Controls on tungsten concentrations in groundwater flow systems: The role of adsorption, aquifer sediment Fe(III) oxide/oxyhydroxide content, and thiotungstate formation
Introduction
A number of recently recognized clusters of childhood leukemia, including clusters in Fallon, Nevada, Sierra Vista, Arizona, and Elk Grove, California, are tentatively linked to the proximity of these sites to tungsten (W)-bearing ore deposits, associated active or inactive smelting operations, and/or hard-metal processing facilities (Seiler et al., 2005, Koutsospyros et al., 2006, Sheppard et al., 2006, Sheppard et al., 2007, Bednar et al., 2007, Bednar et al., 2008). Indeed, the childhood leukemia cluster afflicting Fallon, Nevada, spawned an investigation by the United States Centers for Disease Control and Prevention (CDC), which revealed through urine analysis that residents were exposed to elevated levels of W (Seiler et al., 2005, Koutsospyros et al., 2006). Although it is inherently difficult to directly link these childhood leukemia clusters to environmental exposure to elevated W concentrations, either via inhalation or through consumption of drinking waters with high W concentrations, many studies have shown that W can be toxic and may indeed be carcinogenic (Peão et al., 1993, Marquet et al., 1996, ATSDR (Agency for Toxic Substances, Disease Registry), 2005, Kalinich et al., 2005, U.S. EPA, 2008, Kelly et al., 2013). Moreover, inhalation of W dust in the hard metal manufacturing industry is linked to pulmonary fibrosis and hard metal pneumoconiosis (Edel et al., 1990, Sprince et al., 1994, Sahle et al., 1996). Furthermore, polytungstates like sodium metatungstate (3Na2WO4·9WO3) appear to be more toxic than monomeric tungstate (e.g., Na2WO4; Strigul et al., 2009, Strigul et al., 2010), and may thus pose problems for people consuming drinking water with elevated W concentrations. The CDC investigation led to a study by the U.S. Geological Survey of W in groundwaters of the Carson Desert region of northwest Nevada, which showed elevated W concentrations ranging from 1.47 to 4036 nmol kg− 1 (0.27–742 μg kg− 1; Seiler et al., 2005). These concentrations are high and similar to W concentrations previously reported for groundwaters of the Carson Desert (1029 nmol kg− 1; Johannesson et al., 2000), and may in part be responsible for the high levels of W found in local residents.
In addition to natural sources of W to the environment, interest in the environmental geochemistry of W is also on the rise owing to its increasing use as a replacement for lead (Pb) ammunition for hunting and recreational shooting, for fishing weights, and for military ammunition (i.e., high kinetic energy penetrators and small arms ammunition; Strigul et al., 2005, Koutsospyros et al., 2006, Bednar et al., 2007, Clausen et al., 2007, Bednar et al., 2008, Clausen and Korte, 2009). Here, the use of W was originally conceived as a way to limit the addition of toxic Pb to the environment by what was thought to be a non-toxic, inert metal of low environmental mobility (Koutsospyros et al., 2006, Bednar et al., 2007). In fact, the U.S. Army's “Green Armament Technology” (GAT) program initially sought to limit environmental pollution by recommending the replacement of 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 that W can be toxic and carcinogenic (e.g., Marquet et al., 1996, Sahle et al., 1996, Kalinich et al., 2005, Steinburg et al., 2007, Strigul et al., 2009, Strigul, 2010, Strigul et al., 2010, Kelly et al., 2013), and field-based investigations indicate that it is mobile in the environment (e.g., Johannesson et al., 2000, Dermatas et al., 2004, Petrunic and Al, 2005, Seiler et al., 2005, Dermatas et al., 2006, Arnórsson and Óskarsson, 2007, Clausen et al., 2007, Dave and Johannesson, 2008a, Dave and Johannesson, 2008b, Bednar et al., 2009, Clausen and Korte, 2009, Johannesson and Tang, 2009).
In general, W exhibits similar geochemical behavior to other oxyanion-forming trace elements such as molybdenum (Mo) and arsenic (As) in the environment in that it is strongly sorbed to Fe(III) oxides/oxyhydroxides and other mineral surfaces in oxic waters of low to circumneutral pH, and desorbs from these mineral surface sites at higher pH (Johannesson et al., 2000, Gustafsson, 2003, Koschinsky and Hein, 2003, Seiler et al., 2005, Arnórsson and Óskarsson, 2007, Bednar et al., 2007, Bednar et al., 2008, Bednar et al., 2009, Johannesson and Tang, 2009, Kashiwabara et al., 2013). At low pH (pH < 5) and high dissolved W concentrations (W ≥ 10− 2 mol kg− 1), W tends to form polymerized species in solution, whereas monomers such as the tungstate oxyanion, i.e., WO42 −, predominate at low W concentrations, neutral to alkaline pH, high temperatures, and high ionic strength (Wesolowski et al., 1984, Wood, 1992, Bednar et al., 2007, Rodríguez-Fortea et al., 2008). Despite these observations, little is actually known about the biogeochemistry of W in low-temperature, groundwater flow systems, which are important drinking water sources for much of the world's population (Solley et al., 1993, Clarke and King, 2004). To the best of our knowledge, there are no published studies that have investigated how biogeochemical reactions occurring along groundwater flow paths, including geochemical reactions that alter the solution composition and redox conditions, affect W concentrations and speciation within aquifer systems. For example, although WO42 − is thought to predominate in natural waters, the similarity between W and Mo suggests that thiotungstate species may form in some anoxic, sulfidic waters (e.g., Erickson and Helz, 2000, Dave and Johannesson, 2008a, Dave and Johannesson, 2008b, Mohajerin and Johannesson, 2012).
In this contribution we present W concentrations for groundwaters collected along flow paths in two well-characterized aquifers, the Carrizo Sand (Texas) and the Aquia (Maryland) aquifers. Groundwater W concentrations are evaluated along the studied flow paths in each aquifer along with major solute concentrations, pH, and a number of redox sensitive indicators (i.e., Eh, dissolved oxygen, iron species, and sulfide) to develop a better understanding of the biogeochemical processes that control W concentrations in groundwater flow systems that have not been contaminated by anthropogenic activities.
Section snippets
Carrizo Sand aquifer
The portion of the Carrizo Sand aquifer studied is located in Atascosa and McMullin Counties in southeastern Texas (Fig. 1). Water abstracted from the aquifer is used for both irrigation and water supply (i.e., drinking water) purposes in the study region (e.g., Hamlin, 1988). Within the study area, the Carrizo Sand aquifer is exposed and recharged by precipitation in northwest Atascosa County, where it defines the Carrizo Sand ridge (Fig. 1). From the recharge area groundwater flows southeast
Field sample collection
All sample bottles (HDPE), Teflon® tubing used during sample collection and filtration, and HDPE and Teflon® labware were cleaned prior to use via standard trace element cleaning procedures (e.g., see Johannesson et al., 2004). Groundwater samples for W analysis were collected (March 2006) from 10 different wells finished within the Carrizo Sand aquifer and from 12 wells completed within the Aquia (August 2006) aquifer (Table 1; Fig. 1, Fig. 2). All of the wells sampled from the Carrizo Sand
General geochemistry
Field parameters, major solute concentrations, and redox indicators measured in groundwaters from the Carrizo Sand and Aquia aquifers are presented in Table 4, Table 5, respectively. Variations in the measured field parameters (i.e., pH, conductivity, and temperature are plotted as functions of distance along the flow paths in the Carrizo Sand 2003, 2004, and 2006 field campaigns) and Aquia (2006 field campaign) aquifers in Fig. 3, Fig. 4, respectively. The data for the Carrizo Sand aquifer
Redox conditions in Carrizo Sand and Aquia aquifer groundwaters
Changes in groundwater redox chemistry along the flow paths in both the Carrizo Sand and Aquia aquifers have been previously discussed in detail (Haque and Johannesson, 2006, Tang and Johannesson, 2006, Basu et al., 2007, Haque et al., 2008, Willis and Johannesson, 2011, Johannesson and Neumann, 2013). In brief, the measured redox sensitive parameters indicate zones of Fe(III) reduction along the flow paths in both aquifers as indicated by increases in dissolved Fe(II) concentrations (Fig. 5,
Conclusions
Tungsten concentrations in groundwaters are strongly controlled by the groundwater pH and the Fe(III) oxide/oxyhydroxide content of the aquifer sediments. Specifically, increasing pH and decreasing contents of Fe(III) oxides/oxyhydroxides in aquifer sediments both favor W mobilization. The data presented here indicates that W is initially mobilized by reductive dissolution of Fe(III) oxides/oxyhydroxides in suboxic groundwaters, but that its subsequent transport with flowing groundwater is
Acknowledgments
This work was supported by NSF awards EAR-0510697, EAR-0805332, and EAR-1014946 to K. Johannesson and EAR-1014971 to S. Datta. We are thankful to Larry Akers, John West, and Candi Gonzales of the Evergreen Underground Water Conservation District in Pleasanton, Texas, David Bolton of the Maryland Geological Survey, and John Nickerson of the Maryland Department of the Environment for field sampling assistance. Drs. Jianwu Tang (Tulane University) and Shama Haque (University of British Columbia),
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