F5 refer to the exchangeable,acidic, reducible, oxidizable, and residual fractions, respectively. See text for details.

Fig. 8. Tungsten contents (pg g−1) of the sequential extractions for sediment samplesfrom the Aquia aquifer. Fractions F1, F2, F3, F4, and F5 refer to the exchangeable, acidic,reducible, oxidizable, and residual fractions, respectively. See text for details.

87K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

and subsequently used the non-linear, least-squares optimizationprogram FITEQL 4.0 (Herbelin and Westall, 1999) to obtain intrinsicsurface complexation constants for WO4

2− adsorption onto ferrihydriteusing both the generalized two-layermodel and the charge distributionmultisite model (CD-MUSIC). Between these two seminal studies, threedifferent surface complexation reactions were formulated to describeWO4

2− adsorption onto ferrihydite. These equations are:

≡FeOH0 þWO42− þ 2Hþ þH2O↔≡FeOW OHð Þ5 logK1

int;

ð1Þ

≡FeOH0 þWO42− þHþ↔≡FeWO4

− þH2O logK2int;

ð2Þ

and

≡FeOH0 þWO42−↔≡FeOHWO4

2− logK3int

ð3Þ

where Kiint is the intrinsic surface complexation constant for each reac-

tion. As mentioned above, Dzombak and Morel (1990) used linearfree-energy relationships to estimate K2

int and K3int, and Gustafsson

(2003) was able to achieve satisfactory results with FITEQL 4.0 solvingeither for K2

int and K3int or K1

int and K3int, but not all three reactions simul-

taneously. Gustafsson (2003) concluded that the model employing K1int

and K3int was preferable owing to a small improvement in themodel fits.

Seiler et al (2005) used both of these 2-pK surface complexationmodels to simulate WO4

2− adsorption onto ferrihydrite in groundwatersystems from the Carson Desert. These investigators concluded thatthe model of Gustafsson (2003) produced “…unreasonably high Wadsorption…(greater than 99% at pH 10).” for groundwater from theirstudy site, and consequently chose to employ the model and surfacecomplexation constants estimated by Dzombak and Morel (1990) intheir modeling (Seiler et al., 2005). Application of the K1

int and K3int

model of Gustafsson (2003) to Carrizo Sand and Aquia aquifer ground-waters also appeared to overestimate WO4

2− adsorption, whereas themodel of Dzombak and Morel (1990) underestimated WO4

2− adsorp-tion at our study sites (Dave and Johannesson, 2008a,b).

Consequently, we re-fit the WO42− adsorption data of Gustafsson

(2003) using FITEQL 4.0, the protonation constants for WO42− from

Smith and Martell (2004), and by recasting Eq. (1) above into thefollowing form:

≡FeOH0 þWO42− þ 2Hþ þ↔≡FeHWO4

0 þH2O logK1int

ð4Þ

to be consistent with Eqs. (2) and (3). Because the adsorption exper-iments of Gustafsson (2003) were conducted in a 0.01 M NaNO3

background electrolyte, all equilibrium constants were corrected tothis ionic strength. As an example, the first protonation constant for

b)

Kilometers

10 20 30 40 50 60 70 80 90

Per

cen

t T

un

gst

en

0

20

40

60

80

100

a)

Kilometers0 10 20 30 40 50 60 70

Per

cen

t T

un

gst

en

0

20

40

60

80

100

HWO4-

WO42-

WO3S2-

WO2S22-

WOS32-

WS42-

Fig. 9. Results of the solution complexation modeling of W (percent total W) along theflow paths in (a) the Carrizo Sand aquifer and (b) the Aquia aquifer. The solution com-plexation model predicts that W chiefly occurs as the WO4

2− oxyanion, althoughthiotungstate complexes may occur in groundwaters from the Carrizo Sand aquifer.

Table 8Intrinsic surface complexation constants and associated standard deviations (inparentheses) for WO4

2− adsorption onto ferrihydrite. The surface complexation con-stants represent non-linear least squares fits to the experimental data of Gustafsson(2003) using FITEQL 4.0 (Herbelin and Westall, 1999) and the 2-pK models discussedin the text.

Total Fe (M) log K2int log K3

int WSOS/DF

Tungstate adsorption using reactions 2 and 33 × 10−4 10.28 (0.878) 4.53 (0.074) 131 × 10−3 10.28a 5.85 (0.043) 63 × 10−3 10.28a 6.69 (0.041) 9Weighted meanb 10.28 5.89

Tungstate adsorption using reactions 3 and 43 × 10−4 19.39 (0.067) 6.30 (0.13) 1.91 × 10−3 19.39a 6.72 (0.06) 0.433 × 10−3 19.39a 6.69 (0.04) 9.03Weighted meanb 19.39 6.64

a Held to this value to attain convergence (e.g., Gustafsson, 2003).b Computed using the approach of Dzombak and Morel (1990). See Eqs. (6) and (7)

of the text for details.

88 K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

WO42− (Table 2) was corrected to I = 0.01 M using the following

equation:

KI¼0:01Ma;1 ¼ KI¼0M

a;1 �γHþγWO2−

4

γHWO−4

!ð5Þ

where Ka,1I = 0M is the first protonation constant at infinite dilution, and

γi are the activity coefficients of the relevant species at I = 0.01 M(Wright, 2007).

The best estimates of the surface complexation constants obtainedusing FITEQL 4.0 for WO4

2− adsorption onto ferrihydrite were subse-quently computed using the weighted averaging approach describedin Dzombak and Morel (1990), such that the weighting factor, wi, iscalculated using the expression

wi ¼1=σ logK

� �i

∑ 1=σ logK

� �i

ð6Þ

where (σlogK)i is the standard deviation computed by FITEQL 4.0(i.e., Table 7.5 in the FITEQL output). The best estimate of the intrin-sic surface complexation constant is then computed using

logK ¼ ∑wi logKð Þi ð7Þ

where (log K)i are the different estimates of the surface complexa-tion constant determined for the different sorbate/sorbent ratios(Dzombak and Morel, 1990).

Table 8 presents the results of the refitting exercise of thedata from Gustafsson (2003) for the two different 2-pK models, oneemploying K2

int and K3int and the other using K1

int and K3int. These

models are equivalent to the equations describing WO42− adsorption

from Dzombak and Morel (1990) and Gustafsson (2003), respective-ly, except for the use of Eq. (4) in place of Eq. (1) for the K1

int and K3int

model. The results of the model fits to the data are presented inFig. 10, which demonstrate that the K2

int and K3int model that uses

Eqs. (2) and (3) does a better job at capturing the WO42− adsorption

data of Gustafsson (2003) than the K1int and K3

int model. Moreover,the adsorption edges predicted using the K2

int and K3int model correctly

shift to lower pH values with increasing sorbate (WO42−) to sorbent

(ferrihydrite) ratios, mimicking the adsorption data, whereas the ad-sorption edges predicted using the K1

int and K3int model shift to higher

pH with increasing sorbate/sorbent ratios, which is opposite to thetrends in the actual data. For this reason alone, the K2

int and K3int

model is superior to the K1int and K3

int model, and is consequentlyused hereafter to simulate WO4

2− adsorption onto the Carrizo Sandand Aquia aquifer sediments (Table 3). Nonetheless, it is importantto note that although the K2

int and K3int model does a good job at

capturing WO42− adsorption for the experiments conducted with

1 mmol L−1 of suspended ferrihydrite, it overpredicts WO42− adsorp-

tion for the experiments that employed 3 mmol L−1 ferrihydrite andunderpredicts WO4

2− adsorption, except at high pH, for the experi-ments conducted with 0.3 mmol L−1 ferrihydrite (Fig. 10).

Application of the K2int and K3

int model (Eqs. (2) and (3)) to CarrizoSand and Aquia aquifer groundwaters is presented in Fig. 11. Here,WO4

2− adsorption is considered along with adsorption of silicic acidand the arsenate and arsenite oxyanion species (Table 3), as these an-ions may compete with WO4

2− for surface sites (e.g., Swedlund andWebster, 1999; Gustafsson, 2003; Seiler et al., 2005; Gustafsson andBhattacharya, 2007; Sharif et al., 2011). Symbols represent the pre-dicted amount of WO4

2− adsorption for each groundwater samplecollected from the Carrizo Sand and Aquia aquifers using the datafrom Tables 4 to 6, whereas the dashed lines show the predictedadsorption edges for WO4

2− onto ferrihydrite (Table 3) for a weakelectrolyte solution with an ionic strength equal to the mean ionicstrength of all of the Carrizo Sand and Aquia aquifer groundwaterssampled (i.e., 0.006 M NaCl). As mentioned above, PO4

3− was general-ly below detection (i.e., 0.3–0.4 μmol kg−1) in Carrizo Sand and Aquiaaquifer groundwaters. Nevertheless, we employed a PO4

3− concentra-tion of 0.4 μmol kg−1 in the model simulations (the maximum possi-ble PO4

3− concentrations in these groundwaters) to examine thepossible effects of PO4

3− on WO42− adsorption. Inclusion of PO4

3− inthe model calculations (not shown) resulted in adsorption edgesthat differed by less than 5%, on average, than the adsorption edges

pH

6.0 6.5 7.0 7.5 8.0 8.5 9.0

Fra

ctio

n W

O42-

So

rbed

0.0

0.2

0.4

0.6

0.8

1.0

Carrizo Sand GroundwaterAquia Groundwater0.006 M NaCl, 0.852 g/L Fe0.006 M NaCl, 7.88 g/L Fe

Fig. 11. Results of the K2 and K3 surface complexation model for WO42− adsorption onto

Carrizo Sand and Aquia aquifer sediments (Tables 3, 8). Data points represent modelpredictions for individual groundwater samples from both aquifers employing the geo-chemical data presented in Tables 4–6, and the estimated amounts of available Fe(III)oxide/oxyhydroxide (as ferrihydrite) in sediments from each aquifer (see text fordetails). Dashed lines are modeled WO4

2− adsorption edges for 0.006 M NaCl solutions.

a)Refit using K2 and K3

pH4 5 6 7 8 9 10

Fra

ctio

n W

O42-

So

rbed

0.4

1.0

0.8

0.6

0.2

0.0

3 mM Fe1 mM Fe0.3 mM FeModel - 3 mMModel - 1 mM FeModel - 0.3 mM Fe

b) Refit using K1 and K3

pH4 5 6 7 8 9 10

Fra

ctio

n W

O42-

So

rbed

0.0

0.2

0.4

0.6

0.8

1.0

3 mM Fe1 mM Fe0.3 mM FeModel - 3 mM FeModel - 1 mM FeModel - 0.3 mM Fe

Fig. 10. Results of the refitting exercise for the W adsorption data collected byGustafsson (2003). Data points in both panels represent the results of batch adsorptionexperiments for solutions containing 50 μmol L−1 of WO4

2− and either 0.3, 1, or3 mmol L−1 suspended ferrihydrite (see Gustafsson, 2003). Lines are model fitsusing Eqs. (2) and (3) (panel a; K2 and K3 model) and Eqs. (3) and (4) (panel b;K1 and K3 model) and the corresponding intrinsic surface complexation constantspresented in Table 8.

89K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

calculated without PO43− that are shown in Fig. 11. More specifically,

the position of the WO42− adsorption edge shifted left by less than

0.1 pH unit, on average, in the presence of 0.4 μmol kg−1 PO43−.

These results indicate that the very low PO43− concentrations that

characterize Carrizo Sand and Aquia aquifer groundwaters play aminor competitive role, at most, in influencing WO4

2− adsorption inthese groundwater flow systems.

The model predictions for the weak electrolyte solutions agreewell with the predictions for individual Carrizo Sand and Aquiaaquifer groundwater samples (Fig. 11). The surface complexationmodeling demonstrates that both the pH of the groundwater as wellas the amount of Fe(III) oxide/oxyhydroxide within the aquifer sedi-ments play important roles in controlling the W concentrations ingroundwater.

5. Discussion

5.1. Redox conditions in Carrizo Sand and Aquia aquifer groundwaters

Changes in groundwater redox chemistry along the flow pathsin both the Carrizo Sand and Aquia aquifers have been previouslydiscussed in detail (Haque and Johannesson, 2006; Tang andJohannesson, 2006; Basu et al., 2007; Haque et al., 2008; Willis andJohannesson, 2011; Johannesson and Neumann, 2013). In brief, the

measured redox sensitive parameters indicate zones of Fe(III) reduc-tion along the flow paths in both aquifers as indicated by increases indissolved Fe(II) concentrations (Figs. 5, 6). In addition, a zone of sul-fate reduction occurs further down gradient along the flow path inthe Carrizo Sand aquifer where dissolved S(-II) concentrations in-crease dramatically (Figs. 5, 6). Redox zonation as observed withinthe Carrizo Sand aquifer is a common feature ofmany confined aquifersystems (e.g., Back and Barnes, 1965; Champ et al., 1979; Edmundset al., 1984; Lovley and Goodwin, 1988; Barcelona et al., 1989;Smedley and Edmunds, 2002; Park et al., 2006; Lee et al., 2007). Thedevelopment of such redox zones is thought to reflectmicrobial respira-tion of progressively less energetically favorable terminal electronacceptors (e.g., Fe(III), SO4

2−) during the process of organic carbonremineralization (Lovley and Phillips, 1987; Potsma and Jakobsen,1996; Jakobsen and Postma, 1999; Penny et al., 2003; Park et al., 2006).

In the case of the Carrizo Sand aquifer, the region where the aquiferfirst becomes confined (i.e., ~7.4 km down gradient along the flowpath), corresponds to the initiation of the Fe(III) reduction zone asevidenced by themore than 40-fold increase in dissolved Fe(II) concen-trations (Fig. 5). Specifically, Fe(II) increases from 0.72 μmol kg−1 atwell KS-1 (7.4 km) to a high of 32.7 μmol kg−1 at the G-1 well lessthan 6 km down gradient (Fig. 5). The substantial increase in Fe(II)concentrations strongly suggests that dissimilatory Fe(III) reduction isthe chief redox buffering reaction along this portion of the flow path(Basu et al., 2007). Recent biogeochemical reactive transport modelingsupports this hypothesis (Johannesson and Neumann, 2013). Withflow down gradient beyond the Poteet well (14.5 km), dissolved S(-II)concentrations increase substantially. For example, as groundwatertransits the ~27 km distance from the Poteet well to the Peeler well(41.1 km), dissolved S(-II) increases by a factor of 6 (Table 4, Fig. 5).With further flow down gradient beyond the Peeler well, S(-II) concen-trations increase by close to 15-fold over the next ~25 km. The substan-tial increase in dissolved S(-II) concentrations along the mid- to lowerreaches of the studied flowpath strongly suggests that sulfate reductionbuffers the redox conditions along this portion of the flow path (Haqueand Johannesson, 2006; Basu et al., 2007). The predominance of sulfatereduction along the mid- to lower reaches of the studied flow path ofthe Carrizo Sand aquifer is also supported by biogeochemical reactivetransport modeling (Johannesson and Neumann, 2013).

The subsequent decrease in Fe(II) concentrations in Carrizo Sandgroundwaters with flow beyond the G-1 well can be explained bya number of processes that include: (1) readsorption of Fe(II) onto

90 K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

Fe(III) oxides/oxyhydroxides in the aquifer sediments; (2) precipita-tion of Fe(III) and/or mixed Fe(II)/Fe(III) oxides/oxyhydroxides;(3) precipitation of siderite; and/or (4) precipitation of Fe(II) sulfideminerals (Basu et al., 2007; Haque et al., 2008; Johannesson andNeumann, 2013). Geochemical modeling using the SpecE8 programof the Geochemist's Workbench (version 8.0; Bethke, 2008; Bethkeand Yeakel, 2010) indicates that Carrizo Sand groundwaters areall substantially oversaturated with respect to hematite, magnetite,goethite, and pyrite (Fig. 12). For example, the mean (± SD) satura-tion indexes for hematite, magnetite, goethite, and pyrite for CarrizoSand groundwaters are 12.3 ± 1.26, 10.9 ± 2.6, 5.66 ± 0.63, and5.82 ± 1, respectively. Moreover, Fig. 12 shows that all of the CarrizoSand groundwaters are close to equilibrium (i.e., slightly oversaturated)with respect to ferrihydrite [mean (± SD) SIferrihydrite = 1.27 ± 0.63],and groundwaters from the mid- to lower reaches of the studiedflow path are saturated to slightly oversaturated with respect tomackinawite (FeS) and siderite (FeCO3).

For the Aquia aquifer, Fe(III) predominates in groundwaters alongthe first 52 km of the studied flow path, suggesting possible transportof colloidal iron in the aquifer (Haque et al., 2008). However, withflow down gradient beyond 52 km, Fe(II) is the chief form of ironin Aquia aquifer groundwaters (Fig. 6). The switch from Fe(III)

Kilometers

0 10 20 30 40 50 60 70

Sat

ura

tio

n In

dex

-4

-2

0

2

4

6

8

GoethiteFerrihydrite/HFOPyriteMackinawiteSiderite

Kilometers20 30 40 50 60 70 80 90

Sat

ura

tio

n In

dex

-4

-2

0

2

4

6

8

a)

b)

Ferrihydrite/HFO

Goethite

Pyrite

MackinawiteSiderite

Ferrihydrite/HFO

Goethite

Pyrite

Mackinawite

Siderite

Fig. 12. Trends in saturation indices for selected Fe minerals along the flow pathsin (a) the Carrizo Sand aquifer, and (b) the Aquia aquifer. Groundwaters from bothaquifers are highly oversaturated with respect to goethite and pyrite, and slightlyoversaturated with respect to ferrihydrite. Groundwaters from the lower reaches ofthe Carrizo Sand aquifer are close to saturation with respect to both siderite andmackinawite (i.e., FeS; Rickard et al., 2006; Rickard and Luther, 2007). The majorityof Aquia aquifer groundwaters are in equilibrium with siderite and many are alsosaturated or close to saturation with respect to mackinawite.

to Fe(II) is consistent with Fe(III) reduction buffering the redox con-ditions in Aquia groundwaters with flow beyond 52 km. Further-more, the relatively high NH4

+ and low NO3− concentrations (Fig. 6)

indicate that post dissimilatory nitrate reducing conditions predomi-nate in the Aquia aquifer, which are consistent with the occurrenceof Fe(III) reducing conditions (Burdige, 2006; Haque et al., 2008). Incontrast to the Carrizo Sand aquifer, dissolved S(-II) concentrationsare at or below the limit of detection in Aquia aquifer groundwaters(Table 5, Fig. 6). It is important to note, however, that the absenceof dissolved S(-II) does not rule out the possibility that sulfate reduc-tion is occurring in the Aquia aquifer as the relatively high Fe(II)concentrations of these groundwaters could readily remove S(-II)from solution via precipitation of iron sulfide minerals (Postma andJakobsen, 1996; Jakobsen and Postma, 1999; Bethke et al., 2008). In-deed, all of the Aquia aquifer groundwaters with measureable S(-II)are oversaturated with respect to pyrite [mean (± SD) SIpyrite =4.22 ± 1.02; Fig. 12]. Furthermore, Aquia groundwaters down gradi-ent of where the switch from Fe(III) to Fe(II) occurs are all close toequilibrium with respect to mackinawite (Fig. 12). Hence, sulfate re-duction is probably important within the Aquia aquifer despite thelow dissolved S(-II) concentrations. Nonetheless, the higher dissolvedoxygen concentrations and higher Eh values of Aquia aquifer ground-waters compared to those of the Carrizo Sand aquifer (Figs. 5, 6)generally support Fe(III) reduction as the principal redox bufferingreaction in the Aquia aquifer.

5.2. Tungsten in Carrizo Sand and Aquia aquifer groundwaters

Seiler et al. (2005) evaluated a number of processes that maycontrol dissolved W concentrations in groundwaters from the CarsonDesert of northwestern Nevada. These include: (1) evaporative con-centration of local river waters followed by recharge of the resultinghigh W waters to underlying aquifers; (2) dissolution of W-bearingminerals within the aquifer sediments; (3) mixing with high W geo-thermal waters; (4) reductive dissolution of Fe(III)/Mn(IV) oxides/oxyhydroxides within the aquifer sediments and subsequent releaseof sorbed and/or co-precipitated W to the groundwaters; and (5)pH-related desorption of W from mineral surfaces. Evaporativeconcentration was previously proposed to explain the elevated Wconcentrations of surface waters and some groundwaters from theCarson Desert (Johannesson et al., 2000). However, this mechanismcannot account for all of the high W groundwaters in the region be-cause many of these waters exhibit relatively low Cl concentrations(Seiler et al., 2005). Dissolution of W-bearing minerals and mixingwith W-rich hydrothermal fluids are likely to be locally importantowing to the proximity of tungsten mineralization and geothermalzones in northwestern Nevada and eastern California (Seiler et al.,2005). Reductive dissolution of Fe(III)/Mn(IV) oxides/oxyhydroxidesfollowed by the release of sorbed and/or co-precipitated W is alsolikely to be important, however, direct evidence for this processappears to be lacking for aquifers of the Carson Desert (Seileret al., 2005). Tungsten can still be mobilized from Fe(III)/Mn(IV)oxides/oxyhydroxides without dissolution of these mineral phasesby desorption reactions induced by rising pH (e.g., Miyauchi et al.,1998; Harita et al., 2005; Seiler et al., 2005). The fact that CarsonDesert groundwaters with the highest reported W concentrations(i.e., W > 272 nmol kg−1) all have pH ≥ 8 suggests that pH-relateddesorption is an important mechanism by which W is mobilized inthese aquifers (Seiler et al., 2005; Cutler, 2011). In addition, oncemobilized, pH-related adsorption/desorption reactions can facilitateW transport in neutral to high pH groundwaters, such as those ofthe Carson Desert. These arguments and observations are broadlyconsistent with measured desorption edges for WO4

2− (Gustafsson,2003; Fig. 10), and the reported conservative behavior of W in alka-line groundwaters from a shallow, unconsolidated and fracturedtuffaceous aquifer in southern Nevada (Johannesson and Tang, 2009).

91K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

For the Carrizo Sand and Aquia aquifers, no evidence exists im-plicating evaporative concentration, groundwater interaction withW-rich minerals, or mixing with W-rich geothermal fluids as viableprocesses influencing the dissolved W concentrations. Furthermore,the weak and statistically insignificant relationships between W andFe(II) concentrations in groundwaters from both aquifers suggeststhat, although likely important locally (see below), reductive dissolu-tion of Fe(III) oxides/oxyhydroxides and subsequent release of sorbedand/or coprecipitated W is not the principal mechanism of W mobili-zation in either aquifer. The weak inverse relationship betweenW andFe(II) in Carrizo Sand aquifer groundwaters reflects the large-scaledecoupling of W and Fe(II) concentrations within the region of theaquifer where sulfate reduction predominates (Fig. 5). Specifically,the substantial increase in W concentrations in groundwaters fromthe sulfate reduction zone along the lower reaches of the flow pathcorresponds to decreasing Fe(II) concentrations (Fig. 5). Nevertheless,a closer inspection of the region of the Carrizo Sand aquifer whereFe(III) reduction predominates (i.e., from the KS-1 well to the Poteetwell; Johannesson and Neumann, 2013), reveals a positive relation-ship between dissolved W and Fe(II) concentrations (i.e., r = 0.56).Hence, the small but attendant increase in W with increasing Fe(II)concentrations in groundwaters from the Fe(III) reducing zone indi-cates that reductive dissolution of Fe(III) oxides/oxyhydroxides doesplay at least a minor role in W mobilization within the Carrizo Sandaquifer. Tungsten and Fe(II) also exhibit similar trends in concentra-tion variations along the first ~52 km of the studied flow path withinthe Aquia aquifer (Fig. 6), which could also indicate a relationship be-tween W and Fe(II) release to these groundwaters. Unfortunately,correspondence between W and Fe(II) concentrations is lacking forgroundwaters collected along the lower half of the studied flow path(Fig. 6), explaining the overall weak statistical relationships betweenW and Fe(II) in Aquia aquifer groundwaters.

The apparentminor role that Fe(III) oxide/oxyhydroxide reductionplays in mobilizing W in both the Carrizo Sand and Aquia aquifers is,however, consistent with the surface complexation model (Fig. 11).Specifically, the model predicts that W released to groundwaters byreductive dissolution of Fe(III) oxides/oxyhydroxides will be stronglyre-adsorbed onto Fe(III) oxides/oxyhydroxides unless the groundwaterpH exceeds ~8, or all Fe(III) oxides/oxyhydroxides are stripped from theaquifer sediments by quantitative reductive dissolution. Groundwatersfrom the region of the Carrizo Sand aquifer where Fe(III) reductionis the chief redox buffering reaction all exhibit pH values less than7 (compare Figs. 3 and 5). Hence, the surface complexation modelindicates that any W mobilized during reductive dissolution of Fe(III)oxides/oxyhydroxides at these pH values will be strongly re-adsorbedonto remaining Fe(III) oxides/oxyhydroxides in Carrizo Sand aquifersediments (Fig. 11). In the Aquia aquifer, because groundwater pH iscommonly ~8 [i.e., mean (±SD) pH = 7.97 ± 0.24], any Fe(III) reduc-tion within the aquifer must occur at higher pH values than in theCarrizo Sand aquifer (e.g., 7.6 ≤ pH ≤ 8.0; compare Figs. 4 and 6).Hence, it is reasonable to expect observable W mobilization as a resultof reductive dissolution of Fe(III) oxides/oxyhydroxides in the Aquiaaquifer. However, the surface complexation model indicates thatthe substantially greater Fe(III) oxide/oxyhydroxide content of Aquiaaquifer sediments compared to Carrizo Sand aquifer sediments(i.e., 7.88 g L−1 vs. 0.852 g L−1 for the Aquia and Carrizo Sand aquifersediments, respectively; Fig. 11), essentially buffers the groundwaterW concentrations at lower levels owing to strong surface complexationonto the abundant Fe(III) oxides/oxyhydroxides within the Aquiaaquifer (Fig. 11).

The strong and statistically significant relationship betweenW con-centrations and pH in Carrizo Sand aquifer groundwaters (r = 0.78,p b 0.001) indicates that pH-related desorption is the chief mechanismby which W is mobilized in the aquifer system. These relationships arealso consistent with the surface complexation model (Fig. 11). Moreexplicitly, as pH increases from ~6.2 to >8.5 along the flow path in

the Carrizo Sand aquifer, W concentrations exhibit a ~350-fold increase(Fig. 5). Similar increases in dissolved W concentrations in response toincreasing pH have been reported in lake waters (e.g., Miyauchi et al.,1998; Harita et al., 2005). In contrast, the weak, inverse relationshipbetween W and pH in the Aquia aquifer groundwaters (r = −0.4,p b 0.02) suggests that adsorption/desorption may be less criticalfor W mobilization within the Aquia aquifer. However, as notedabove, the surface complexation model demonstrates that althoughpH-related W desorption is important in groundwater flow systems,the amount of available Fe(III) oxide/oxyhydroxide substrate withinthe aquifer sediments is also central to controlling W concentrationsin groundwaters (Fig. 11). Specifically, aquifers with lower contents ofFe(III) oxides/oxyhydroxides (e.g., Carrizo Sand aquifer) are more sus-ceptible to pH-related W desorption at moderately alkaline pH valuesthan aquifers with higher contents of Fe(III) oxides/oxyhydroxides,such as the Aquia aquifer, where W mobilized by reductive dissolutionis subsequently immobilized by re-adsorption onto mineral surfacesites (Fig. 11). Indeed, the surface complexation model predicts thatgroundwater pH values close to 9 are required to achieve similar frac-tions of W desorption in the Aquia aquifer as observed in the CarrizoSand aquifer (Fig. 11).

Therefore, although a number of investigations support thatWO4

2− is mobile in oxic, high pH natural waters (e.g., Johannessonet al., 2000; Gustafsson, 2003; Seiler et al., 2005; Bednar et al., 2008;Johannesson and Tang, 2009; Cutler, 2011), the data and geochemicalmodeling presented here also demonstrate that WO4

2− can be mobilein anoxic groundwaters (Figs. 5, 9). Indeed, W concentrations inCarrizo Sand groundwaters increase dramatically between the regionof the aquifer where Fe(III) reduction is supplanted by sulfate reduc-tion as the primary redox buffering reaction (Fig. 5). The relativelylow Fe(III) oxide/oxyhydroxide contents of Carrizo Sand aquifersediments, compared, for example, to Aquia aquifer sediments, in con-junction with the high S(-II) concentrations attained in groundwatersfrom down gradient regions of the aquifer, suggests that the bulkof the Fe(III) oxides/oxyhydroxides within the Carrizo Sand are con-sumed by reductive dissolution along the flow path and subsequentlyconverted to siderite and/or Fe sulfideminerals (e.g., Johannesson andNeumann, 2013). Such quantitative reductive dissolution of Fe(III)oxides/oxyhydroxides in the aquifer sediments effectively eliminatesthe chief mineral scavenger of WO4

2− (e.g., Koschinsky and Hein,2003; Kashiwabara et al., 2013) in the Carrizo Sand aquifer, allowingtheWO4

2− concentration to build up in solutionwithin the groundwa-ters along the lower reaches of the studied flowpath. These argumentsare complicated by the fact that the sulfidic waters of the Carrizo Sandaquifer exhibit pH values greater than 8, and consequently, the highWconcentrations of these waters also reflect pH-related desorption(Fig. 11). In contrast, themuch higher Fe(III) oxide/oxyhydroxide con-tents of Aquia aquifer sediments and the non detectable S(-II) concen-trations of the associated groundwaters suggest that Fe(III) oxides/oxyhydroxides within the aquifer sediments are sufficiently abundantthat reductive dissolution of these oxides/oxyhydroxides continuesto produce enough dissolved Fe(II) to scavenge, by precipitationof Fe sulfide minerals, any S(-II) produced by sulfate reduction inthe Aquia aquifer system. Consequently, the higher Fe(III) oxide/oxyhydroxide contents of Aquia aquifer sediments effectively bufferthe WO4

2− concentrations at much lower levels than observed in theCarrizo Sand groundwaters.

The strong, positive correlation between W and S(-II) in ground-waters from the Carrizo Sand aquifer (i.e., r = 0.95, p ≤ 0.001) isintriguing and may indicate that thiotungstate complexes form inthe anoxic (i.e., sulfidic) groundwaters along the lower reaches of thestudied flow path (Dave and Johannesson, 2008a,b). Preliminary solu-tion complexation modeling further supports this conjecture (Fig. 9),although, again, caution is warranted here because we have employedthe formation constants for analogous thiomolybdate species as proxies(see above). By comparison, W concentrations in Aquia groundwaters

92 K.H. Johannesson et al. / Chemical Geology 351 (2013) 76–94

are generally lower than in the Carrizo Sand aquifer [e.g., the mean(± SD) W concentrations in Aquia and Carrizo Sand groundwatersare 63.9 ± 59 pmol kg−1 and 245 ± 442 pmol kg−1, respectively],and more importantly, Aquia aquifer groundwaters do not attainsulfidic conditions (Figs. 5, 6). Consequently, the preliminary solutioncomplexation modeling predicts that thiotungstate complexes arenegligible in Aquia aquifer groundwaters (Fig. 9). To the best of ourknowledge, the formation of thiotungstate complexes has not beeninvestigated in groundwaters or other low-temperature, naturalwaters despite the recognition that thiomolybdate complexes can beimportant in such environments (e.g., Helz et al., 1996; Ericksonand Helz, 2000; Helz et al., 2011). It is important to note, however,that if thiotungstates also form in sulfidic waters, analogy withthiomolybdates suggests that they may likewise be scavenged ontothe surfaces of Fe sulfide minerals such as pyrite and/or mackinawite(e.g., Bostick et al., 2003; Vorlicek et al., 2004), and consequentlywould be removed to the aquifer sediments. More intensive inves-tigation of thiotungstate interactions with Fe oxides and Fe sulfideminerals is required, however, to quantitatively evaluate the mineralsurface reactivity of thiotungstate anions (e.g., Bostick et al., 2003;Vorlicek et al., 2004; Xu et al., 2006).

5.3. Tungsten speciation in Carrizo Sand and Aquia aquifer sediments

An interesting feature of the sequential extraction data for theCarrizo Sand and Aquia aquifer sediments is that besides the “residual”fraction, the “oxidizable” fraction represents the most important reser-voir of W in these sediments. The “oxidizable” fraction presumablyrepresents metal sulfides or possibly SOM, whereas the “residual”fraction is the portion of W that is “locked” in the crystalline structureof primary and secondaryminerals within aquifer sediments. The resid-ual fraction is not considered labile in either of these groundwaterflow systems. Consequently, the bulk of the labile, solid phase W inboth aquifers appears to be associated with either metal sulfides orSOM. Specifically how these sequential extraction results correspondwith the trends in W concentrations within Carrizo Sand and Aquiaaquifer groundwaters requires additional investigation. One possibility,however, is that iron sulfide minerals within the aquifer sedimentsmay also represent an important sink for W (e.g., Bostick et al.,2003; Vorlicek et al., 2004). We note that the concentrations of W inCarrizo Sand groundwaters are strongly correlated with the computedsaturation indexes for pyrite (r = 0.93, p b 0.001) and mackinawite(r = 0.81, p b 0.005). Similar relationships are not observed, however,in the Aquia aquifer. Finally, although bulk EXAFS study of W in CarrizoSand and Aquia aquifer sediments is underway (Datta and Johannesson,in prep.), these investigations have been hampered by the low W con-tent of the aquifer sediments. Consequently, more spectroscopic dataare necessary before conclusions regarding the solid phase speciationof W can be confidently made for either aquifer.

6. Conclusions

Tungsten concentrations in groundwaters are strongly controlled bythe groundwater pH and the Fe(III) oxide/oxyhydroxide content ofthe aquifer sediments. Specifically, increasing pH and decreasing con-tents of Fe(III) oxides/oxyhydroxides in aquifer sediments both favorW mobilization. The data presented here indicates that W is initiallymobilized by reductive dissolution of Fe(III) oxides/oxyhydroxides insuboxic groundwaters, but that its subsequent transport with flowinggroundwater is hindered by strong re-adsorption onto Fe(III) oxides/oxyhydroxides and other mineral surfaces in neutral pH groundwaters.As groundwater pH rises owing to calcite dissolution, silicate mineralweathering, and/or alkalinity production during microbial respiration,W is mobilized by pH-related desorption as WO4

2−. As pH continuesto rise, the affinity of themobilizedWO4

2− for mineral surface sites pro-gressively decreases, allowingW concentrations to increase in solution.

As redox conditions in groundwaters are progressively lowered owingto microbial respiratory processes, some of the mobilized W may reactwith dissolved S(-II), forming thiotungstate species, themineral surfacereactivity of which remains to be elucidated. Our results demonstratethat W can be mobile as WO4

2− even in anoxic groundwaters. Hence,the paradigm that oxyanion-forming trace elements are only mobilein high pH, oxic waters is clearly in need of revision.

Acknowledgments

This work was supported by NSF awards EAR-0510697,EAR-0805332, and EAR-1014946 to K. Johannesson and EAR-1014971to S. Datta. We are thankful to Larry Akers, John West, and CandiGonzales of the Evergreen Underground Water Conservation Districtin Pleasanton, Texas, David Bolton of the Maryland Geological Survey,and John Nickerson of the Maryland Department of the Environmentfor field sampling assistance. Drs. Jianwu Tang (Tulane University)and Shama Haque (University of British Columbia), and Dr. StephanieWillis (The University of Texas at Arlington) assisted with samplingand analysis. The comments of two anonymous reviewers and thoseof the editor, Dr. Carla M. Koretsky, improved this paper.

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