Chromium(VI) generation in vadose zone soils and alluvial sediments of the southwestern Sacramento Valley, California: A potential source of geogenic Cr(VI) to groundwater

Applied Geochemistry 26 (2011) 1488–1501

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Chromium(VI) generation in vadose zone soils and alluvial sediments of the southwestern Sacramento Valley, California: A potential source of geogenic Cr(VI) to groundwater Christopher T. Mills ⇑, Jean M. Morrison, Martin B. Goldhaber, Karl J. Ellefsen United States Geological Survey, Crustal Geophysics & Geochemistry Science Center, Denver Federal Center, MS 964D, Denver, CO 80225, USA

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Article history: Received 19 October 2010 Accepted 28 May 2011 Available online 2 July 2011 Editorial handling by C. Reimann

a b s t r a c t Concentrations of geogenic Cr(VI) in groundwater that exceed the World Health Organization’s maximum contaminant level for drinking water (50 lg L1) occur in several locations globally. The major mechanism for mobilization of this Cr(VI) at these sites is the weathering of Cr(III) from ultramafic rocks and its subsequent oxidation on Mn oxides. This process may be occurring in the southern Sacramento Valley of California where Cr(VI) concentrations in groundwater can approach or exceed 50 lg L1. To characterize Cr geochemistry in the area, samples from several soil auger cores (approximately 4 m deep) and drill cores (approximately 25 m deep) were analyzed for total concentrations of 44 major, minor and trace elements, Cr associated with labile Mn and Fe oxides, and Cr(VI). Total concentrations of Cr in these samples ranged from 140 to 2220 mg per kg soil. Between 9 and 70 mg per kg soil was released by selective extractions that target Fe oxides, but essentially no Cr was associated with the abundant reactive Mn oxides (up to 1000 mg hydroxylamine-reducible Mn per kg soil was present). Both borehole magnetic susceptibility surveys performed at some of the drill core sites and relative differences between Cr released in a 4-acid digestion versus total Cr (lithium metaborate fusion digestion) suggest that the majority of total Cr in the samples is present in refractory chromite minerals transported from ultramafic exposures in the Coast Range Mountains. Chromium(VI) in the samples studied ranged from 0 to 42 lg kg1, representing a minute fraction of total Cr. Chromium(VI) content was typically below detection in surface soils (top 10 cm) where soil organic matter was high, and increased with increasing depth in the soil auger cores as organic matter decreased. Maximum concentrations of Cr(VI) were up to 3 times greater in the deeper drill core samples than the shallow auger cores. Although Cr(VI) in these vadose zone soils and sediments was only a very small fraction of the total solid phase Cr, they are a potentially important source for Cr(VI) to groundwater. Enhanced groundwater recharge through the vadose zone due to irrigation could carry Cr(VI) from the vadose zone to the groundwater and may be the mechanism responsible for the correlation observed between elevated Cr(VI) and NO 3 concentrations in previously published data for valley groundwaters. Incubation of a valley subsoil showed a Cr(VI) production rate of 24 lg kg1 a1 suggesting that field Cr(VI) concentrations could be regenerated annually. Increased Cr(VI) production rates in H+-amended soil incubations indicate that soil acidification processes such as nitrification of ammonium in fertilizers could potentially increase the occurrence of geogenic Cr(VI) in groundwater. Thus, despite the natural origin of the Cr, Cr(VI) generation in the Sacramento Valley soils and sediments has the potential to be influenced by human activities. Published by Elsevier Ltd.

1. Introduction 1.1. Overview Chromium is a redox active element under typical soil and groundwater conditions and is present in two oxidation states, Cr(III) and Cr(VI) (Bartlett and James, 1979; Ball and Nordstrom, ⇑ Corresponding author. Tel.: +1 303 236 5529. E-mail address: [email protected] (C.T. Mills). 0883-2927/$ - see front matter Published by Elsevier Ltd. doi:10.1016/j.apgeochem.2011.05.023

1998). Chromium(III) is a micronutrient with relatively low toxicity. In contrast, Cr(VI) is a known carcinogen and irritant through inhalation and a potential health hazard through ingestion (Salnikow and Zhitkovich, 2007; Stout et al., 2008). The US Environmental Protection Agency (EPA) maximum contaminant level (MCL) for total Cr in drinking water is 100 lg L1 although individual states may set lower standards. The California MCL of 50 lg L1 is also the MCL recommended by the World Health Organization. Chromium can be introduced into the environment through anthropogenic activities such as industrial releases or the use of

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phosphate fertilizers (Molina et al., 2009). Alternatively, elevated concentrations of Cr in soils and water can be a result of natural processes, most prominently, the alteration and weathering of ultramafic rocks (Robles-Camacho and Armienta, 2000; Chung et al., 2001; Cooper, 2002; Fantoni et al., 2002; Ball and Izbicki, 2004; Gonzalez et al., 2005; Garnier et al., 2006; Oze et al., 2007; Izbicki et al., 2008; Wood et al., 2010). The Pacific coast of the United States has numerous exposures of ultramafic rocks (Oze et al., 2007; Krevor et al., 2009) including extensive outcrops in the Coast Range, Sierra Nevada, and Klamath Mountains that border the Sacramento Valley (Goldhaber et al., 2009; Morrison et al., 2009). A geochemical survey of surface soils in the Sacramento Valley showed that concentrations of Cr were up to 100 times greater than the geometric mean (13 mg kg1) for the conterminous USA, likely due to the erosion and transport to the valley of sediments derived from ultramafic rocks in the surrounding mountains (Morrison et al., 2009). Chromium(VI) can be reduced by a host of abiotic and microbially-mediated reactions with reduced compounds including organic C, sulfide, and Fe(II) (Bartlett and Kimble, 1976; Wittbrodt and Palmer, 1995; White and Peterson, 1996; Wielinga et al., 2001;Tseng and Bielefeldt, 2002). However, the only recognized pathway for the conversion of Cr(III) to Cr(VI) at circumneutral pH in natural environments is its oxidation by Mn(III, IV) oxides (Bartlett and James, 1979; Eary and Rai, 1987; Fendorf and Zasoski, 1992). The thermodynamic favorability of Cr(III) oxidation increases with pH, but the low solubility of Cr(III) at circumneutral pH kinetically limits its transport to the surfaces of Mn oxides (Bartlett, 1991; Fendorf and Zasoski, 1992; Ball and Nordstrom, 1998; Dai et al., 2009). This kinetic limitation is likely greater for Cr derived from ultramafic rocks which occurs predominantly as highly insoluble chromite (FeCr2O4) (Oze et al., 2007). Oze et al. (2007) observed that the rate of Cr(VI) generation in a suspension of chromite and Mn oxide particles was approximately 6 times faster at pH 5.0 than at pH 6.7 and not detectable at pH 8.0. Thus, acid-producing processes in soils that contain both geogenic Cr(III) and Mn oxides could have a substantial effect on Cr(VI) production. Elevated concentrations of Cr(VI) in groundwater have been attributed to geogenic sources in several areas of California (Deverel et al., 1984; Ball and Izbicki, 2004; Gonzalez et al., 2005; Oze et al., 2007; Izbicki et al., 2008; Ndung’u et al., 2010) including the southern Sacramento Valley where concentrations of Cr in some groundwaters have been reported to approach or exceed 50 lg L1 (Chung et al., 2001; Luhdorff and Sclamanini, 2004; Dawson et al., 2008; Morrison et al., 2009). Although these studies have suggested that Cr(VI) is produced when naturally-occurring Cr(III) is oxidized by Mn oxides, little work has been performed to investigate the geochemical controls on this mechanism. Chung et al. (2001) incubated drill core material from the Sacramento Valley in the laboratory to show that Cr(VI) is generated in situ from naturally occurring minerals. However, freeze-dried samples collected from a small area with a history of anthropogenic perturbation were used for these experiments. The current study sought to develop a more thorough understanding of processes by which naturally-occurring Cr(III) is oxidized and mobilized to groundwater in the Sacramento Valley. Geogenic Cr(VI) is one of several groundwater quality issues that are of concern for municipalities and water districts in the valley (Water Resources Association of Yolo County, 2007). Relatively 1 high concentrations of NO ; Dawson et al., 3 (up to 19 mg N L 2008) in valley groundwaters are also of concern, and are evidence of the impact of extensive agricultural activity on groundwater quality. There is a potential link between groundwater Cr(VI) and agriculturally-derived NO 3 . As a result of irrigation (for which NO 3 is a tracer), the hydrology of the Central Valley has been dramatically altered with a large increase in the amount of groundwa-

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ter recharged through soils (Bertoldi et al., 1991; Davisson and Criss, 1993). Thus leaching of mobile elements, including Cr(VI), from the vadose zone may contribute substantially to groundwater quality issues in the Central Valley. Previous studies have described characteristics of Cr in ultramafic source material and the weathering and transport of this material on a regional scale (Goldhaber et al., 2009; Morrison et al., 2009; Morrison, 2010). The goals of the current study were: (1) to determine the concentrations and major residences of Cr in soils and vadose sediments of the southwestern Sacramento Valley; (2) to determine the occurrence and content of Cr(VI) in the subsurface; (3) to correlate Cr(VI) concentrations to soil geochemistry or characteristics; and (4) to understand the effect of acid production on the generation of Cr(VI) in valley soils. Soils and vadose zone alluvial sediments were sampled from depths up to 32 m below the surface. Samples were analyzed for total element concentrations and leachable Cr(VI) content. Selective extractions for Mn and Fe oxides and associated trace elements were performed on select samples. In addition, borehole electrical conductivity and magnetic susceptibility surveys were performed as a proxy for sediment grain size and the presence of Cr-bearing spinel phases (e.g. chromite), respectively. Lastly, valley groundwater data were analyzed for evidence of processes tied to Cr(VI) generation and these data were compared to the results of soil incubations designed to mimic in situ processes.

2. Study area The Sacramento Valley encompasses the northern third of the Central (Great) Valley (Fig. 1), an approximately 103,600 km2 (40,000 mile2) basin that is one of the world’s most productive agricultural areas. Major crops in the valley include rice, nuts, tomatoes and grain. Due to the area’s Mediterranean climate with hot, dry summers, much of the cropland is irrigated by surface and groundwater. The present research was further focused on Yolo County in the southern Sacramento Valley (Fig. 1). Approximately half of the land and 95% of water usage for Yolo County is dedicated to agriculture (Water Resources Association of Yolo County, 2007). About 30% of irrigation water is provided by groundwater and the remainder from surface water, but the majority of municipal water is obtained from groundwater (Water Resources Association of Yolo County, 2007). The Sacramento Valley is comprised of Quaternary alluvium which is underlain on the western edge by the Pliocene Tehama Formation, a cemented silt, sand, gravel and clay formation derived from the Klamath Mountains and Coast Range. The Tehama Formation overlies the Great Valley Sequence, a thick sequence of marine sedimentary rocks that outcrops in the eastern foothills of the Coast Range (Goldhaber et al., 2009; Morrison et al., 2009). Samples were collected from an area in the southern-central portion of Yolo County near the city of Davis (Fig. 1). Groundwater well depths in Yolo County range from 12 to 460 m (California Groundwater Bulletin 118). There is no regional confining layer in the Sacramento Valley aquifer, but aquifers can be locally confined by clay lenses. Deeper wells penetrate the Tehama Formation. The Tehama Formation is up to 2500 m thick in the central part of the Sacramento Valley (Olmsted and Davis, 1961). Shallower wells penetrate Quaternary alluvium transported by Putah Creek (Davisson and Criss, 1993). Most of the core samples were collected from this alluvium. Putah Creek (Fig. 1) is one of only three perennial streams that enter the Sacramento Valley from the Coast Range. It drains an area with substantial outcrops of ultramafic rocks and serpentine soils (Bertoldi et al., 1991; Goldhaber et al., 2009; Morrison et al., 2009). Three of the cores studied (D6–D8) were obtained from a US EPA Superfund site on

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Fig. 1. Maps showing the location of the study area and of drill cores (triangles, D1–D8) and shallow soil cores (circles, S9–S14).

the campus of the University of California, Davis. Chromium(VI) concentrations as high as 600 lg L1 in shallow groundwater have been measured near a historic landfill area at this site (R. Devany, Weiss Associates, 2011, pers. comm.; Chung et al., 2001). It has not been determined whether this Cr(VI) is contamination from historic activities on the site or occurs naturally. Cores D6–D8 collected from this site were >100 m outside of the area with anomalously high concentrations of Cr(VI) in shallow groundwater. 3. Methods 3.1. Sample collection Fresh soil core samples from six boreholes (D1–D6) were obtained from the University of California, Davis between August and November, 2007 (Fig. 1). The core samples from groundwater monitoring wells drilled with a hollow stem auger. The boreholes were between 17 and 26 m deep. Discrete sediment samples 15 cm long by 5 cm diameter were collected approximately every 1.5 m. These samples were placed in plastic bags and transported on ice to USGS laboratories in Denver, CO where they were stored at 4 °C. Within 4 days of collection, core material was removed from the original core liner, homogenized, and subsampled for various analyses. Discrete samples from two archived, 32 m deep boreholes (D7–D8) were collected at approximately 0.6 m intervals. Because these core samples had been air dried and stored for several years, they were analyzed for total element concentrations only. Six continuous soil cores (S9–S14) were collected in May 2008 by hand using an 8 cm diameter auger. These soil cores were up to 4 m deep and core material was placed in sterile polyethylene bags. Fresh samples were transported overnight to the laboratory and immediately stored at 4 °C. Field soil moisture content was determined by gravimetric difference between moist and oven-dried (105 °C) soils. 3.2. Total chemistry Samples were sieved to <2 mm and then machine ground to pass through a 100 mesh (<150 lm) sieve before chemical analysis.

Two different analyses were performed by the USGS contract laboratory, SGS Minerals in Toronto, Ontario, Canada: (1) a 4-acid digestion (HNO3, HCl, HF, HClO4) followed by inductively coupled plasma-atomic emission spectrometry (ICP-AES) and inductively coupled plasma-mass spectrometry (ICP-MS) analyses and (2) a Li metaborate fusion followed by ICP-AES analysis. Standard reference materials (SRMs) SRM 2709 (NIST) and DGPM-1 (USGS) as well as 4 USGS-produced SRMs were used to monitor quality control. Details of methods and quality control are reported by Smith et al. (2009). Some samples from cores S9 to S14 were combined based on similar soil texture and color for total chemical analyses. Inorganic C was determined by coulometric titration at SGS laboratories. Total C and N were both determined by pyrolysis at the Colorado State University (CSU) Soil, Water & Plant Testing Laboratory (Fort Collins, Colorado). Organic C was determined by difference of total and inorganic C. Soil pH was determined on slurries of 1:1 air dried soil:deionized water. 3.3. Selective extractions Two selective extraction techniques were performed on field moist samples from cores S10, S12, D1, and D6 that had been stored under refrigeration for several months. Approximately 10 g of soil from each sample were homogenized by hand before being subsampled for the extractions. Selective extractions were also performed on air-dried, machine-ground samples from cores D7 and D8. Seven air-dried ground samples from cores D1, D6, and S10 were also extracted to compare the different sample preparation and storage techniques. Manganese and trace elements associated with easily reducible Mn oxides were determined using a hydroxylamine hydrochloride (HA) extraction. Two hundred mg (dry weight) soil was shaken in 400 mL 0.1 M hydroxylamine hydrochloride (pH  3.6) for 2 h (Neaman et al., 2004a,b). Samples were filtered (0.45 lm pore size) and filtrates analyzed by ICP-MS at the USGS laboratories in Denver, Colorado. Element concentrations in extraction solutions were blank-corrected and then converted to concentrations in the original soil sample (dry weight). Iron and trace metals associated with amorphous Fe oxides were determined by a sodium citrate–bicarbonate–dithionite (CBD) extraction. The procedure was similar to HA extractions except

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that 200 mg samples were stirred for 1 h at 80 °C in 10 mL of 1 M NaHCO3, 80 mL 0.3 M sodium citrate, and 2 g sodium dithionite (Neaman et al., 2004a). Coefficients of variation for hydroxylamine triplicate extractions on three different samples were 1–5% (Mn), 1–6% (Co), and 3–6% (Ni) and for dithionite–citrate triplicate extractions on three samples were 4–14% (Mn), 2–16% (Fe), 3– 14% (Co), 7–18% (Ni), and 6–12% (Cr). Oze et al. (2004) explored the possibility that chromite-Cr might be released during sequential extractions, but found that less than 0.5 and 1.4 mg kg1 Cr were released from pure chromite during hydroxylamine–acetic acid and HF extractions, respectively. 3.4. Leachable Cr(VI) measurements Within 2 weeks of collection, field moist samples (35–40 g wet weight) were shaken in 50 mL centrifuge tubes with 25 mL of a Ca/ Mg sulfate solution (5 mM in each of CaSO4 and MgSO4) for 15 min (Chung et al., 2001). The purpose of the added SO2 was to ex4 change with any sorbed Cr(VI). Samples were centrifuged at 2500 rpm for 5 min and filtered (0.45 lm pore size). The concentration of Cr(VI) in the supernatant was determined colorimetrically using a modification of EPA Method 7196A. One milliliter of diphenylcarbazide reagent (0.50 g diphenylcarbazide in 30 mL acetone) and 50 lL of 10% H2SO4 were added to 10 mL of sample. Color was allowed to develop for approximately 10 min and absorbance was measured at 540 nm using a 5 cm optical glass cell (Beckman DU-64 spectrophotometer). Seven of the D core samples that had been stored field-moist at 5 °C for several months were analyzed again. Both sulfate solution and deionized water extractions were performed on select samples to assess what portion of Cr(VI) was sorbed and required anion exchange to be mobilized. 3.5. Cr(III) oxidation test To test the capacity of samples to oxidize Cr(III)–Cr(VI), a modified version of the Cr oxidation test described by Bartlett and James (1979) was used. Three grams of moist soil was shaken with 15 mL of 1 mM CrCl3 in a 50 mL centrifuge tube for 2 h. Samples were centrifuged and the concentration of Cr(VI) in the supernatant was determined by the diphenylcarbazide method previously described. 3.6. Borehole magnetic susceptibility and electrical conductivity measurements Magnetic susceptibility measurements quantify the induced magnetism in a material and are controlled by Fe-rich minerals; electrical conductivity quantifies the ability of a material to pass electrical current and is controlled by the dissolved ions within pore water and sedimentary lithology (Hearst et al., 2000). Both properties were measured in ground water wells with logging tools manufactured by Mt. Sopris Instruments, Inc., and the logs were collected in accordance with the procedures recommended by the manufacturer. Measurements were made as the instrument descended each well and data was checked for consistency as the tool was retracted. Occasionally, the magnetic susceptibility and electrical conductivity measurements were affected by metal centralizers within the well casings. The affected measurements span intervals of less than 0.5 m, and these measurements were replaced by values interpolated from the measurements just above the and below the affected interval. 3.7. Soil incubations A homogenous soil for use in incubation experiments was prepared by combining equivalent amounts of core S11 samples ranging from 1.0 to 3.6 m depth. Because this soil was collected on a

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grassland nature preserve, it was considered to be the most unaltered by agriculture or other land use impacts of all the soils collected. Subsamples of this homogenized soil were used for all incubations. Six subsamples (12.5 g wet weight) of soil were prepared in 50 mL glass beakers. Aliquots of various concentrations of trace metal grade HCl were added to the subsamples resulting in a final moisture of 16 wt.%. This was as close to the field moisture content (15 wt.%) as possible. The final amount of H+ added ranged from 0 to 70 mmoles per kg soil (dry weight). The beakers were sealed with laboratory film to allow exchange of air, but not moisture. After 2 weeks, subsamples of the incubated soils were used for pH and Cr(VI) measurements as previously described. In addition, a subsample was shaken for 1 h with ultrapure deionized water. This water extract was then filtered (0.45 lm pore size), acidified, and analyzed for major cations and trace metals by ICPMS. An additional five replicate incubations of the highest H+ concentration were sampled as a time sequence. A control incubation time series was conducted similarly except that 70 g portions of soil were incubated in 250 mL glass Erlenmeyer flasks and a soil moisture of 16% was obtained by adding deionized water. 4. Results 4.1. Sample descriptions The field characteristics of soils from S cores generally followed the type description of the mapped soil series; Hillgate loam (S12), Sacramento soils flooded (S14), Capay silty clay loam (S13), Brentwood silty clay loam (S11), Yolo silt loam (S10), and Yolo silty clay loam (S9) (Soil Survey Staff, National Resources Conservation Service). Small (3 mm diameter) red and black concretions presumed to be Fe and Mn oxides were observed in many of the deeper samples. Soil pH generally ranged from 7.0 to 8.8, but was less than 7.0 for a few, typically shallower, samples (Figs. 2 and A1). Organic C concentrations were as high as 2.3% in the surface soils but decreased with depth and were typically <0.5% below 1 m depth. Concentrations of inorganic C were low (<0.1%) in the near surface samples, but concentrations as high as 0.9% were observed in the deeper samples of cores S11–S14. Soil moistures ranged from 3.6 to 28 wt.% (median 16%) (online data file). Sediments from the drill cores were predominately clay-sized material with irregularly occurring lenses of fine grained sand or gravel. The texture of the samples generally matched the classifications obtained from drill logs (Figs. 3 and A2). Geophysical logging surveys of boreholes D1, D2, and D5–D8 indicated that high electrical conductivities and low magnetic susceptibilities are usually associated with clays, whereas low conductivities and high susceptibilities are usually associated with sands (Figs. 3 and A2). Most clay-rich samples contained extensive color mottling of gray, red and black areas. Small (3 mm diameter) to large (>1 cm diameter) concretions of Fe and Mn oxides and carbonate were present in many of the samples. Sample pH was between 7.0 and 8.6 (Figs. 3 and A2). Soil moistures ranged from 4.6 to 27 wt.% (median 17%) (online data file). Concentrations of inorganic C varied from <0.003% to 2%. Organic C concentrations were determined for select samples from cores D1 and D6 and were highest (0.6% by wt.) in the first depth interval for both cores. 4.2. Total element concentrations The complete data set for total concentrations of 44 elements in core samples is available in an online data file. The distributions of the concentrations of elements typically associated with ultramafic rocks (Fe, Mg, Cr and Ni) as well as Mn are shown in Fig. 4. The concentrations of all of these elements are elevated with respect to the

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Fig. 2. Depth profiles of elements and species associated with Cr cycling in cores S10–S12 collected in the southern Sacramento Valley. For all cores except S10, some samples from adjacent depths were combined for total element analysis. Profiles for cores S9, S13, and S14 are presented in Fig. A1 (online).

geometric mean determined for surface soil in the conterminous USA (Shacklette and Boerngen, 1984) (Fig. 4). For a given sample, the Cr concentration determined using the 4-acid digestion (75– 599 mg kg1) was always lower than that determined using Li metaborate fusion (140–2220 mg kg1). The ratio of the two concentrations ranged from 0.16 to 0.98. The value determined from the Li-metaborate fusion analysis is subsequently referred to as total Cr (CrT) and that determined by 4-acid digestion as 4-acid Cr (Cr4A).

Median CrT concentrations were notably lower for cores S12– S14 (190–230 mg kg1) than for all of the other cores (Fig. 5). Median CrT concentrations for cores D1–D8 and S9–S11 ranged from 250 to 615 mg kg1 with no consistent trend toward higher median concentrations observed for the deeper cores. However, CrT was much more variable with depth in the deeper cores with some samples having CrT concentrations exceeding 1000 mg kg1 (Figs. 3 and A2). In the D cores, CrT was highest in coarser-grained material as indicated by drill logs and electrical conductivity measurements

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Fig. 3. Depth profiles of elements and species associated with Cr cycling in cores D1, D5, and D6 of vadose zone alluvium in the southern Sacramento Valley. Profiles for cores D2–D4, D7, and D8 are presented in Fig. A2 (online).

(Figs. 3 and A2) and was positively correlated with the borehole magnetic susceptibility measurements (Figs. 3, 6 and A2). Total Mn (MnT) concentrations for all samples ranged from 267 to 1250 mg kg1 with one outlying value of 2170 mg kg1 for the

12.2–12.5 m depth sample from core D8 (Fig. 4). Although the range in MnT was generally greater in D cores than in S cores, all cores had similar median concentrations (663–812 mg kg1) (Fig. 5).

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Fig. 4. Distributions of the concentrations of elements associated with ultramafic rocks and Mn in S (light gray bars) and D cores (dark gray bars). Solid and dotted vertical lines indicate geometric mean and 95th percentile, respectively, for total element concentrations in surface soils of the conterminous USA (Shacklette and Boerngen, 1984). Geometric means (gm) for S and D cores are listed below each distribution plot.

4.3. Selective extractions Selective extractions were performed on select samples from D and S cores to determine the relative amounts of more reactive Cr associated with Mn and Fe oxides and as a measure of the amounts of reactive Mn oxides in samples. 4.3.1. Hydroxylamine hydrochloride (HA) extractions The complete results for selective extractions are presented in the online data file. Hydroxylamine–hydrochloride extractions have been shown to dissolve easily reducible Mn oxides and associated trace elements (Neaman et al., 2004a,b, 2008) When possible, selective extractions were performed on samples that had been stored refrigerated under field moist conditions because air drying has been shown to affect reactive soil Mn oxides (Ross et al., 2001). When field-moist samples were not available, extractions were performed on air-dried, ground samples. For seven samples, extractions were performed on splits that had been stored in each manner to assess sample storage effects (Table 1). Concentrations of Mn in the HA extractions (MnHA) determined for field moist samples ranged from 12% to 111% (mean = 64%) of respective MnT concentrations with one outlier of 143% for the 13.1–13.4 m depth sample from core D6 (Figs. 2, 3 and A2 For air-dried samples, MnHA concentrations ranged from 19% to 91% (mean = 51%) of MnT. The concentrations of MnHA in the field-moist samples that were

greater than 100% of MnT were likely a result of increased heterogeneity due to greater difficulty homogenizing moist samples. However, overall there was no significant difference (paired t-test) between MnHA for field-moist versus air dried samples (Table 1). Concentrations of MnHA in the deep core samples correlated well with Cr(III) oxidation potential measurements (Fig. 7) suggesting that easily reducible Mn oxides were responsible for oxidation of added Cr(III) during the tests. This result also suggests that Cr(III) oxidation potential measurements can be used as a proxy for the amount of easily reducible Mn oxides in the deep core samples for which selective extractions were not performed. Chromium was not detected in the HA extractions for any of the samples, but Co (6 and 26 mg per kg bulk soil) and Ni (4 and 64 mg kg1) were released (online data file). The most abundant cations released in these extractions were Ca (up to 3 wt.% of soil) and Mg (up to 1 wt.%). Because of the low pH of the solution, HA extractions also likely dissolve carbonates which were abundant in some of the deep core samples. 4.3.2. Citrate–bicarbonate–dithionite (CBD) extractions Citrate–bicarbonate–dithionite (CBD) extractions have been shown to dissolve both Fe oxyhydroxides and Mn oxides and their associated trace elements (Neaman et al., 2008). The concentrations of both major and trace elements released from air-dried samples were significantly greater (paired t-test) than for

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et al., 2004a,b). The CBD extractions released 25–64% of FeT and 1–18% of SiT. Between 9 and 70 mg Cr per kg soil was also released in the CBD extractions (Figs. 2, 3 and A2 and the amounts of Co (5– 22 mg kg1) and Ni (20–180 mg kg1) released in the CBD extractions were almost always similar to or greater than the amounts released in the HA extractions. A stronger positive linear correlation was observed between Cr and Si than between Cr and Fe released in the extractions (Fig. 8). 4.4. Leachable Cr(VI)

Fig. 5. Box and whisker plots of the concentrations of CrT, MnT, Cr(VI) in soil and sediment, estimated Cr(VI) in pore water, and Cr(III) oxidation potential in S and D cores. The top and bottom of box indicate the 25th and 75th quartiles, respectively, and the line within the box indicates the median. The whiskers indicate the maximum and minimum data values that lie within 1.5 of the interquartile range. Data that lie outside of the whiskers are indicated by circles. See text for estimation of Cr(VI) in soil pore water (SPW).

Fig. 6. Relationship between borehole magnetic susceptibility measurements and CrT measured in samples from cores D1, D2 and D5–D7.

field-moist treatments (Table 1), likely due to more reactive surface area exposed by grinding of the air-dried sample (Neaman

Concentrations of leachable Cr(VI) determined using an extraction solution with 10 mM SO2 ranged from below detection to 4 42 lg per kg soil in all samples from S and D cores (Figs. 2, 3, A1 and A2). Chromium(VI) was not detected in surface soils and its concentrations typically increased with depth in the S cores (up to a maximum of 15 lg kg1 in core S11) (Figs. 2 and A1 The magnitude of this increase of Cr(VI) with depth was substantially less for cores S12–S14 than for cores S9–S11. The maximum Cr(VI) concentrations observed in the D cores were higher than for the S cores (Figs. 3 and A2). Each D core exhibited a maximum Cr(VI) concentration, but the depth of this maximum concentration varied between cores. There were no apparent correlations between concentrations of Cr(VI) and concentrations of CrT, Cr4A, CrCBD, or Cr(III) oxidation potential. Seven samples from D cores that had been stored field-moist at 5 °C were extracted again with both sulfate solution and deionized water to assess what percentage of Cr(VI) measured in the sulfate solution extractions was released by anionic exchange. The Cr(VI) concentrations in the sulfate solution extractions made on these stored samples were between 66% and 135% of the Cr(VI) concentrations in the original sulfate extractions (Fig. 9) suggesting that, during storage, Cr(VI) was reduced in some soils and oxidized in others. Chromium(VI) concentrations in the deionized water extractions made on the stored samples were similar to or greater than concentrations in the sulfate solution extractions performed at the same time (Fig. 9). This indicates that very little of the Cr(VI) measured in the original sulfate solution extracts was strongly sorbed to soil particles. It is unclear why Cr(VI) concentrations were significantly greater in the deionized versus sulfate solution extracts for some of the samples. 4.5. Long term and acid titration soil incubations Samples of homogenized soil from core S11 (1.0–3.6 m depth) were incubated at 16% soil moisture for 130 days. Concentrations of Cr(VI) in these incubations increased linearly with time whereas soil pH remained relatively constant (Fig. 10a). The Cr(VI) concentration in the incubation sampled at the end of the experiment (130 days) was almost twice that of the initial concentration (0 days). Concentrations of water-leachable total Cr were near or below detection limit for all samples. Chromium(VI) was also measured at the end of 2-week soil incubations which were amended with various amounts of HCl (Fig. 10b). The concentration of Cr(VI) (83 lg kg1) in the incubation with the highest amount of H+ added (70 mmoles kg1) was more than 6 times greater than Cr(VI) (13 lg kg1) in the incubation with no acid addition. The concentrations of water-leachable major cations (Na, K, Mg, Ca) also increased with increasing acid addition although Mg plus Ca comprised the majority (75–94%) of these cations (Fig. 10b). Furthermore, Mg was 5–8 times more abundant than Ca. Replicates of the soil incubation with 70 mmoles H+ per kg soil were harvested at different times during the 2 week incubation (Fig. 10c). Approximately half of the Cr(VI) observed after 2 weeks was generated within the first day of incubation. The rate of

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Table 1 Comparison of concentrations of elements released in selective extractions made on field moist versus air-dried samples. Core

S10

Depth interval (m)

3.2–3.6

D1

D6

Student’s paired t-test

3.6–4.0

1.5–1.8

13.7–14.0

18.3–18.6

13.1–13.4

14.3–14.6

95% confidence interval of difference (p value)a

Hydroxylamine–hydrochloride (HA) Mn (mg kg1) Moist 763 Dried 404 1 Al (g kg ) Moist 0.08 Dried 0.15 Moist 1.5 Mg (g kg1) Dried 1.4 Ni (mg kg1) Moist 26 Dried 19 Co (mg kg1) Moist 13 Dried 11

211 427 0.04 0.20 0.7 1.4 11 22 6 12

682 556 0.14 0.23 4.7 3.6 59 45 20 17

328 369 0.07 0.11 8.2 9.5 14 12 6.9 8.2

201 118 0.06 0.67 2.6 2.6 4.0 4.7 3.4 2.7

1530 978 0.13 0.23 8.4 10 44 26 26 18

267 238 0.07 0.10 4.6 2.8 5.5 1.4 4.1 3.9

109 to +364 (0.24)

Citrate–bicarbonate–dithionite (CBD) Fe (wt.%) Moist 1.5 Dried 1.6 Mn (mg kg1) Moist 481 Dried 571 1 Al (g kg ) Moist 1.6 Dried 2.3 Mg (g kg1) Moist 4.3 Dried 6.4 SiO2 (g kg1) Moist 14 Dried 24 Cr (mg kg1) Moist 31 Dried 46 Ni (mg kg1) Moist 73 Dried 104 Co (mg kg1) Moist 16 Dried 19

1.6 1.9 464 565 1.6 3.5 4.2 13 19 48 33 64 80 142 17 19

1.9 2.1 628 707 2.1 4.6 3.9 5.5 16 25 37 58 81 113 20 22

1.5 1.6 222 493 1.1 2.2 10 18 18 22 27 40 46 54 6.2 12

1.5 1.8 172 273 1.2 4.3 3.0 4.1 17 33 26 52 36 71 6.8 8.3

1.5 2.1 940 1180 1.5 3.7 4.7 19.2 12 24 23 36 38 64 18 22

1.3 1.6 287 366 1.2 2.9 4.8 9.0 14 31 21 41 20 45 5.6 8.1

0.34 to +0.03 (0.088) 1.2 to +1.1 (0.86) 4.1 to +13.6 (0.24) 3.0 to +4.9 (0.58)

0.43 to 0.11 (0.006) 213 to 62 (0.004) 2.64 to 1.13 (0.0009) 10.3 to 1.2 (0.02) 21.2 to 6.5 (0.004) 26.2 to 13.5 (0.0003) 46.2 to 16.3 (0.002) 4.4 to 1.6 (0.002)

a Two tailed t-test. Difference = moist minus dried. p-Value <0.05 indicates that the null hypothesis ‘‘there is no difference between values for moist and dried samples’’ can be rejected at the 95% confidence level.

Fig. 7. Cr(III) oxidation potential versus MnHA. Cores D1 and D6 (circles) and S10 and S12 (crosses). Linear fit is to cores D1 and D6 only.

generation quickly slowed and was constant at 1.7 lg kg1 day1 for the last 3 times. Concentrations of water-leachable Ca plus Mg rapidly increased within the first day of incubation and then remained relatively constant for 2 weeks. Soil pH was at a minimum of 6.3 after 1 day of incubation and then gradually increased to a pH of 7.3 over the 2 weeks.

5. Discussion The large concentrations of CrT and relatively high Cr oxidation potentials of the soils and sediments studied suggest that they are well poised for the generation of Cr(VI). Indeed, measured Cr(VI) concentrations of up to 42 lg kg1 indicate the generation of this

toxic substance in situ. However, the lack of correlation between Cr(VI) concentrations and CrT, MnHA, or Cr oxidation potential suggest that Cr(VI) generation is kinetically limited in the alkaline pH range of the soil and sediments studied. Results of the laboratory incubations show increased rates of Cr(VI) with added H+. Increased mobilization of Cr(III) through cation exchange reactions and/or lowered soil pH likely increases the rate of Cr(III) oxidation on abundant Mn oxide surfaces. Given that the pH of the majority of core samples was 8.0 or greater, it is likely that much of the Cr(VI) in the soils and alluvium studied was not strongly sorbed to soil particles. Above pH 8 only a small fraction of Cr(VI) is predicted to sorb to hydroxyl groups on soil Fe oxides, Al oxides and clays (Rai et al., 1989; Bradl, 2004). Comparison of sulfate solution and deionized water extractions confirmed this (Fig. 9). Taking into account the small sorbed fraction, the soil Cr(VI) concentration, and soil moisture data in combination show that pore waters contained up to 64 and 180 lg L1 Cr(VI) for S and D cores, respectively (Fig. 5). These concentrations are on the same order as those observed in some area groundwaters. Thus, despite the exceedingly small fraction of total Cr that is present as Cr(VI), leaching of vadose zone soils could be a significant source of Cr(VI) in valley ground water. The potential for oxidation of Cr(III) to Cr(VI) in the valley soils and sediments likely depends less on CrT concentrations and more on the relative reactivity of the Cr, which is affected by transport and weathering. Variable CrT concentrations with depth, particularly in the deep cores, suggest episodes of greater relative transport of ultramafic clastic material by Putah Creek from the Coast Range. This ultramafic material likely mixed with varying amounts of sediment derived from the marine Great Valley sequence that outcrops extensively along Putah Creek as well as Sacramento River flood deposits (Davisson and Criss, 1993; Goldhaber et al., 2009). For example, the lower concentrations of CrT in cores S12 and S13, located >5 km north of Putah Creek, and S14, located in an area

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Fig. 8. Relationship between (a) Cr and Fe and (b) Cr and Si released in citrate dithionite selective extractions. Concentrations are based on bulk dry weight soil. Data include extractions performed on both field moist and air dried samples from cores D1, D6–D9 (circles) and S10, S12 (crosses). Linear fits are to all data points shown.

Fig. 10. Behavior of Cr(VI), water leachable Mg2+ + Ca2+, and pH in control and acid amended laboratory incubations of soil from core S11. (a) Response over time for control incubations. (b) Increase in [Cr(VI)] and water leachable Mg2+ + Ca2+ with increasing additions of HCl. All measurements were made after 2 weeks of incubation. (c) Response over time for addition of 70 mmoles kg1 HCl. Linear regression shown is for last four time points only.

Fig. 9. Comparison of Cr(VI) extracted from selected samples using Ca/Mg SO2 4 (10 mM) and deionized water (DI). Samples were stored field moist in the refrigerator between fall 2007 and spring 2009 extractions. Error bars are ±half the range of duplicate or triplicate extractions.

that is annually flooded by the Sacramento River, may reflect less input of ultramafic material than for cores located near Putah Creek. The positive correlations between CrT and grain size as inferred from drill logs and borehole electrical conductivity (Figs. 3 and A2 and between CrT and magnetic susceptibility (Figs. 3, 4, and A2) suggest that sediments with the greatest CrT concentrations result from high energy (hence coarse grain size) transport events of ultramafic material eroded from the Coast Range and transported by Putah Creek. The co-occurrence of chromite and magnetite in serpentinized ultramafic rocks and the relatively high magnetic susceptibility of chromite itself are well-documented (Oze et al., 2004; Morrison et al., 2009) and have been used as geophysical signals for chromite prospecting (e.g. Yungul, 1956). These two minerals have similar densities and may have been transported and deposited together, resulting in sediments with relatively high CrT content and magnetic susceptibility.

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Chromite transported from the Coast Range appears to be the dominant Cr-bearing mineral in the valley soils and sediments as evidenced by the large discrepancy between Cr4A and CrT concentrations in many samples and by scanning electron microscopy studies (Morrison et al., 2009; Morrison, 2010). Morrison et al. (2009) showed that acid digestions etch the surface of, but do not dissolve the chromite. Thus the ratio of Cr4A to CrT (Cr4A:CrT) is expected to decrease with increasing chromite grain size. A further analysis of the role of grain size in determining Cr abundance can be made by assuming an approximately spherical geometry for chromite grains, and that Cr4A and CrT represent the surface area and volume of a sphere, respectively. With these assumptions, the following relationship is derived:

Cr4A : CrT ¼ AðCrT Þ0:333

ð1Þ

where A is a fitting constant. Accounting for any more labile forms of Cr that are completely dissolved in the 4-acid digestion (represented by the constant B), the relationship becomes:

Cr4A : CrT ¼ AðCrT  BÞ0:333

ð2Þ

A plot of Cr4A:CrT versus CrT (Fig. 11) shows this relationship fits the data well and that samples with the highest CrT concentrations also contain the largest average grains of chromite. This is likely one reason that Cr(VI) concentrations do not correlate with CrT. Much of the Cr in samples with very high CrT is sequestered in the interior of chromite grains and unreactive. Apparent outliers are samples from cores D7 and D8 that are deeper (26–32 m) than samples from other cores. Sample grinding as preparation for digestions may have partly obscured a particle size effect. However, grinding criteria only require ground soils to pass a 100 mesh sieve, which likely still allows grain size effect expression for most of the finer grained samples. The outlier samples from core D7 were classified as gravels in the driller’s log. Grinding of the <2 mm fraction of these samples likely obscured grain size effects by breaking up larger chromite particles and exposing more surface area in the 4-acid digestions. Fig. 11 shows that the best fit of Eq. (2) was obtained with B = 147 mg kg1. This value is a rough estimate of the amount of Cr that is not associated with chromite, but instead with more labile residences. This more labile Cr includes that released in the

Fig. 11. Relationship between the ratio of Cr4A/CrT and CrT. The dashed line represents the function y = 2.48(x147)0.333. This function was determined with a linear regression on log transformed data. Parameter B (Eq. (2)) was adjusted so that the power factor was 0.333. The apparent outlying data for the deepest samples of cores D7 and D8 (open circles) were not included in the fit.

CBD extractions (9–70 mg per kg bulk soil) as well as up to 65 mg Cr per kg bulk soil measured in the clay-sized fraction of Sacramento Valley soils and sediments by Morrison (2010). These two different quantifications are not exclusive due to the likely presence of a portion of the Fe oxide particles residing in the clay-sized fraction. It is interesting that a stronger positive correlation was observed between Cr and Si released in the CBD extractions than between Cr and Fe (Fig. 8). Morrison (2010) observed Cr-containing, nano-crystalline, siliceous ferrihydrite in the less than 2 lm fraction in valley soils using transmission electron microscopy and it can be assumed that a similar relationship would be likely in the coarser grained material. The extent to which the CBD extractions could dissolve other silicate phases is not clear, although it has been suggested that clay minerals should be stable in the CBD extractant due to its near neutral pH (Neaman et al., 2004a,b). Morrison (2010) also observed Cr associated with clays (smectite–illite) as well as Cr associated with silicate-rich rims surrounding chromite grains in serpentine soils (Oze et al., 2004; Morrison, 2010). However, these rims are absent in chromite in valley soils suggesting they have been weathered away during transport (Morrison, 2010). The lack of any Cr associated with Mn oxides despite the association of relatively large amounts of ultramafic trace elements (Ni, Co) has been previously reported (Neaman et al., 2004a, 2008). It is likely that any Cr that sorbs to Mn oxides is rapidly oxidized to Cr(VI) (Bartlett and James, 1979) which is unlikely to sorb to soil particles in the pH range of the valley soils (Rai et al., 1989; Bradl, 2004). Even though only a small fraction of CrT appears to be relatively labile, this more reactive Cr(III) is still quite abundant. Reactive Mn oxides are also quite abundant and, thus, it is unlikely that the capacity for Cr(VI) production is limited by the abundance of these two reactants. This conclusion is supported by the lack of correlation between concentrations of Cr(VI) and CrT, Cr associated with Fe oxides, or Cr(III) oxidation potential. It is more likely that the low solubility and mobility of Cr(III) in the pH regime of the valley soils kinetically limits Cr(VI) production. This result is in contrast to Cr(VI) concentrations that correlate with Mn oxide abundance in a lateritic regolith developed on ultramafic rocks in New Caledonia (Fandeur et al., 2009). These lateritic soils are highly weathered and acidic (pH  4–6), conditions that promote the mobility of Cr(III). Despite low Cr(III) mobility in the Sacramento Valley soils, it is apparent from the laboratory soil incubation experiments that Cr(VI) production in the vadose zone is environmentally important. The generation rate of 24 lg Cr(VI) per kg soil per year (0.065 lg kg1 day1) determined from Fig. 10a suggests that the Cr(VI) concentrations observed in core material could be regenerated on a yearly basis. Leaching of Cr(VI) through the vadose zone could be responsible for the Cr(VI) maxima detected in the deep cores with the depth of the maximum depending on local hydrological conditions and rates of Cr(VI) production in the soils and sediments. The absence of detectable Cr(VI) in the near-surface soils is likely due to reduction of Cr(VI) by abundant organic matter outpacing Cr(VI) production (Bartlett and Kimble, 1976; Wittbrodt and Palmer, 1995; Banks et al., 2006). Chung et al. (2001) measured Cr(VI) production rates for vadose zone material (2–22 m depth) in the vicinity of cores D6–D8 that were between 200 and 1000 lg Cr(VI) kg1a1. Faster rates may have been due to differences in the deeper material incubated. However, they also lyophilized and subsequently rehydrated the material before incubation which may have substantially disturbed the steady state dynamics of sorbed Cr(III) and promoted Cr(VI) generation. Chromium(VI) generated in the vadose zone is likely susceptible to leaching to groundwater. Irrigation in the Great Valley has substantially increased groundwater recharge through the vadose

C.T. Mills et al. / Applied Geochemistry 26 (2011) 1488–1501

zone relative to atmospheric precipitation alone (Bertoldi et al., 1991). Davisson and Criss (1993) suggest that up to half of the groundwater in some areas of the Sacramento Valley can be attributed to irrigation return water as indicated by d18 OH2 O values and elevated NO 3 concentrations. Fig. 12 shows a positive correlation between Cr(VI) and NO 3 for groundwater data collected as part of the Ground-Water Ambient Monitoring and Assessment (GAMA) program (Dawson et al., 2008). This correlation is consistent with the leaching of Cr(VI) generated in the vadose zone by irrigation return water high in NO 3 . A similar mechanism may also be occurring in portions of the Mojave Desert (Izbicki et al., 2008). In addition to providing increased leaching of the vadose zone, irrigation return water may contain solutes that influence Cr(VI) generation rates and mobility. For example, Becquer et al. (2003) observed soil pore water concentrations of up to 700 lg L1 Cr(VI) in phosphate fertilized soils compared to 20 lg L1 in a similar soil not under agricultural production. This was the result of phosphate exchange with sorbed Cr(VI). A similar mechanism is unlikely in the Sacramento Valley soils. The pH of the soils studied by Becquer et al. (2003) (pH  5) was much lower than that of Sacramento Valley soils which likely increased the rate of production and decreased the mobility of Cr(VI) (Rai et al., 1989; Bradl, 2004). In the Sacramento Valley soils, it is more likely that cations associated with irrigation return (e.g. H+, K+) will increase the mobility of Cr(III) and its subsequent oxidation on Mn oxides. Increased concentrations of H+ commonly result from the microbial oxidation of ammonium in fertilizers:

NHþ4 þ 2O2 ¼ NO3 þ 2Hþ þ H2 O

ð3Þ

Complete nitrification of a typical application of ammonium fertilizer (100 mg N per kg soil) results in the input of about 14 mmole H+ per kg soil. The results of the soil titration experiment (Fig. 10b) show that the valley soil studied is buffered against this order of magnitude of H+ addition by cation exchange with Mg2+ and Ca2+. The positive correlation between water leachable Mg plus Ca and generated Cr(VI) are consistent with the hypothesis that small amounts of Cr(III) are also mobilized through cation exchange and quickly oxidized on Mn oxides to Cr(VI). Incubation experiments have previously shown increased mobility of trace

Fig. 12. Relationship between Cr and NO 3 in southern Sacramento Valley groundwaters as reported by Dawson et al. (2008). Diamonds represent wells along the Putah Creek, triangles represent other wells sampled in the southern Sacramento Valley, and open circles represent wells with P10 mg L1 dissolved Mn suggesting that NO 3 concentrations are near zero due to more reducing conditions. The majority of GAMA wells sampled were between 30 and 120 m depth. Approximately 15% of wells were between 120 and 400 m depth (Dawson et al., 2008).

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metals in soils after the addition of ammonium fertilizers (Tsadilas et al., 2005; Liu et al., 2007) and KCl (Liu et al., 2007). Unlike some other trace metal cations, Cr(III) can be oxidized by Mn oxides before it re-partitions into the solid phase. Thus, fertilizer applications may have a much more pronounced effect on Cr mobility and toxicity than on other geogenic trace elements in the soil (e.g. Ni, Co). The large initial increase in Mg plus Ca and Cr(VI) during the first day of the incubation with 70 mmole H+ kg1 (Fig. 10b and c) indicates that both cation exchange and Cr oxidation processes are quite rapid. The continued generation of Cr(VI) throughout the 14 day incubation despite no measurable increase in water leachable Ca plus Mg may have been due to the increased activity of base cations in the soil or the adsorption of H+ as the pH slowly increased. It is also possible that the continued Cr(VI) production was due to the dissolution of Cr(III) from chromite. Oze et al. (2007) showed that the dissolution of chromite and subsequent oxidation of Cr(III) on Mn oxides is a viable mechanism for Cr(VI) production in soils and sediments at pH 6 7. Using the rate equation Oze et al. (2007) derived from experimental data at pH = 7, up to 1.6 lg Cr(VI) per kg soil could be generated from chromite dissolution (assumptions: 560 mg kg1 CrT in soil is all chromite, chromite is 34 wt.% Cr, chromite density is 4.5 g cm3, chromite particles are spheres with an average diameter of 50 lm). This result is similar to the Cr(VI) generation rate of 1.7 lg kg1 day1 observed during the last four time points of the incubation in the current study (Fig. 10c).

6. Conclusions Ultramafic inputs into the Sacramento Valley alluvium have resulted in soils and aquifer materials that are extremely elevated in total Cr compared to average continental abundances. Much of the Cr is present in the refractory mineral chromite, but an appreciable amount (on the order of 100 mg per kg soil) is associated with Fe oxides and clay minerals. At the typical alkaline pH of the valley soils and alluvium, it is this latter, more labile Cr that is more likely to be oxidized to Cr(VI) by abundant Mn oxides. Although Cr(VI) represents a small fraction of the total Cr in the soil, concentrations are high enough to impact ground water quality. The rate of Cr(VI) generation in laboratory soil incubations as well as the shape of the Cr(VI) depth profiles in drill cores suggests that Cr(VI) could undergo an annual cycle of production in near surface soils and sediments followed by leaching into groundwater. Thus, increased leaching of soils and vadose zone sediments due to irrigation could exacerbate natural Cr(VI) groundwater enrichment. Increased Cr(VI) production was observed in incubations in which varying amounts of H+ was added to valley soil. Very rapid Cr(VI) production and concomitant release of water leachable Ca2+ and Mg2+ ions during the incubations suggest that the major mechanism of increased Cr(VI) production was the exchange of H+ for sorbed Cr(III) and the subsequent oxidation of the Cr(III) by Mn oxides. It is also possible that Cr(III) from enhanced chromite dissolution at lowered soil pH was also oxidized to Cr(VI). These results suggest that cationic and acid-producing constituents in irrigation return water, such as ammonium, may accelerate Cr(VI) production rates in the valley soils and alluvial sediments. Overall the results of this study suggest that Cr(VI) is generated naturally in Sacramento Valley soils and sediments. However, the generation rate and leaching of Cr(VI) to the groundwater may be influenced by human activities such as agriculture. The mechanisms outlined in this study may be important to consider for the protection of groundwater resources in areas where natural Cr(VI) enrichment is a concern.

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Acknowledgments This work was supported by a US Geological Survey Mendenhall Postdoctoral Fellowship to CTM. We thank Sue Fields and Andrew Fulks (University of California, Davis), Yolo County Parks Department, and anonymous landowners for providing access to sampling sites and samples; Chuck Frey (Brown and Caldwell) for assistance with geophysical logging of boreholes; JoAnn Holloway, Todd Hoefen and Monique Adams (USGS) for assistance with field collections and laboratory analyses; Rich Wanty and John Izbicki (USGS) for insightful comments on previous versions of this manuscript; Clemens Reimann for editorial handling and two anonymous reviewers for comments which substantially improved the manuscript. Any use of trade, product, or firm names is for descriptive purposes only and does not imply endorsement by the US Government. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.apgeochem.2011.05.023. References Ball, J.W., Izbicki, J.A., 2004. Occurrence of hexavalent chromium in ground water in the western Mojave Desert, California. Appl. Geochem. 19, 1123–1135. Ball, J.W., Nordstrom, D.K., 1998. Critical evaluation and selection of standard state thermodynamic properties for chromium metal and its aqueous ions, hydrolysis species, oxides, and hydroxides. J. Chem. Eng. Data 43, 895–918. Banks, M.K., Schwab, A.P., Henderson, C., 2006. Leaching and reduction of chromium in soil as affected by soil organic content and plants. Chemosphere 62, 255–264. Bartlett, R.J., 1991. Chromium cycling in soils and water: links, gaps, and methods. Environ. Health Perspect. 92, 17–24. Bartlett, R., James, B., 1979. Behavior of chromium in soils: III Oxidation. J. Environ. Qual. 8, 31–35. Bartlett, R., Kimble, J.M., 1976. Behavior of chromium in soils: II Hexavalent forms. J. Environ. Qual. 5, 383–386. Becquer, T., Quantin, C., Sicot, M., Boudot, J.P., 2003. Chromium availability in ultramafic soils from New Caledonia. Sci. Total Environ. 301, 251–261. Bertoldi, G.L., Johnston, R.H., Evenson, K.D., 1991. Ground water in the Central Valley, California – a summary report. US Geol. Surv. Prof. Paper 1401-A. Bradl, H.B., 2004. Adsorption of heavy metal ions on soils and soils constituents. J. Colloid Interface Sci. 277, 1–18. California’s Groundwater Bulletin 118. Sacramento Groundwater Basin, Yolo Subbasin. Last updated 27.02.04. (accessed 09.03.10). Chung, J.-B., Burau, R.G., Zasoski, R.J., 2001. Chromate generation by chromate depleted subsurface materials. Water Air Soil Pollut. 128, 407–417. Cooper, G.R.C., 2002. Oxidation and toxicity of chromium in ultramafic soils in Zimbabwe. Appl. Geochem. 17, 981–986. Dai, R., Liu, J., Yu, C., Sun, R., Lan, Y., Mao, J.D., 2009. A comparative study of oxidation of Cr(III) in aqueous ions, complex ions and insoluble compounds by manganese-bearing mineral (birnessite). Chemosphere 76, 536–541. Davisson, M.L., Criss, R.E., 1993. Stable isotope imaging of a dynamic groundwater system in the southwestern Sacramento Valley, California, USA. J. Hydrol. 144, 213–246. Dawson, B.J.M., Bennett, G.L., Belitz, K., 2008. Ground-water quality data in the Southern Sacramento Valley, California, 2005 – results from the California GAMA program. US Geol. Surv. Data Ser. 285. Deverel, R.J., Gilliom, R.J., Fujii, R., Izbicki, J.A., Fields, J.C., 1984. Areal distribution of selenium and other inorganic constituents in shallow ground water of the San Luis drain service area, San Joaquin Valley, California: a preliminary study. US Geol. Surv. Water Resour. Invest. Rep. 84-4319. Eary, L.E., Rai, D., 1987. Kinetics of chromium(III) oxidation to chromium(VI) by reaction with manganese dioxide. Environ. Sci. Technol. 21, 1187–1193. Fandeur, D., Juillot, F., Morin, G., Olivi, L., Cognigni, A., Webb, S.M., Ambrosi, J.-P., Fritsch, E., Guyot, F.O., Brown, J.G.E., 2009. XANES evidence for oxidation of Cr(III) to Cr(VI) by Mn-oxides in a lateritic regolith developed on serpentinized ultramafic rocks of New Caledonia. Environ. Sci. Technol. 43, 7384–7390. Fantoni, D., Brozzo, G., Canepa, M., Cipolli, F., Marini, L., Ottonello, G., Zuccolini, M., 2002. Natural hexavalent chromium in groundwaters interacting with ophiolitic rocks. Environ. Geol. 42, 871–882. Fendorf, S., Zasoski, R.J., 1992. Chromium(III) oxidation by d-MnO2. 1. Characterization. Environ. Sci. Technol. 26, 79–85. Garnier, J., Quantin, C., Martins, E.S., Becquer, T., 2006. Solid speciation and availability of chromium in ultramafic soils from Niquelândia, Brazil. J. Geochem. Explor. 88, 206–209.

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