Quantifying Cr(VI) Production and Export from Serpentine Soil of the

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Quantifying Cr(VI) production and export from serpentine soil of the California Coast Range Cynthia N. McClain, Scott Fendorf, Samuel M. Webb, and Kate Maher Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b03484 • Publication Date (Web): 22 Nov 2016 Downloaded from http://pubs.acs.org on November 27, 2016

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Environmental Science & Technology

Quantifying Cr(VI) production and export from serpentine soil of the California Coast Range Cynthia N. McClaina,*, Scott Fendorfb, Samuel M. Webbc, Kate Mahera,* a b

Department of Geological Sciences, Stanford University, Stanford, CA, 94305, USA Department of Earth System Science, Stanford University, Stanford, CA, 94305, USA c Stanford Synchrotron Radiation Lightsource, Menlo Park, CA 94025, USA * Corresponding author E-mail: [email protected], Phone: +1-650-223-5790 * Corresponding author E-mail: [email protected], Phone: +1-650-725-0927

ABSTRACT

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Hexavalent chromium (Cr(VI)) is generated in serpentine soils and exported to

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surface and groundwaters at levels above health-based drinking water standards.

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Although Cr(VI) concentrations are elevated in serpentine soil pore water, few studies

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have reported field evidence documenting Cr(VI) production rates and fluxes that govern

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Cr(VI) transport from soil to water sources. We report Cr speciation (i) in four serpentine

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soil depth profiles derived from the California Coast Range serpentinite belt and, (ii) in

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local surface waters. Within soils, we detected Cr(VI) in the same horizons where Cr(III)-

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minerals are co-located with biogenic Mn(III/IV)-oxides, suggesting Cr(VI) generation

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through oxidation by Mn-oxides. Water-extractable Cr(VI) concentrations increase with

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depth constituting a 7.8 to 12

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produced at a rate of 0.3 to 4.8 kg Cr(VI)/km2/yr and subsequently flushed from soil

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during water infiltration, exporting 0.01 to 3.9 kg Cr(VI)/km2/yr at concentrations

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ranging from 25 to 172 µg/L. Although soil-derived Cr(VI) is leached from soil at

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concentrations exceeding 10 µg/L, due to reduction and dilution during transport to

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streams, Cr(VI) levels measured in local surface waters largely remain below California’s

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drinking water limit.

kg/km2 reservoir of Cr(VI) in soil. Here, Cr(VI) is

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TOC

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INTRODUCTION

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Surface and groundwater are contaminated by Cr(VI) at levels above drinking

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water limits (e.g. World Health Organization – 50 µg Cr/L, California – 10 µg Cr(VI)/L)

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as a result of geogenic Cr(III) mineral oxidation.1,2 In the United States, 0.92 to 74

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million people are exposed to hazardous Cr(VI) concentrations (> 10 µg/L) in drinking

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water,3,4 which poses a human health threat through an increased risk for stomach and

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intestinal cancer.5,6

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Due to the high Cr content, ultramafic rock (2400 mg/kg),7 their metamorphic

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derivatives (serpentinite), and weathering products (serpentine soil and sediments) are the

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most important sources of Cr(VI) produced from geogenic sources within drinking water.

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Serpentine soils develop during weathering of ultramafic bedrock, creating ideal

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biogeochemical conditions for Cr(VI) generation, accumulation, and export. Chromium

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minerals accumulate in soil, increasing Cr concentrations (29 to 80000 mg/kg),8,9 with

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coinciding increases in Mn concentrations resulting from weathering of ultramafic 2 ACS Paragon Plus Environment

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rocks.10,11 Chromium(III)-bearing primary minerals within weathered ultramafic rock

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have morphological features characteristic of dissolution, suggesting Cr(III) release.12,13

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Chromium(VI) is thought to be generated in soil via oxidation of Cr(III) by proximal

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Mn(III/IV)-oxide minerals,14-16 reaching concentrations up to ~ 7000 µg/kg.16-19 As a

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consequence, Mn oxidizing bacteria can accelerate the Cr(VI) generation rate through the

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production of the primary Cr(III) oxidant.20 Cr(VI) production21 rates vary widely

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depending on Cr(III) source (e.g. silicates, spinels, FexCr1-x(OH)3, aqueous), oxidation

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pathway (e.g. different Mn-oxide minerals with varying Mn(III/IV) content, O2, reactive

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oxygen species), and pH.1,22-29 Countering production, Cr(VI) can be attenuated by

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reduction (e.g. by Fe(II), sulfide, organic matter, microbes) and adsorption onto mineral

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surfaces.30,31 As a result of the complexity underlying Cr release and oxidation rates, the

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rates associated with natural systems have not yet been quantified.

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Serpentine soils have shown Cr levels in soil water that exceed health based

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drinking water standards, ranging from 5 – 250 µg/L. 32-35 For comparison, surface waters

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in ultramafic catchments have Cr(VI) concentrations up to 19 µg/L.2,33,36 Due to the high

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Cr concentration in soil water, soil-derived Cr(VI) may be an important source of Cr(VI)

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to surface and groundwater water. However, the inventory of Cr(VI) within serpentine

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soils, along with the rates of production and export, are largely unresolved.

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California, USA, is a region with (i) abundant ultramafic bedrock (2860 km2) and

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soils (Figure S1),37 and (ii) elevated Cr(VI) levels in surface and groundwater.2,33,38-42 In

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California, 0.1 to 7.8 million individuals may be exposed to hazardous concentrations of

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Cr(VI) and other trace elements in drinking water.3,43 Here, we studied soil derived from

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the northern California Coast Range serpentinite belt. Our objectives were (i) to quantify



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the Cr(VI) reservoir, production rate and export fluxes from serpentine soil, and (ii) to

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compare these fluxes to Cr(VI) fluxes in local surface water. We collected depth resolved

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soil samples from four profiles and water samples, with concurrent discharge rates, from

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a nearby creek. We determined physical and mineralogical properties of soil, chemical

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composition and speciation of soil solids and water extracts (e.g. Cr(VI), Cr(III)), centi-

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to micro-meter scale imaging of the spatial distribution and oxidation state of Cr and Mn

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minerals, and tested for Mn oxidizing bacteria. Based on these results, we evaluate

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approaches for determining in situ Cr(VI) production rates. Our results provide a basis for

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understanding the biogeochemistry of Cr(VI) generation in the context of the hydrologic

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processes governing Cr(VI) release from serpentine soils.

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MATERIALS AND METHODS

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Study Area. Serpentine soils, derived from the California Coast Range

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serpentinite belt, are abundant in the Putah Creek/Berryessa watershed of northern

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California, and constitute a potential source of Cr(VI) contamination to drinking water

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(Figure S1). The major surface water reservoir in the watershed, Lake Berryessa, serves

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as a drinking water source for more than 0.35 million people (Vallejo, Fairfield,

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Vacaville, Suisun City) and supplies groundwater recharge (natural and irrigation return)

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in the southwestern Sacramento Valley where groundwater is used for drinking (e.g.

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Davis). Our field site lies near the headwaters of the Putah Creek/Berryessa watershed, in

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upper Hunting Creek watershed (16.8 km2), within University of California’s

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McLaughlin Natural Reserve (38.86, -122.404 [WGS 84]) (Figure S1).


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At our field site, Montara series Mollisols predominate, which are well-drained

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soils derived from the Berryessa serpentinite complex.44,45 Grasses and manzanita

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(Arctostaphylos viscida) proliferate on southwest facing slopes.

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Precipitation in the Mediterranean climate (Xeric moisture regime, mean annual

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temperature of 15°C, mean annual precipitation of 720 mm),46 during winter months

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leads to through-flow and runoff, contributing to groundwater recharge and surface water

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flow. Chromium concentrations in Hunting Creek and its tributaries range from below

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detection (< 0.3 µg/L) to 19 µg/L.2

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Soil Sample Collection. Field-moist soil samples were collected every 5-10 cm

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with depth from 4 profiles (numbered 1-4) along a toposequence on the shoulder of the

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meadow by hand digging soil pits or hand augering until refusal depth ~ 60 cm (bedrock

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interface), and stored in airtight vials at 4°C prior to laboratory analysis to minimize

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oxygen exposure. Bulk density samples were collected using the short core method

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(Table S1).47 Mollisol profiles typically have brown A, and Bt horizons (0-32 cm)

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underlain by olive colored saprolite (C horizon, 32-60 cm) (Table S2). Bioturbation from

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rooting of vegetation, and burrowing gophers was limited to < 50 cm. Saturated hydraulic

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conductivity was measured at multiple depths in soil profiles using a Guelph

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Permeameter, operating under constant head conditions.48

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Water Sample Collection. Water samples were collected during a rainstorm on

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April 1, 2014 from Hunting Creek, tributaries draining the meadow where soil samples

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were collected, and from overland flow on the meadow. Water samples were filtered

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using Geotech in-line high capacity 0.45 µm cartridges, split into two subsamples

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(unacidified and acidified to pH < 2 with nitric acid), and stored at 4°C prior to laboratory



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analysis. Volumetric flow rates were measured by the velocity-area method. Watershed

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area was calculated using Google Earth’s polygon tool to determine contributing area

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above each sampling location, delineated by topographic ridges. Stream-based specific

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discharge (runoff) was calculated by normalizing the measured volumetric discharge to

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watershed area.

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Solid-Phase Analysis and Sample Preparation. Gravimetric water content was

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determined by difference between field moist and oven dry soil (104°C). Particle size

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distribution (< 2 mm) and soil texture were determined at the University of Idaho by

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sieving and the pipette method (Table S3). X-ray diffraction (XRD) was performed on

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powdered bulk soil samples and clay separates as described in Supporting Information

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(Figures S2, S3, Table S4). Thin sections were prepared from oven dried soil aggregates

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embedded in Epotek 301 2-FL epoxy resin, that were cut and polished to 30 µm using

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heat and oxygen sensitive techniques by Spectrum Petrographics (Vancouver, WA).

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Mineral modal abundance was determined by optical petrography and point counting of

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thin sections on a Nikon Optiphot polarizing petrographic microscope (Table S5).

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Regolith is composed of serpentine minerals (lizardite, antigorite, clinochrysotile), clay

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(smectite, vermiculite, chlorite), actinolite, and oxide minerals (e.g. Fe-Cr-spinels,

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secondary Mn-Fe-oxides) (Figure S4). Thin section imaging and determination of Fe,

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Mn, Ni and Cr content of minerals was performed by Electron Microprobe Analysis

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(EMPA) using a JEOL JXA-8230 SuperProbe. Oven dried soil samples were also

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disaggregated by gentle grinding with an agate mortar and pestle and analyzed for bulk

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major and trace element content by X-ray fluorescence spectrometry (Spectro XRF-

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XEPOS).


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Extractions. Extractions of field moist soil samples were performed to quantify

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major and trace element content of two pools: “water-exchangeable” and “dithionite-

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extractable” (nominally from Fe- and Mn-oxides [citrate-bicarbonate-dithionite –

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CBD]).49 Extractions for the “water-exchangeable” pool were performed for 24 h with 30

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g of soil and 30 g of deionized water, vacuum filtered through Millipore 0.2 µm Steriflip

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filters, split into two subsamples (unacidified and acidified to pH < 2 with nitric acid),

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and stored at 4°C prior to chemical analysis. Concentrations are reported in mg or µg per

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kg of dry soil. Mean normalized concentrations for profiles i = 1 – 4 are expressed as:

1 = 4

(1)

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where X is concentration in mg/kg measured at each sampling depth (z) or at the deepest

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sample from near the saprolite-bedrock interface. Chromium(VI) concentrations in soil

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water (c [µg Cr(VI)/L of soil water]) are estimated from water-exchangeable Cr(VI)

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concentrations (c [µg Cr(VI)/kg of dry soil]), using measured gravimetric water content

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(w [g water/g dry soil]) and water density (ρw [g/cm3) according to: c = C × ρw / w.

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Aqueous Chemical Analysis. Acidified water samples and extracts were

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analyzed for major and trace element content with inductively coupled plasma optical

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emission spectroscopy (Thermo Scientific ICAP 6300 Duo View ICP-OES) and mass

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spectroscopy (Thermo Scientific XSERIES 2 ICP-MS), respectively. Unacidified water

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samples and water-exchangeable extracts were measured for pH (ranging from 7.2 to

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8.3), inorganic anions by ion chromatography (Dionex DX-500), alkalinity by titration

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with 0.04 M sulfuric acid to a colorimetric endpoint (pH 4.5) using bromocresol green-

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methyl red indicator, and dissolved organic carbon (DOC, measured as non-purgeable



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organic carbon) on a Shimadzu TOC-L. Chromium(VI) concentrations were determined

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by

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spectrophotometer.14,50 Chromium(III) concentrations were calculated by difference

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between Cr(TOT), measured by ICP-MS, and Cr(VI).

a

modified

s-diphenyl

carbazide

method

on

a

Shimadzu

UV-1601

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Microbiological Analysis. To enrich Mn oxidizing bacteria from field moist soils

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(from 0-4 cm depth), serial dilutions (1/104) in artificial seawater media were plated on

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agar-solidified K-media containing 20 mM HEPES buffer and 100 µM MnCl2 (pH 7.6).

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Mn(III/IV)-oxides were confirmed with the Leucoberbelin Blue (LBB) colorimetric

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assay.51

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Synchrotron Spectroscopy and Element Mapping. Synchrotron-based X-ray

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microprobe measurements of element distributions and oxidation state in regolith thin

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sections from profile 1 were conducted on beam lines 10-2 and 2-3 at the Stanford

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Synchrotron Radiation Lightsource as described in Supporting Information. Element

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maps were generated by rastering the sample around the X-ray beam, collecting X-ray

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fluorescence data, and using the Microanalysis Toolkit for processing.52 Each micro X-

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ray absorption near edge structure (µXANES) spectrum was collected from 200 eV

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below the Mn or Cr K-edge to 400 eV above the edge. All spectra were background

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subtracted and normalized to a unit step edge with the iXAFS Athena software package.53

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Soil Hydraulic Properties. To calculate the flux of Cr(VI) under unsaturated

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conditions, we use a special case of Darcy’s law; the infiltration flux, is equal to the

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unsaturated hydraulic conductivity, assuming a unit hydraulic gradient that may result

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from either constant pressure head or local extreme of pressure head within a vertical

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section of soil.54 Unsaturated hydraulic conductivity was calculated from measured


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saturated hydraulic conductivity, using the Van Genuchten equation (1980) as described

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in Supplemental Information.55

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RESULTS

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Trends in Major and Trace Element Chemistry of Serpentine Regolith. The

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vertical distribution of major and trace elements systematically varies with depth in

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relation to soil horizons and mineralogy. The saprolite is characterized by low total,

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dithionite-extractable and water-exchangeable concentrations of major and trace elements

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(Figure 1). In contrast, the soil, which has more abundant clay and Fe/Mn-oxides (Table

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S5), is enriched in Al, Fe, Mn and Cr. Additionally, the concentrations of water-

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exchangeable trace elements are highest in the soil (Figure 2).



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Figure 1. Mean normalized concentration () of major (Si, Mg, Ca, Al) and trace (Fe, Mn, Cr, Ni) elements from 4 serpentinite regolith profiles as a function of depth, normalized to the concentration in each extraction at the saprolite/bedrock interface. (a, d) Total (XRF), (b, e) dithionite-extractable (CBD), (c, f) water-extractable. Filled circles indicate sampling locations and are shown only on Si and Fe profiles for clarity.

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Chromium Concentrations. Chromium concentrations are 1.2 to 3 times higher

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in soil compared to saprolite, across all Cr reservoirs (Figures 1 and 2). Approximately

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4% of total Cr resides in the dithionite-extractable fraction (decreasing with depth), and

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0.001% is water-extractable. Water-extractable Cr is mainly Cr(VI), with maximum 10 ACS Paragon Plus Environment

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Cr(VI) levels in soil of 27 µg/kg, compared to 12 µg/kg in saprolite from 50-60 cm depth.

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Water-exchangeable Cr(III) concentrations increase (up to 15 µg/kg) in near surface soils

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(5 cm depth), coinciding with high DOC and alkalinity. However, in two profiles, water-

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exchangeable Cr(III) levels are highest (21 and 57 µg/kg) near the soil-saprolite interface

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(~ 40 cm). Water samples collected during a rainstorm from both overland flow in the

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meadow, and surface water runoff have a narrow range of Cr(VI) concentrations from 5.5

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to 6.9 µg/L, and low Cr(III) concentrations (< 0.6 µg/L).

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Figure 2. Water-extractable concentrations of (a) Cr(VI), (b) Cr(III), (c) dissolved organic carbon (DOC) and, (d) alkalinity per kg dry soil from 4 serpentinite regolith depth profiles (filled circles). Mean concentrations shown as black squares with error bars (1 SD). Occurrence and Oxidation State of Cr and Mn in Regolith Minerals at the

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Micrometer Scale. Chromium(III) and Mn(III/IV) minerals are juxtaposed, residing

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within microns of each other in all soil and saprolite samples (Figures 3 and S4).

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Chromium(III)-bearing minerals are present both as small diameter (50 µm), etched and

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embayed, high Cr concentration minerals (e.g. Cr- and Fe-spinels such as chromite

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[spectrum 4 in Figure 3] and magnetite), and as more diffuse sources (e.g. serpentine



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minerals [spectrum 2 in Figure 3], Fe(Cr)-oxides, and clay minerals) (Table S5).

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Although water-extractable Cr(VI) was measured in soil, it resided at levels below

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detection by our µXANES spectroscopic analysis (Figure 3). Manganese oxides and Mn-

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bearing silicates in serpentine the regolith have varying oxidation state, based on the

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position of the Mn K-edge in the energy range from 6547 to 6552 eV.

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Manganese(III/IV)-oxides similar to birnessite are found coating mineral grains (spectra

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7 and 12) and as < 2 µm clusters (spectrum 9), both in close proximity to Cr(III)

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minerals. Additionally, Mn oxidizing bacteria were detected in surface soil (Figure S5).

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Figure 3. X-ray fluorescence maps showing relative intensity of Cr (magenta), Mn (yellow), and Fe (cyan) in serpentinite regolith (profile 1) from (a) 4 cm depth, (b, d) 26 cm depth, and (c) 39 cm depth. Representative, normalized µXANES spectra of reference minerals and from locations numbered in panels a-d for (e) Cr K-edge and (f) Mn K-edge. Shaded gray regions in panel f indicate energies characteristic of Mn(II) (6547-6553 eV), Mn(III) (6554-6559 eV) and Mn(IV) (6560-6562 eV).


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DISCUSSION

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Cr(VI) Generation in Serpentine Soil. Chromium(VI) is produced in serpentine

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soil as a result of both long-term (100’s to 1000’s of years) soil formation that

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concentrates and redistributes Cr and Mn, and contemporary biogeochemical reactions

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that produce Cr(VI). Vertical and lateral transport of solutes further redistribute reaction

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products, with a dependence on landscape position, soil hydraulic properties, and

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precipitation (rainfall) intensity. Figure 4 shows our conceptual model of geochemical

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reactions and movement of water through serpentine soil that results in Cr(VI)

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production, accumulation, and export.

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Figure 4. Conceptual cycle of Cr(VI) production and export from serpentinite regolith. 13 ACS Paragon Plus Environment


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During pedogenesis Cr-bearing primary minerals (e.g. spinels and serpentine)

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dissolve (Figure S4), repartitioning ~ 4% of Cr into secondary minerals resulting in Cr

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accumulation in clay and oxyhydroxides (Tables S1c and S6), as also noted by Oze et al.

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(2004) and Morrison et al. (2015).13,56 The patterns of Cr and Mn partitioning we

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observed are comparable to other studies in temperate climates where ~ 10 % of Cr was

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associated with Fe-oxides,57,58 and Cr repartitioned to oxyhydroxides in the B

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horizon.10,11 Some Cr will also be lost to the fluid phase during weathering.9,59 However,

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the majority of Cr (~ 96%) remains in low solubility primary minerals that preferentially

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accumulate as the surrounding, more easily weathered, soil matrix collapses. In contrast

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to Cr, more Mn (> 10%) in soil is hosted in secondary minerals (e.g. oxides), having been

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redistributed from primary minerals during soil formation. Given the enrichment of Mn

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oxidizing bacteria from surface soil (0-4 cm depth, [Figure S5]), neoformed Mn-oxides

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may be biologically promoted, giving rise to highly reactive Mn-oxide phases.60 Mixed

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Mn oxidation state (III/IV) provides evidence of Mn redox cycling,61 whereby Mn-oxide

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surfaces can be regenerated with Mn2+, a byproduct of Cr(III) oxidation. Neoformed

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Mn(III/IV) minerals are co-located (within microns) with etched and pitted Cr(III)-

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bearing spinel and silicate minerals, and at various locations, they surround Cr(III)-

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bearing mineral surfaces (Figures 3 and S4). Within soil, water-extractable Cr(VI)

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concentrations are also high (Figure 2). Together, these observations support the potential

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for continual dissolution of Cr(III) from oxides and silicates and subsequent Cr(III)

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oxidation by biogenic Mn(III/IV)-oxides in serpentine soils developed in temperate

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climates, similar to previous studies in tropical climates,15,16 and in the laboratory.20,24-26


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As these soils have low iron oxide content and alkaline pH values, little Cr(VI) will

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adsorb to soil matrix, and Cr(VI) will be readily transported vertically to saprolite and

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groundwater, and laterally to surface water.

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Coupled biogeochemical and hydrologic processes generate depth gradients in

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water-exchangable Cr. Water-exchangeable Cr concentrations are highest in the B

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horizon, where Cr(VI) and Cr(III) concentrations reach up to 25 and 15 µg/kg,

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respectively (Figure 2). Here, total Cr and Mn, and dithionite extractable Cr levels are

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highest. In this horizon, dissolved organic acids (e.g. acetate, citrate) and increased

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carbon dioxide fugacity may enhance Cr(III) mineral dissolution,62,63 stabilize Cr(III) in

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solution,64,65 or enhance Cr(VI) mobility due to decreased Cr(VI) sorption in the presence

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of carbonate.66,67 Near the soil-saprolite interface (40 cm), anomalously high Cr(III)

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concentrations may result from heterogeneity or localized reducing conditions near

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manzanita roots. In the saprolite (from 50 to 60 cm), Cr(VI) concentrations may be lower

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due to lower total Cr concentrations, Cr(VI) reduction in areas with limited oxygen

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supply, or from flushing of Cr(VI) due to preferential lateral flow above the saprolite-

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bedrock interface.

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Calculated Cr(VI) concentrations in soil pore water under unsaturated conditions

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(c [µg Cr(VI)/L of soil water]) all exceed California’s drinking water limit of 10 µg/L,

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ranging from 25 to 172 µg/L (Table S7). These calculations, based on water extractions

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performed in the laboratory, may overestimate field measurements due to lower

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soil/water ratios, different water chemistry, or release of exchangeable Cr(VI) from pores

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and surfaces that may not otherwise be accessible. Nevertheless, calculated Cr(VI)

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concentrations in soil water are similar to measurements of dissolved plus fine colloids



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(filtered to < 0.45 µm) in soil water collected by tension-free lysimetry by Gasser et al.

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(1994) (9-104 µg/L).32 This suggests that calculating soil water Cr(VI) concentrations

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from measured water content and Cr(VI) in water extractions and may be a good proxy

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for in-situ soil water Cr(VI) levels when field measurements are not available.

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Importantly, Cr may also be present as colloids,32 however in this study we focus on

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dissolved Cr(VI) and use calculated soil water Cr(VI) concentrations to estimate Cr(VI)

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fluxes and export rates from serpentine soils.

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Cr(VI) Accumulation in Serpentine Soil. Over the top 40 cm of the regolith, we

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integrate the profile of water-extractable Cr(VI) concentrations per mass of soil (C

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[µg/kg]) with depth (z [m]) (Figure 2), estimating the inventory of water-extractable

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Cr(VI) to be 7.8 to 12 kg/km2 (I = C × z × ρb [kg Cr(VI)/km2], where ρb is bulk density

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[g/cm3]) (Figure 4, Tables S1 and S7). This Cr(VI) reservoir is ~ 16 times smaller than in

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unsaturated ultramafic-bearing sediments of southwest Sacramento Valley, California

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because serpentine soils are shallower than the valley’s unsaturated zone.68 In these

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California Coast Range serpentine soils, Cr(VI) only makes up ~ 0.001% of the total Cr

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reservoir in soil (1.1 × 106 kg/km2), yet contributes to geogenic Cr(VI) contamination of

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surface water and groundwater during infiltration, runoff and recharge.

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Cr(VI) Flux through Serpentine Soil. During infiltration through unsaturated

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soil Cr(VI) will be transported downward according to the water infiltration flux (q

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[mm/yr], here assumed to equal unsaturated hydraulic conductivity), at Cr(VI)

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concentrations in soil water (c [µg Cr(VI)/L of soil water]) (Tables S1 and S7).

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Infiltration fluxes calculated here for unsaturated serpentine soil are spatially variable,

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ranging from 0.03 to 65 mm/yr, with a mean travel time (τ ) of 7.3 years through the top 16 ACS Paragon Plus Environment

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40 cm of soil (τ = θ × z ⁄ q, [yr], volumetric water content is θ = w × ρb / ρw, where ρw is

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water density (g/cm3)) (Figure 4, Table S7). These water fluxes have large uncertainties,

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primarily due to the variability in saturated hydraulic conductivity measurements, which

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typically have uncertainties of up to two orders of magnitude.69 However, because

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hydraulic conductivity can vary by orders of magnitude with depth and along the

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hillslopes (where we measured),54,70 these calculated infiltration fluxes are thought to be

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representative of natural variability in this meadow. We estimate that the resulting flux, J,

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of Cr(VI) (J = q × c [kg/km2/yr]) through meadow soil to surface or groundwater is

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spatially variable, ranging from 0.01 to 3.9 kg/km2/yr, with a mean of 1.1 kg/km2/yr

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(Figure 4, Table S7). This flux provides a minimum constraint on the long-term average

329

Cr(VI) production rate.

330

Cr(VI) Production Rate. To our knowledge, no field-estimates of Cr(VI)

331

production in soils have been developed. Such estimates are important for evaluating the

332

role of soils and unsaturated zones in contributing to Cr(VI) contamination of

333

groundwaters and surface waters. Here, we calculate the net rate of Cr(VI) generation in

334

three different ways: we estimate it using two different mass balance approaches that

335

require knowledge of the infiltration flux, and then compare these results to a laboratory-

336

determined rate law for Fe-Cr-spinel oxidation by birnessite that does not depend on the

337

infiltration flux.1 First, we simulate the top 40 cm of each profile as a completely mixed

338

flow reactor (CMFR), assuming the system has spatially uniform properties and the only

339

source of Cr(VI) to the system is in situ generation, due to the low Cr(VI) concentrations

340

in rainwater.71 The net in situ Cr(VI) production rate (R [µg/L/yr]) can be expressed as:



=

1 + "#

Page 18 of 29

(2)

341

where m is the water to soil ratio (w × ρb × z /ρw), I (kg/km2) is the amount of Cr(VI) that

342

accumulates in soil during the time (τ [yr]) it takes water to travel through soil, J is the

343

Cr(VI) export flux (kg/km2/yr) (Tables S1 and S7). We estimate the Cr(VI) production

344

rate in serpentine soil to range from 1.2 to 17.3 µg/L/yr, which is equivalent to 0.19 to

345

2.35 kg/km2/yr using the conversion factor m (Table S7). This CMFR model is a

346

reasonable representation of the relatively constant Cr(VI) concentration with depth for

347

profiles 3 and 4 (located farther downslope than profiles 1 and 2), where the assumption

348

of spatially uniform properties may be valid (Figure S6). This homogeneity may originate

349

from the lower infiltration fluxes and/or greater mixing due to complex, two-dimensional,

350

water and soil movement (e.g. eluviation, creep, deposition) down the hillslope.

351

Second, we simulate profiles 1 and 2 as plug flow reactors (PFRs), which is

352

consistent with Cr(VI) concentrations that increase with depth. The change in Cr(VI)

353

concentration in soil water with depth (c(z) [µg/L]) during steady-state vertical advective

354

transport, with Cr(VI) generation described by zeroth-order kinetics (typical for surface

355

catalyzed reactions) can be expressed as: $%&' = $ +

(& )

(3)

356

where co is Cr(VI) concentration at the ground surface (25 µg/L) (equation development

357

and assumptions described in Supporting Information). Using equation (2) to fit our

358

Cr(VI) concentration data with depth by adjusting R, the best fits are Cr(VI) production

359

rates of 21 and 30 µg/L/yr (2.92 and 2.96 kg/km2/yr) for profiles 1 and 2, respectively,

360

assuming mean θ and q for each profile (Figure S6a, Table S7). The PFR model provides 18 ACS Paragon Plus Environment

Page 19 of 29


361

a better fit than the CMFR for profiles 1 and 2 (located closer to the summit) suggesting

362

that depth gradients in concentration arise from the balance between net Cr(VI)

363

production and net downward transport due to infiltration, as per equation (3) above. The

364

fit of two different mass balance models for Cr(VI) concentration in soil water (CMFR,

365

PFR) reveals how coupled hydrologic and geochemical processes influence the depth

366

distribution of Cr(VI) at different landscape positions along a hillslope.

367

Third, the Cr(VI) production rate from oxidation of Fe-spinel (the mineral that

368

hosts the most Cr by mass [Table S6]) by the Mn(IV)-oxide birnessite, parameterized by

369

data measured at our field site, can be described after Oze et al. (2007) by:1 :. 99% of Cr is immobile, mineral-bound Cr(III). Soil Cr(III)-minerals, while

436

limited in solubility, constitute a seemingly infinite source of Cr that can be oxidized to

437

Cr(VI) and transmitted to the hydrosphere, given the proper conditions. We show that

438

Cr(III)-minerals reside proximal to Mn(III/IV)-oxide minerals, providing a pathway for

439

generating a reservoir of Cr(VI). Although it constitutes < 0.001% of the Cr reservoir in

440

serpentine soils, Cr(VI) is flushed to surface and groundwater at concentrations above 10

441

µg/L during infiltration. However, soil-derived Cr(VI) is attenuated by dilution and/or

442

reduction before reaching surface water ways, leading to Cr(VI) concentrations largely 22 ACS Paragon Plus Environment

Page 22 of 29

Page 23 of 29


443

below California’s drinking water limit in surface water of the Berryessa/Putah Creek

444

watershed.2 Nevertheless, Cr removed by reduction in streambed sediments may be

445

transported downstream, loading depositional environments such as alluvial plains with

446

Cr(III)-phases susceptible to re-oxidation. This study summarizes the size of water-

447

extractable Cr(VI) reservoir in serpentine soil (~10 kg/km2), provides the first estimate of

448

in situ Cr(VI) production rate (~2.5 kg/km2/yr), and constrains the amount of Cr(VI)

449

released from soil during infiltration (~1.1 kg/km2/yr), relative to Cr(VI) levels measured

450

in surface water used for drinking.

451 452

ASSOCIATED CONTENT

453

Supporting Information. Additional method and equation details; list of abbreviations,

454

symbols, equations, and parameters; tables; figures. This information is available free of

455

charge via the Internet at http://pubs.acs.org/.

456 457

ACKNOWLEDGEMENTS

458

This work was supported by the U.S. National Science Foundation (Graduate Research

459

Fellowship DGE-114747 to CNM, and EAR-1254156 to KM), and Stanford University’s

460

McGee Grant, and conducted at the University of California’s Donald and Sylvia

461

McLaughlin Natural Reserve. We thank Cathy Koehler, Paul Aigner, Scott Moore, Dr.

462

Guangchao Li, Doug Turner, Dr. Juan Lezama Pacheco, Dr. Adam Jew, Anita Falen and

463

numerous students from Stanford University for field and laboratory assistance. Use of

464

the Stanford Synchrotron Radiation Lightsource, SLAC National Accelerator Laboratory,

465

is supported by the U.S. Department of Energy, Office of Science, Office of Basic



466

Energy Sciences under Contract No. DE-AC02-76SF00515. The SSRL Structural

467

Molecular Biology Program is supported by the DOE Office of Biological and

468

Environmental Research, and by the National Institutes of Health, National Institute of

469

General Medical Sciences (including P41GM103393). The contents of this publication

470

are solely the responsibility of the authors and do not necessarily represent the official

471

views of NIGMS or NIH.

472 473 474 475 476 477 478 479 480 481 482 483 484 485 486 487 488 489 490 491 492 493 494 495 496 497 498 499 500 501 502 503 504 505

REFERENCES (1) Oze, C.; Bird, D. K.; Fendorf, S. Genesis of hexavalent chromium from natural sources in soil and groundwater. Proc. Natl. Acad. Sci. 2007, 104 (16), 6544– 6549. (2) McClain, C. N.; Maher, K. Chromium fluxes and speciation in ultramafic catchments and global rivers. Chem. Geol. 2016, 426, 135–157. (3) Occurrence estimation methodology and occurrence findings report for the sixyear review of existing national primary drinking water regulations; EPA-815R-03-006; United States Environmental Protection Agency Office of Water: Washington, DC, 2003; www.epa.gov/sites/production/files/201412/documents/815r03006.pdf. (4) Chromium-6 in U.S. tap water; Environmental Working Group: Washington, DC, 2010; www.ewg.org/chromium6-in-tap-water. (5) Beaumont, J. J.; Sedman, R. M.; Reynolds, S. D.; Sherman, C. D.; Li, L.-H.; Howd, R. A.; Sandy, M. S.; Zeise, L.; Alexeeff, G. V. Cancer mortality in a Chinese population exposed to hexavalent chromium in drinking water. Epidemiology 2008, 19 (1), 12–23. (6) Toxicology and carcinogenesis studies of sodium dichromate dehydrate in F344/N rats and B6C3F1 mice; NTP TR 546; National Toxicology Program: Research Triangle Park, NC, 2008; ntp.niehs.nih.gov/ntp/htdocs/lt_rpts/tr546.pdf. (7) Nriagu, J. O.; Nieboer, E. Chromium in the natural and human environments; Wiley-Interscience: New York, 1988. (8) Oze, C.; Fendorf, S.; Bird, D. K.; Coleman, R. G. Chromium geochemistry of serpentine soils. Int. Geol. Rev. 2004, 46 (2), 97–126. (9) Kierczak, J.; Neel, C.; Bril, H.; Puziewicz, J. Effect of mineralogy and pedoclimatic variations on Ni and Cr distribution in serpentine soils under temperate climate. Geoderma 2007, 142 (1-2), 165–177. (10) Lee, B. D.; Graham, R. C.; Laurent, T. E.; Amrhein, C. Pedogenesis in a wetland meadow and surrounding serpentinitic landslide terrain, northern California, USA. Geoderma 2004, 118 (3), 303–320. (11) Caillaud, J.; Proust, D.; Philippe, S.; Fontaine, C.; Fialin, M. Trace metals distribution from a serpentinite weathering at the scales of the weathering profile 24 ACS Paragon Plus Environment

Page 24 of 29

Page 25 of 29

506 507 508 509 510 511 512 513 514 515 516 517 518 519 520 521 522 523 524 525 526 527 528 529 530 531 532 533 534 535 536 537 538 539 540 541 542 543 544 545 546 547 548 549 550 551


(12) (13)

(14) (15)

(16)

(17) (18) (19)

(20) (21)

(22)

(23)

(24)

(25) (26)

(27)

and its related weathering microsystems and clay minerals. Geoderma 2009, 149 (3-4), 199–208. Garnier, J.; Quantin, C.; Guimarães, E.; Becquer, T. Can chromite weathering be a source of Cr in soils? Mineral. Mag. 2008, 72 (1), 49–53. Morrison, J. M.; Goldhaber, M. B.; Mills, C. T.; Breit, G. N.; Hooper, R. L.; Holloway, J. M.; Diehl, S. F.; Ranville, J. F. Weathering and transport of chromium and nickel from serpentinite in the Coast Range ophiolite to the Sacramento Valley, California, USA. Appl. Geochem. 2015, 61, 72–86. Bartlett, R. Behavior of chromium in soils. III. Oxidation. J. Environ. Qual. 1979, 8 (1), 31–34. Garnier, J.; Quantin, C.; Guimarães, E. M.; Vantelon, D.; Montargès-Pelletier, E.; Becquer, T. Cr(VI) genesis and dynamics in Ferralsols developed from ultramafic rocks: The case of Niquelândia, Brazil. Geoderma 2013, 193-194, 256–264. Fandeur, D.; Juillot, F.; Morin, G.; Olivi, L.; Cognigni, A.; Webb, S. M.; Ambrosi, J.-P.; Fritsch, E.; Guyot, F.; Brown, G. E., Jr. 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. Env. Sci. Technol. 2009, 43 (19), 7384–7390. Becquer, T. Chromium availability in ultramafic soils from New Caledonia. Sci. Total Environ. 2003, 301 (1-3), 251–261. Cooper, G. Oxidation and toxicity of chromium in ultramafic soils in Zimbabwe. Appl. Geochem. 2002, 17 (8), 981–986. Garnier, J.; Quantin, C.; Martins, E. S.; Becquer, T. Solid speciation and availability of chromium in ultramafic soils from Niquelândia, Brazil. J. Geochem. Explor. 2006, 88 (1-3), 206–209. Murray, K. J.; Tebo, B. M. Cr(III) Is Indirectly Oxidized by the Mn(II)-oxidizing bacterium Bacillus sp. strain SG-1. Env. Sci. Technol. 2007, 41 (2), 528–533. Landrot, G.; Ginder-Vogel, M.; Sparks, D. L. Kinetics of chromium (III) oxidation by manganese (IV) oxides using quick scanning X-ray absorption fine structure spectroscopy (Q-XAFS). Env. Sci. Technol. 2010, 44 (1), 143–149. Rajapaksha, A. U.; Vithanage, M.; Ok, Y. S.; Oze, C. Cr(VI) Formation related to Cr(III)-muscovite and birnessite interactions in ultramafic environments. Env. Sci. Technol. 2013, 47 (17), 9722–9729. Namgung, S.; Kwon, M. J.; Qafoku, N. P.; Lee, G. Cr(OH)3(s) Oxidation induced by surface catalyzed Mn(II) oxidation. Env. Sci. Technol. 2014, 48 (18), 10760–10768. Weaver, R. M.; Hochella, M. F. J. The reactivity of seven Mn-oxides with Cr3+aq: A comparative analysis of a complex, environmentally important redox reaction. Am. Mineral. 2003, 88 (11-12), 2016. Nico, P. S.; Zasoski, R. J. Importance of Mn(III) Availability on the rate of Cr(III) oxidation on δ-MnO2. Env. Sci. Technol. 2000, 34 (16), 3363–3367. Tang, Y.; Webb, S. M.; Estes, E. R.; Hansel, C. M. Chromium(III) oxidation by biogenic manganese oxides with varying structural ripening. Env. Sci. : Processes Impacts 2014, 16, 2127–2136. Eary, L. E.; Rai, D. Kinetics of chromium(III) oxidation to chromium(VI) by 25 ACS Paragon Plus Environment


552 553 554 555 556 557 558 559 560 561 562 563 564 565 566 567 568 569 570 571 572 573 574 575 576 577 578 579 580 581 582 583 584 585 586 587 588 589 590 591 592 593 594 595 596 597

(28) (29)

(30)

(31) (32)

(33)

(34)

(35)

(36)

(37) (38)

(39)

(40)

(41)

(42)

reaction with manganese dioxide. Env. Sci. Technol. 1987, 21 (12), 1187–1193. Oze, C.; Sleep, N. H.; Coleman, R. G.; Fendorf, S. Anoxic oxidation of chromium. Geology 2016, G37844.1. Rock, M. L.; James, B. R.; Helz, G. R. Hydrogen peroxide effects on chromium oxidation state and solubility in four diverse, chromium-enriched soils. Env. Sci. Technol. 2001, 35 (20), 4054–4059. Fendorf, S.; Wielinga, B. W.; Hansel, C. M. Chromium transformations in natural environments: The role of biological and abiological processes in chromium(VI) reduction. Int. Geol. Rev. 2000, 42 (8), 691–701. Fendorf, S. E. Surface reactions of chromium in soils and waters. Geoderma 1995, 67 (1-2), 55–71. Gasser, U. G.; Juchler, S. J.; Sticher, H. Chemistry and speciation of soil water from serpentinitic soils: Importance of colloids in the transport of Cr, Fe, Mg, and Ni. Soil Sci. 1994, 158 (5), 314. Gough, L.; Meadows, G.; Jackson, L.; Dudka, S. Biogeochemistry of a highly serpentinized, chromite-rich ultramafic area, Tehama County, California; United States Geological Survey Bulletin 1901: Denver, CO; 1989; pubs.usgs.gov/bul/1901/report.pdf. Anderson, A. J.; Meyer, D. R.; Mayer, F. K. Heavy metal toxicities: levels of nickel, cobalt and chromium in the soil and plants associated with visual symptoms and variation in growth of an oat crop. Aust. J. Agric. Res. 1973, 24 (4), 557–571. Becquer, T.; Quantin, C.; Boudot, J. P. Toxic levels of metals in Ferralsols under natural vegetation and crops in New Caledonia. Eur. J. Soil Sci. 2010, 61 (6), 994–1004. Novák, M.; Chrastný, V.; Čadkova, E.; Farkaš, J.; Bullen, T. D.; Tylcer, J.; Szurmanova, Z.; Cron, M.; Prechova, E.; Curik, J.; et al. Common occurrence of a positive δ53Cr shift in central European waters contaminated by geogenic/industrial chromium relative to source values. Env. Sci. Technol. 2014, 48 (11), 6089–6096. Kruckeberg, A. R. California serpentines: flora, vegetation, geology, soils, and management problems; University of California Press, Ltd.: Berkeley, CA, 1984. Izbicki, J. A.; Wright, M. T.; Seymour, W. A.; McCleskey, R. B.; Fram, M. S.; Belitz, K.; Esser, B. K. Cr(VI) occurrence and geochemistry in water from public-supply wells in California. Appl. Geochem. 2015, 63, 203-217. Manning, A. H.; Mills, C. T.; Morrison, J. M.; Ball, L. B. Insights into controls on hexavalent chromium in groundwater provided by environmental tracers, Sacramento Valley, California, USA. Appl. Geochem. 2015, 62, 186–199. Gonzalez, A. R.; Ndung'u, K.; Flegal, A. Natural occurrence of hexavalent chromium in the Aromas Red Sands aquifer, California. Env. Sci. Technol. 2005, 39 (15), 5505–5511. Abu-Saba, K. E.; Flegal, A. R. Temporally variable freshwater sources of dissolved chromium to the San Francisco Bay Estuary. Env. Sci. Technol. 1997, 31 (12), 3455–3460. Morrison, J. M.; Goldhaber, M. B.; Lee, L.; Holloway, J. A. M.; Wanty, R. B.; Wolf, R. E.; Ranville, J. F. A regional-scale study of chromium and nickel in 26 ACS Paragon Plus Environment

Page 26 of 29

Page 27 of 29

598 599 600 601 602 603 604 605 606 607 608 609 610 611 612 613 614 615 616 617 618 619 620 621 622 623 624 625 626 627 628 629 630 631 632 633 634 635 636 637 638 639 640 641 642 643


(43)

(44)

(45)

(46) (47)

(48) (49) (50)

(51)

(52)

(53)

(54)

(55) (56)

(57) (58)

soils of northern California, USA. Appl. Geochem. 2009, 24 (8), 1500–1511. Belitz, K.; Fram, M. S.; Johnson, T. D. Metrics for assessing the quality of groundwater used for public supply, CA, USA: Equivalent-population and area. Env. Sci. Technol. 2015, 49 (14) 8330-8338. SoilWeb; United States Department of Agriculture Natural Resources Conservation Service SSURGO Database; casoilresource.lawr.ucdavis.edu/gmap/. Hopson, C. A.; Mattinson, J. M.; Pessagno, E. A.; Luyendyk, B. P. California Coast Range ophiolite: Composite middle and late Jurassic oceanic lithosphere. In Ophiolites, Arcs, and Batholiths: A Tribute to Cliff Hopson; Wright, J. E., Shervais, J., Eds.; The Geological Society of America: Boulder, CO; 2008. Knoxville Creek California; www.wrcc.dri.edu/cgi-bin/rawMAIN.pl?caCKNO. Blake, G. R.; Hartge, K. H. Bulk Density. In Methods of Soil Analysis Part 1 Physical and Mineralogical Methods; Klute, A., Ed.; Soil Science Society of America: Madison, WI, 1986. Guelph Permeameter Kit; www.soilmoisture.com/GUELPH-PERMEAMETERKIT/. Loppert, R. L.; Inskeep, W. P. Iron. In Methods of Soil Analysis; Bigham, J. M., Ed.; Soil Science Society of America: Madison, WI, 1996; pp 639–664. Arar, E. J.; Pfaff, J. D.; Martin, T. D. Determination of dissolved hexavalent chromium in drinking water, groundwater and industrial wastewater effluents by ion chromatography; Method 218.6, Revision 3.3.; U.S. Environmental Protection Agency: Cincinnati, OH, 1994; https://www.epa.gov/sites/production/files/2015-08/documents/method_2186_rev_3-3_1994.pdf Krumbein, W. E.; Altmann, H. J. A new method for the detection and enumeration of manganese oxidizing and reducing microorganisms. Helgol. Wiss. Meeresunters. 1973, 25, 347–356. Webb, S. The MicroAnalysis Toolkit: X-ray fluorescence image processing software. AIP Conference Proceedings. 2011, 1365 (1), 196-199; www.samsxrays.com/#!smak/n44a5. Ravel, B.; Newville, M. ATHENA, ARTEMIS, HEPHAESTUS: Data analysis for X-ray absorption spectroscopy using IFEFFIT. J. Synchrotron Radiat. 2005, 12, 537–541. Stonestrom, D. A.; White, A. F.; C Akstin, K. Determining rates of chemical weathering in soils—solute transport versus profile evolution. J. Hydrol. 1998, 209 (1-4), 331–345. Van Genuchten, M. T. A closed-form equation for predicting the hydraulic conductivity of unsaturated soils. Soil Sci. Soc. Am. J. 1980, 44 (5), 892–898. Oze, C. Chromium geochemistry in serpentinized ultramafic rocks and serpentine soils from the Franciscan complex of California. Am, J. Sci. 2004, 304 (1), 67–101. Gasser, U. G..; Juchler, S.; Sticher, H. The fate of chromium and nickel in subalpine soils derived from serpentinite. Can. J. Soil Sci. 1995, 75 (2), 187-195. Gasser, U. G.; Dahlgren, R. A. Solid-Phase Speciation and Surface Association of Metals in Serpentinitic Soils. Soil Sci. 1994, 158 (6), 409-420. 27 ACS Paragon Plus Environment


644 645 646 647 648 649 650 651 652 653 654 655 656 657 658 659 660 661 662 663 664 665 666 667 668 669 670 671 672 673 674 675 676 677 678 679 680 681 682 683 684 685 686 687 688 689

(59)

(60)

(61)

(62)

(63)

(64)

(65) (66)

(67)

(68)

(69)

(70) (71)

(72)

(73)

Traore, D.; Beauvais, A.; Chabaux, F.; Peiffert, C.; Parisot, J. C.; Ambrosi, J. P.; Colin, F. Chemical and physical transfers in an ultramafic rock weathering profile: Part 1. Supergene dissolution of Pt-bearing chromite. Am. Mineral. 2008, 93 (1), 22–30. Tebo, B. M.; Bargar, J. R.; Clement, B. G.; Dick, G. J.; Murray, K. J.; Parker, D.; Verity, R.; Webb, S. M. Biogenic manganese oxides: Properties and mechanisms of formation. Annu. Rev. Earth Planet. Sci. 2004, 32 (1), 287–328. Bargar, J. R.; Tebo, B. M.; Bergmann, U.; Webb, S. M.; Glatzel, P.; Chiu, V. Q.; Villalobos, M. Biotic and abiotic products of Mn(II) oxidation by spores of the marine Bacillus sp.strain SG-1. American Mineralogist 2005, 90 (1), 143–154. Duckworth, O. W.; Akafia, M. M.; Andrews, M. Y.; Bargar, J. R. Siderophorepromoted dissolution of chromium from hydroxide minerals. Env. Sci. : Processes Impacts 2014, 16 (6), 1348–1359. Rai, D.; Moore, D. A.; Hess, N. J.; Rosso, K. M.; Rao, L.; Heald, S. M. Chromium(III) hydroxide solubility in the aqueous K+-H+-OH−-CO2-HCO3− CO32−-H2O system: A thermodynamic model. J. Solution Chem. 2007, 36 (10), 1261–1285. Gabriel, C.; Raptopoulou, C. P.; Terzis, A.; Tangoulis, V.; Mateescu, C.; Salifoglou, A. pH-Specific synthesis and spectroscopic, structural, and magnetic studies of a Chromium(III)−citrate species. Aqueous solution speciation of the binary Chromium(III)−citrate system. Inorg. Chem. 2007, 46 (8), 2998–3009. Hamada, Y. Z.; Holyfield, H.; Rosli, K.; Burkey, T. Equilibrium models of Cr3+ and Cu2+ with glutamate. J Coord. Chem. 2009, 62 (5), 721–733. Zachara, J. M.; Girvin, D. C.; Schmidt, R. L.; Resch, C. T. Chromate adsorption on amorphous iron oxyhydroxide in the presence of major groundwater ions. Env. Sci. Technol. 1987, 21 (6), 589–594. Villalobos, M.; Trotz, M. A.; Leckie, J. O. Surface complexation modeling of carbonate effects on the adsorption of Cr(VI), Pb(II), and U(VI) on goethite. Env. Sci. Technol. 2001, 35 (19), 3849–3856. Mills, C.; Morrison, J. M.; Goldhaber, M. B.; Ellefsen, K. J. 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. Appl. Geochem. 2011, 26 (8), 1488–1501. Sherlock, M. D.; Chappell, N. A.; McDonnell, J. J. Effects of experimental uncertainty on the calculation of hillslope flow paths. Hydrol. Processes 2000, 14, 2457–2471. Chappell, N.; Ternan, L. Flow path dimensionality and hydrological modelling. Hydrol. Processes 1992, 6, 327–345. Kieber, R. J.; Willey, J. D.; Zvalaren, S. D. Chromium speciation in rainwater: Temporal variability and atmospheric deposition. Env. Sci. Technol. 2002, 36 (24), 5321–5327. Malmström, M. E.; Destouni, G.; Banwart, S. A.; Strömberg, B. H. E. Resolving the Scale-Dependence of Mineral Weathering Rates. Env. Sci. Technol. 2000, 34 (7), 1375–1378. White, A. F.; Brantley, S. L. The effect of time on the weathering of silicate minerals: why do weathering rates differ in the laboratory and field? Chem. Geol. 28 ACS Paragon Plus Environment

Page 28 of 29

Page 29 of 29

690 691 692 693 694 695 696 697 698 699 700 701


(74)

(75) (76) (77)

2003, 202 (3-4), 479–506. Madison, A. S.; Tebo, B. M.; Mucci, A.; Sundby, B.; Luther, G. W. Abundant porewater Mn(III) is a major component of the sedimentary redox system. Science 2013, 341 (6148), 875–878. Di Baldassare, G.; Montanari, A. Uncertainty in river discharge observations: a quantitative analysis. Hydrol. Earth Syst. Sci. 2009, 13, 913–921. Maher, K.; Chamberlain, C. P. Hydrologic Regulation of Chemical Weathering and the Geologic Carbon Cycle. Science 2014, 343 (6178), 1502–1504. Unruh, J. R.; Moores, E. M. Quaternary blind thrusting in the southwestern Sacramento Valley, California. Tectonics 1992, 11 (2), 192–203.


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