The Influence of a Fluctuating Water Table on Arsenic Mobility in a Western U.S. Aquifer

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1 Utah State University All Graduate Theses and Dissertations Graduate Studies The Influence of a Fluctuating Water Table on Arsenic Mobility in a Western U.S. Aquifer Allia Maher Abu-Ramaileh Utah State University Follow this and additional works at: Part of the Civil and Environmental Engineering Commons Recommended Citation Abu-Ramaileh, Allia Maher, "The Influence of a Fluctuating Water Table on Arsenic Mobility in a Western U.S. Aquifer" (2015). All Graduate Theses and Dissertations This Thesis is brought to you for free and open access by the Graduate Studies at DigitalCommons@USU. It has been accepted for inclusion in All Graduate Theses and Dissertations by an authorized administrator of DigitalCommons@USU. For more information, please contact dylan.burns@usu.edu.

2 THE INFLUENCE OF A FLUCTUATING WATER TABLE ON ARSENIC MOBILITY IN A WESTERN U.S. AQUIFER by Allia Maher Abu-Ramaileh A thesis submitted in partial fulfillment of the requirements for the degree of MASTER OF SCIENCE in Civil and Environmental Engineering Approved: Professor Joan E. McLean Major Professor Professor R. Ryan Dupont Committee Member Professor Paul Grossl Committee Member Dr. Mark McLellan Vice President for Research and Dean of the School of Graduate Studies UTAH STATE UNIVERSITY Logan, Utah 2015

3 ii Copyright Allia Maher Abu-Ramaileh 2015 All Rights Reserved

4 iii ABSTRACT INFLUENCE OF A FLUCTUATING WATER TABLE ON ARSENIC MOBILITY IN A WESTERN U.S. AQUIFER by Allia Maher Abu-Ramaileh, Master of Science Utah State University, 2015 Major Professor: Joan E. McLean Department: Civil and Environmental Engineering Arsenic (As) concentrations in groundwater that exceed the Maximum Contaminant Level (MCL) (10 µg/l) for drinking water have been reported throughout the United States, with higher occurrences in the Southwestern basin-fill aquifers. Levels of As above the MCL were measured in wells throughout the Cache Valley Basin, Utah. The As is naturally occurring in geologic material from the soil surface to depths of groundwater. This study reports on the mechanisms of retention and solubilization of As through these zones using geochemical modeling and microcosm studies. Two cores (NP 9 and NP 13) were collected from the soil surface to the depth of groundwater and sectioned based on observed redoximorphic features. Pore water was analyzed for As and iron( Fe) redox species, general water quality parameters and solid phase As, Fe and Mn using sequential extractions. These data were used in PHREEQC and MINTEQ geochemical models to predict mechanisms of As retention. Microcosm studies were performed using sediments from the water table zone. The sediments were

5 iv exposed to oxidized, reduced, and poisoned conditions over time to evaluate the effect of the seasonal fluctuating water table on As release. Modeling results indicated As(V) was dominantly sorbed to hydrous ferric oxides (HFO) throughout both profiles. Although much less As(V) was sorbed to CaCO 3, the percentage associated with calcite was 1.7 to 3.3% and 6 to 59% in the surface and water table zones for NP 9 and NP 13, respectively. As(III) solubility was controlled by the formation of an As-S mineral, orpiment. Microcosm findings, over 113-day incubation, concluded that regardless of treatment condition, As is released. For reduced samples As in solution was primarily As(III), while oxidized and poisoned samples only released As(V). The release of As under every condition, and the lack of reduced As and Fe in the poisoned samples, indicates that As release is abiotically controlled, while reduction is microbially driven. Carbonate minerals were the source of As(V) under treatment conditions as determined using an acetate extraction. Desorption of As(V) from carbonate minerals and the reduction of As(V) to As(III) played a significant role in explaining solution phase As(III) concentrations. (171 pages)

6 v PUBLIC ABSTRACT THE INFLUENCE OF A FLUCTUATING WATER TABLE ON ARSENIC MOBILITY IN A WESTERN U.S. AQUIFER Allia Maher Abu-Ramaileh Significant amounts of Arsenic (As) are seen in groundwater used for drinking water in the United States, especially in the Southwest (SW). Long-term exposure to As concentrations over the drinking water limit of 10 µg/l can cause skin and lung cancer, along with several other adverse health effects. Previous studies have evaluated the hydro-biogeochemical processes leading to the exposure of millions of people in West Bengal, Bangladesh, and Southeast Asia to unsafe levels of As in drinking water. There are some similarities between Southeast Asia and the SW U.S., but only a few studies analyzing As have been performed in the SW U.S. Levels of As above the drinking water limit were measured in groundwater wells throughout Cache Valley, UT. Field research performed by our research team found that As is naturally occurring. The processes controlling As release from the soil to groundwater in this area have not been extensively studied. Modeling was used to determine which As mineral forms exist in the soil and lab experiments were conducted to see if changing groundwater levels affect As release. Solids prediction modeling results found that As(III), the most toxic form of As, forms an As-S mineral, orpiment. However, when considering sorption mechanisms as well,

7 vi modeling results indicated most of the As(V) sorbs onto hydrous ferric oxides (HFO). Although much less As(V) was sorbed to CaCO 3, the percentage associated with calcite was 1.7 to 3.3% and 6 to 59% in the surface and water table zones for NP 9 and NP 13, respectively. The lab experiments took soil from groundwater depth and looked at As release under different conditions; 1) exposed to air, 2) oxygen free and exposed to nitrogen gas, and 3) poisoned to eliminate microbial activity. Despite the condition, large amounts of As were released into the water from the soil. The samples not exposed to air released both As(III) and As(V), but the samples exposed to air and the poisoned samples only released As(V). Since As was released under every condition, but only As(III) was seen in the samples not exposed to air, As release was deduced to be chemically controlled. However, in order to release As(III), microbes were needed. Chemical extractions were conducted to determine the amount of As and iron (Fe) associated with carbonate minerals. Desorption of As(V) from carbonate minerals and the reduction of As(V) to As(III) played a significant role in explaining As(III) in solution.

8 vii ACKNOWLEDGMENTS This thesis would not have been possible without the guidance and the help of several individuals who in one way or another contributed and extended their valuable assistance in the preparation and completion of my Master s Degree. First and foremost, I would like to acknowledge Joan McLean, my advisor, professor, and friend. Joan provided excellent guidance and direction throughout my research, always encouraging me to consider the data in different light and consider alternate mechanisms. I would also like to acknowledge my other committee members, Dr. Ryan Dupont and Dr. Paul Grossl, for their help, guidance, and expertise throughout my research. I would like to extend my gratitude to the Utah Water Research Laboratory for funding my degree and providing all the services that made a difference in advancing my thesis. My research would not have been possible without the Utah Water Research Laboratory and the support of its excellent staff. My gratitude is also extended to scholarships I earned throughout my degree for their substantial financial support and generosity. I am deeply honored by their recognition. The scholarships had a great impact on my research and personal life. Last but not the least, I would like to thank my family and friends for their support throughout my degree. I am incredibly grateful to my husband, Adel, for his patience and encouragement throughout my degree, especially during the tough and challenging times. My dad, sisters, and brother were all incredibly understanding

9 throughout the whole process as well. Without the support of those mentioned above, my thesis would not have been possible. "ق ل إ ن ص ال ي ت و ن س ي ك و م ح ي اي و م م ا ي ت ي ل ل ي ر يب إل ع ال يمني " Say, Indeed, my prayer, my rites, my living and my dying are for Allah, Lord of the worlds (Quran, Surat al-an'am: 162) viii

10 ix CONTENTS Page ABSTRACT... iii PUBLIC ABSTRACT...v ACKNOWLEDGMENTS... vii LIST OF TABLES... xii LIST OF FIGURES... xiv INTRODUCTION...1 LITERATURE REVIEW Arsenic Sources Arsenic Toxicity Arsenic Mineralogy Iron Mineralogy Arsenic Sorption Arsenic Sorption to Iron and Manganese Oxides Arsenic Sorption to Carbonate Minerals Desorption of Arsenic from Mineral Surfaces: Bicarbonate Natural Organic Matter (NOM) Arsenic Reductive Microbial Processes Reductive Dissolution of Iron Oxides Direct Arsenic Reduction Abiotic Redox Processes Alternating Redox Conditions Geochemical Equilibrium Modeling Modeling Inorganic Species Precipitation and Dissolution Reactions Double Layer FeOH Model Uncertainty in Equilibrium Models Successful Geochemical Modeling Applications in Arsenic Research STUDY BACKGROUND...21

11 3.1 Hypotheses and Objectives MATERIALS AND METHODS Objective 1- Geochemical Modeling Characterization of Collected Soil Model Selection Objective 1 Data Analysis PHREEQC Modeling MINTEQ Modeling Model Alterations Objective 2 Redox Experiment Poisoned Controls Quality Control and Statistics Objective 2 Data Analysis RESULTS NP 9 and NP 13 Pore Water Data NP 9 and NP 13 Solid Phase Extractions NP 9 and NP 13 Geochemical Modeling Results - Solid Phases Controlling Arsenic Solution Chemistry Sorption onto FeOH and CaCO NP 9 and NP 13 Modeled Eh Redox Experiment Results -As Solubilization and Fe Reduction Arsenic Associated with the Solid Phase Additional Analyzed Parameters Control Experiment Results DISCUSSION Summary and Conclusions ENGINEERING SIGNIFICANCE...94 REFERENCES...96 PORE WATER AND SEQUENTIAL EXTRACTION PROCEDURE PHREEQC AND MINTEQ MODEL INPUTS SHORT-TERM PRELIMINARY EH STUDY SOLIDS PREDICTED IN PHREEQC REDOX RESULTS: EH, EC, AND PH x

12 xi REDOX RESULTS REDOX EXPERIMENT ACETATE EXTRACTABLE CA, MN, AND MG CONTROL RESULTS CONTROL ATP RESULTS...156

13 xii LIST OF TABLES Table Page 1 Several Iron(III) Oxy-Hydroxides (Cornell and Schwertmann 2003) Summary of Arsenic Sorption to Iron and Manganese Oxides Studies Summary of NOM-As Related Research (Adapted from Wang and Mulligan (2006); includes Cornu et al. (1999), Grafe et al. (2001), Redman et al. (2002), and Thanabalasingam and Pickering (1986) Summary of NP 9-5 and NP 13-3 Parameters NP 9 and NP 13 Groundwater Properties Sequential Extraction Results, As associated minerals Sequential Extraction Results, Fe associated minerals NP 9 As(III) Distribution and Sorption onto HFO as Predicted by MINTEQ (small h denotes high affinity sites on HFO) NP 9 As(V) Distribution and Sorption onto HFO and CaCO3 as Predicted by MINTEQ (small h denotes high affinity sites on HFO) NP 13 As(III) Distribution and Sorption onto HFO as Predicted by MINTEQ (small h denotes high affinity sites on HFO) NP 13 As(V) Distribution and Sorption onto HFO and CaCO 3 as Predicted by MINTEQ (small h denotes high affinity sites on HFO) A.1 Aqueous Analytes and Analytical Methods A.2 Sequential Extraction Procedure B.1 NP 9 Inputs for PHREEQC and MINTEQ Models B.2 NP 13 Inputs for PHREEQC and MINTEQ Models B.3 NP 9 Ca and solids concentrations for CaCO3 entered into MINTEQ B.4 NP 13 Ca and solids concentrations for CaCO 3 entered into MINTEQ C.1 Pilot Study Aqueous As, µg/l, in NP 9-5 and NP C.2 Pilot Study Average Aqueous As, µg/l, in NP 9-5 and NP C.3 Pilot Study Aqueous Fe, µg/l, in NP 9-5 and NP C.4 Pilot Study Average Aqueous Fe, µg/l, in NP 9-5 and NP C.5 Pilot Study Aqueous Fe(II), mg/l, in NP 9-5 and NP C.6 Pilot Study Average Aqueous Fe(II), mg/l, in NP 9-5 and NP C.7 Pilot Study ph in NP 9-5 and NP C.8 Pilot Study Average ph in NP 9-5 and NP C.9 Pilot Study Eh, mv in NP 9-5 and NP C.10 Pilot Study Average Eh, mv, in NP 9-5 and NP C.11 Pilot Study HCl As, µg/kg, in NP 9-5 and NP C.12 Pilot Study Average HCl As, µg/kg, in NP 9-5 and NP

14 C.13 Pilot Study HCl Fe, mg/kg, in NP 9-5 and NP C.14 Pilot Study Average HCl Fe, mg/kg, in NP 9-5 and NP C.15 Pilot Study HCl Fe(II), mg/kg, in NP 9-5 and NP C.16 Pilot Study Average HCl Fe(II), mg/kg, in NP 9-5 and NP C.17 Pilot Study HCl Fe(III), mg/kg, in NP 9-5 and NP C.18 Pilot Study Average HCl Fe(III), mg/kg, in NP 9-5 and NP D.1 Solids Predicted in PHREEQC for NP NP D.2 Solids Predicted in PHREEQC for NP 9-4 and NP D.3 Solids Predicted in PHREEQC for NP NP D.4 Solids Predicted in PHREEQC for NP NP D.5 Solids Predicted in PHREEQC for NP 13-1 and NP D.6 Solids Predicted in PHREEQC for NP 13-3 and NP D.7 Solids Predicted in PHREEQC for NP NP D.8 Solids Predicted in PHREEQC for NP NP F.1 Summary of 9-5 Oxidized HCl and Acetate Extractable Fe and As F.2 Summary of 9-5 Reduced HCl and Acetate Extractable Fe and As F.3 Summary of 13-3 Oxidized HCl and Acetate Extractable Fe and As F.4 Summary of 13-3 Reduced HCl and Acetate Extractable Fe and As I.1 Control Experiment ATP Results I.2 Control Experiment Average ATP Results xiii

15 xiv LIST OF FIGURES Figure Page 1 Arsenic (μg/l) Prediction at 200 Foot Aquifer Penetration Depth Southwestern Basin (Anning et al. 2012) Southwestern U.S. Principal Aquifers and Basin Locations (Anning et al. 2012) Sampling Sites, NP 9 and NP 13, Near the Logan Landfill Schematic of NP 9 (A) NP 13 (B) Soil Profile Depths and Zones Soil Profile Sketch. Larger Ellipses Indicate Higher Pore Water and Solid Phase Concentrations Sampling and Bubbling Dates. Red circles indicate sampling and yellow stars represent bubbling Pore Water NP 9As, μg/l (A), NP 9 Fe, mg/l (B) NP 13 As, µg/l (C) and NP 13 Fe mg/l (D). The shaded region indicates the transition zone Percent of As extracted at each extraction step down the NP 9(A) and NP 13 (B) Profiles The Concentration of As(III) (A) and As(V) (B) Extracted in NP 9, µg/kg and the Concentration As(III) (C) and As(V) (D) Extracted in NP 13, µg/kg Percent of Fe extracted at each extraction step down the NP 9 (A) and NP 13 (B) Profiles Predicted Solids in NP 9-4 and NP 9-5 (A) and Predicted Solids in NP 13-3 and NP 13-4 (B) Modeled Eh Calculated Using As, N, Fe, and S Redox Pairs Down the NP 9 (A) and NP 13 (B) Profiles Measured Redox down the NP 9 Profile Measured Redox down the NP 13 Profile HCl Extractable Fe(II) and Aqueous As(III) and As(V) in NP 9-5 (A) and NP 13-3 (B) Reduced Samples. Red circles indicate point at which reduction occurred. Error bars represent ± Tukey's HSD HCl Extractable Fe(II) and Aqueous As(III) and As(V) in NP 9-5 (A) and NP 13-3 (B) Oxidized Samples. Error bars represent ± Tukey's HSD Acetate Extractable Fe(II) (mg/kg) (A) and As (III) and As(V) (µg/kg) (B) in NP 9-5 Under Oxidized (solid line) and Reduced (dashed line) Conditions. Error bars represent ± Tukey's HSD

16 18 Acetate Extractable Fe(II) (mg/kg) (A) and As (III) and As(V) (µg/kg) (B) in NP 13-3 Under Oxidized (solid line) and Reduced (dashed line) Conditions. Error bars represent ± Tukey's HSD NP 9-5 Distribution of As in the Carbonate Mineral and Aqueous Phases Over 113 Days Under Reduced Conditions NP 13-3 Distribution of As in the Carbonate Mineral and Aqueous Phases Over 113 Days Under Reduced Conditions NP 9-5 Distribution of As in the Carbonate Mineral and Aqueous Phases Over 113 Days Under Oxidized Conditions. (Only days with significant differences from time 0 are displayed) NP 13-3 Oxidized Distribution of As in the Carbonate Mineral and Aqueous Phases Over 113 Days Under Reduced Conditions. (Only days with significant differences from time 0 are displayed) NP 9-5 Poisoned and Active Aqueous As(V). Error bars represent ± Tukey's HSD across all treatments NP 9-5 Poisoned and Active HCl Extractable Total As over 21 days. Error bars represent ± Tukey's HSD across all treatments NP 9-5 HCl Extractable Fe(II) in Active and Poisoned Samples. Error bars represent ± Tukey's HSD across all treatments NP 13-3 Poisoned and Active Aqueous As(V). Error bars represent ± Tukey's HSD NP 13-3 Poisoned and Active HCl Extractable Total As over 21 days. Error bars represent ± Tukey's HSD across all treatments NP 13-3 HCl Extractable Fe(II) in Active and Poisoned Samples. Error bars represent ± Tukey's HSD E.1 Eh in NP 9-5 (A) and NP 13-3 (B) under Oxidized (solid line) and Reduced Conditions (dashed line). Error bars represent ± Tukey's HSD E.2 NP 9-5 EC Values over time. Solid and dashed lines indicate oxidized and reduced conditions, respectively. Error bars represent Tukey's HSD E.3 NP 13-3 EC Values over time. Solid and dashed lines indicate oxidized and reduced conditions, respectively. Error bars represent Tukey's HSD E.4 ph in NP 9-5 under Oxidized (solid line) and Reduced Conditions (dashed line). Eh, ph, As(III), and As(V) error bars represent Tukey's HSD E.5 NP 13-3 ph under Oxidized (solid line) and Reduced Conditions (dashed line). Eh, ph, As(III), and As(V) error bars represent Tukey's HSD F.1 HCl extractable total As (µg/kg) (B) in NP 9-5 under oxidized (solid line) and reduced (dashed line). Error bars represent Tukey's HSD F.2 NP 9-5 HCl extractable total As (µg/kg). Error bars represent 95% confidence interval xv

17 F.3 NP 13-3 HCl extractable total As (µg/kg). Error bars represent 95% confidence interval F.4 NP 9-5 Oxidized and Reduced HCl Extractable Fe(III). Error bars represent Tukey's HSD F.5 NP 13-3 Oxidized and Reduced HCl Extractable Fe(III). Error bars represent Tukey's HSD F.6 NP 9-5 Aqueous Fe(III) F.7 NP 13-3 Aqueous Fe(III) G.1 NP 9-5 Oxidized Acetate Extractable Calcium, mg/kg. Error bars represent Tukey's HSD G.2 NP 9-5 Reduced Acetate Extractable Calcium, mg/kg. Error bars represent Tukey's HSD G.3 NP 13-3 Oxidized Acetate Extractable Calcium, mg/kg. Error bars represent Tukey's HSD G.4 NP 13-3 Reduced Acetate Extractable Calcium, mg/kg. Error G.5 bars represent Tukey's HSD G.5 NP 9-5 Oxidized Acetate Extractable Manganese, mg/kg. Error bars represent Tukey's HSD G.6 NP 9-5 Reduced Acetate Extractable Manganese, mg/kg. Error bars represent Tukey's HSD G.7 NP 13-3 Oxidized Acetate Extractable Manganese, mg/kg. Error bars represent Tukey's HSD G.8 NP 13-3 Reduced Acetate Extractable Manganese, mg/kg. Error bars represent Tukey's HSD G.9 NP 9-5 Oxidized Acetate Extractable Magnesium, mg/kg. Error bars represent Tukey's HSD G.10 NP 9-5 Reduced Acetate Extractable Magnesium, mg/kg. Error bars represent Tukey's HSD G.11 NP 13-3 Oxidized Acetate Extractable Magnesium, mg/kg. Error bars represent Tukey's HSD G.12 NP 13-3 Reduced Acetate Extractable Magnesium, mg/kg. Error bars represent Tukey's HSD G.13 Acetate extractable As (V) (µg/kg) in NP 9-5 under oxidized (solid line) and reduced (dashed line). Error bars represent Tukey's HSD G.14 Acetate extractable As (V) (µg/kg) in NP 13-3 under oxidized (solid line) and reduced (dashed line). Error bars represent Tukey's HSD G.15 NP 13-3 Acetate Ext. As(V) Differences in Condition. Error bars represent 95% confidence interval xvi

18 xvii H Poisoned and Active sample ph. Error bars represent Tukey's HSD across all treatments H Poisoned and Active sample ph. Error bars represent Tukey's HSD H Poisoned Eh, mv, over 21 days. Error bars represent Tukey's HSD across all treatments H Poisoned and Active sample Eh, mv, over 21 days. Error bars represent Tukey's HSD

19 CHAPTER 1 INTRODUCTION Drinking water contaminated by arsenic (As), a well-known carcinogen (IARC 2004), is a worldwide crisis, affecting millions of people across the globe (Mukherjee et al. 2011; Walvekar et al. 2007). Extensive research (Acharyya et al. 1999; Anawar et al. 2003; Nickson et al. 1998; Nickson et al. 2000; Smith et al. 2000) investigated sources and mechanisms behind the calamitous As poisoning of millions of people in Bangladesh and West Bengal, India, who use groundwater as their main drinking water source. This poisoning was deemed "the largest mass poisoning of a population in history" (Smith et al. 2000). Researchers found microbial-driven reductive dissolution of iron and manganese oxides releasing geological sources of As into groundwater to be the major release mechanisms behind the tragedy (McArthur et al. 2001; Nickson et al. 2000). Naturally occurring As is also prevalent in the Southwestern (SW) United States (U.S.). Anning et al. (2012) predicted that approximately 43% of the area within basinfill aquifers in the SW U.S. equals or exceeds the Environmental Protection Agency's (EPA) Maximum Contaminant Level (MCL) of 10 µg/l. Common aspects of these basin-fill aquifers are geology associated with volcanic activity and hydrology driven by evapotranspiration. The SW U.S. includes parts of Arizona, California, Colorado, Nevada, New Mexico, and Utah (Anning et al. 2012). Welch (2000) also found that the Rocky Mountain and Interior Plains regions of the U.S. generally have higher concentrations of As in groundwater and drinking water than the Eastern U.S. Geological and geographical similarities and differences exist between Southeast Asia

20 2 and the SW U.S., but only a few studies evaluating As groundwater contamination (Busbee et al. 2009; Welch and Lico 1998) have been performed in the SW U.S. Levels of As above the EPA's MCL have been measured in monitoring wells near the Logan City Landfill, UT and throughout the Cache Valley Basin (Sanderson and Lowe 2002). Field research performed by the research team demonstrated that As is naturally occurring in geologic material from the soil surface to depths below the groundwater table. Arsenic is associated with minerals and mineral surfaces controlled by biogeochemical environments in the vadose zone, water table zone, transition zone, and depletion zone. The mechanisms of retention and solubilization of As through these soil zones have not been extensively studied. The Cache Valley Basin lies on the eastern edge of the Southwestern Basin and Range province. The center of the valley is filled with lacustrine and alluvial sediments associated with lake cycles of Lake Bonneville, underlain by tertiary sedimentary and volcanic rocks. These tertiary sediments are exposed at high elevations in the surrounding mountains (Evans and Oaks 1996) and may be the source of As observed in the valley groundwater. From research by Meng (2015), the near surface soil (0-1 m depth) at the study site in the Cache Valley Basin contains primary and secondary As minerals. Primary As minerals, presumably of volcanic origin that are exposed to atmospheric oxygen, dissolve with oxidation to As(V), and are transported into a transition zone (1-4 m depth). The transition zone contains secondary As(V) minerals, iron oxides, and carbonate minerals. With these minerals present, the As is retained through sorption and precipitation reactions. In deeper (4-6 m depth), strongly reduced sediments, sulfate is

21 3 reduced to sulfide, forming As-Fe-S minerals. Meng (2015) also identified a zone between the vadose and transition zone where As(III) and As(V) co-occur. This zone was located within the seasonally oscillating groundwater wetting front. Understanding groundwater chemistry, mineralogy, biology, and As speciation is necessary to identify As solubility and to predict if As will be released to groundwater (Hussam et al. 2003). This research identified the different As minerals and Asassociated minerals formed under reduced, oxidized, and redox transition conditions down a soil profile, from the soil surface to below the water table, using geochemical modeling. Experimental results were also collected, analyzing the release and retention of As under different redox conditions using microcosm studies. These results help to better understand and predict As retention and solubilization in semi-arid, basin-fill aquifers. Results may also be used by water managers to identify actions that help avoid As contaminated drinking water supplies.

22 4 CHAPTER 2 LITERATURE REVIEW 2.1 Arsenic Sources The twentieth most abundant element in the Earth's crust, As, can be found throughout the environment, including groundwater and aquifer sediments. Arsenic possesses both metallic and non-metallic properties and is sensitive to mobilization under oxidizing and reducing conditions. Arsenic is often found naturally in organic-rich shale, geothermal environments, and volcanogenic sources (Nordstrom 2002). It can occur as arsenic hydrides and trivalent or pentavalent organic and inorganic compounds. Though often naturally occurring, As can also be introduced to a system by anthropogenic activities (Mandal and Suzuki 2002) including mining, and through the use of pesticides, wood preservatives, and alloying agents. 2.2 Arsenic Toxicity Several adverse effects are caused by long-term ingestion of As, including skin, lung, and bladder cancers, neurological effects, and reproductive toxicity (Schuhmacher- Wolz et al. 2009). Currently, there is no medical treatment for chronic ingestion of arsenic, called arsenicosis, and the only way to prevent it is to avoid ingesting arsenic contaminated water (Petrusevski et al. 2007). To help prevent arsenicosis and other adverse health effects, the EPA set a drinking water standard (MCL) of 10 µg/l (US- EPA 2001).

23 5 2.3 Arsenic Mineralogy There are more than 300 naturally occurring primary and secondary As minerals (Drahota and Filippi 2009). Understanding As mineralogy is important in helping to predict its stability, mobility, and toxicology. The most common primary arsenic bearing minerals include arsenopyrite (FeAsS), cobaltite (CoAsS), enargite (Cu 3 AsS 4 ), gersdorffite (NiAsS), loellingite (FeAs 2 ), orpiment (As 2 S 3 ), arsenopyrite (FeAsS), realgar (AsS), and tennantite ((Cu, Ag, Fe, Zn) 12 As 4 S 13 ) (Drahota and Filippi 2009). Arsenopyrite is the most abundant Fe-As-S species, in addition to orpiment. Scorodite, a well-crystallized iron arsenate, is a common secondary mineral in acidic environments. In typical groundwater phs (6-9), scorodite dissolves into iron hydroxides and arsenate oxyanions (H 2 AsO - 4 or HAsO 2-4 ) (Drahota and Filippi 2009). Under reduced conditions, sulfate is reduced and As can precipitate out as several different sulfide minerals. Arsenopyrite is common in reduced environments (Eh= -0.4 to -0.7 volts) and at typical groundwater ph. 2.4 Iron Mineralogy Iron oxide minerals are often the main source and sink for arsenic in groundwater (Anawar et al. 2003). Elevated As has been attributed to desorption of As from iron oxide surfaces or the release of As as a result of reductive dissolution of iron minerals. In aquatic environments, Fe(II) can be present as Fe 2+, Fe(OH) +, Fe(OH) 2 (aq), and Fe(OH) 3- ; and Fe(III) as Fe 3+, Fe(OH) 2+, Fe(OH) 2 +(aq), and Fe(OH) 3 (Bose and Sharma 2002). Iron minerals undergo redox potential changes similar to As throughout the soil profile. In surface layers, Fe(III) oxides are dominant (Table 1) with higher Fe(II)

24 6 concentrations in the transition and depletion zones. In the transition and depletion zones, Fe(II) can precipitate with carbonates or sulfide minerals, such as siderite (FeCO 3 ) or pyrite (FeS 2 ). Table 1: Several Iron(III) Oxy-Hydroxides (Cornell and Schwertmann 2003) Name Chemical Formula Bernalite Fe(OH) 3 Goethite α-feooh Ferrihydrite Fe 5 HO 8 4H 2 O Feroxyhyte Ψ-FeOOH Hematite α-fe 2 O 3 Lepidocrocite γ-feooh Magnetite Fe 3 O 4 (Fe 2+ /Fe 3+ O 4)

25 7 2.5 Arsenic Sorption Dzombak and Morel (1987) explain the importance of understanding sorption of inorganic contaminants since many inorganic compounds bind through a site-specific process. For example, As(III) and As(V) form inner-sphere complexes on iron, aluminum, and manganese oxide surfaces. It has been reported in the literature that natural organic matter, carbonates, and phosphates can compete with arsenic for sorption sites on iron oxides (Ghosh et al. 2006; Holm 2002). 2.6 Arsenic Sorption to Iron and Manganese Oxides Many studies have looked into the interactions between As and iron oxyhydroxides. Most of these experiments were performed in a laboratory under very controlled environmental conditions using synthetic minerals. Pierce and Moore (1980; 1982) and Raven et al. (1998) investigated As sorption onto amorphous iron hydroxides as affected by ph. Pierce and Moore (1982) reported maximum sorption of As(III) at ph 7 and of As(V) at ph 4. Gimenez et al. (2007) found that both As(III) and As(V) sorb more efficiently on hematite than magnetite or goethite. Sorption of both As(III) and As(V) decreased at alkaline ph (>9). In addition, As(III) sorbed similarly to both natural and synthetic oxides, with constant sorption onto goethite and magnetite from acidic to neutral ph, and decreased sorption at alkaline ph. Other studies have also shown As(V) sorption on goethite decreases with increasing ph (6-11) (Grossl and Sparks 1995) and that As(III) and As(V) sorption on goethite decreases at neutral to alkaline ph (Matis et al. 1997). Several studies also show As(V) sorption on iron oxyhydroxides follows first-

26 8 order kinetics and data properly fits a Langmuir isotherm (Gimenez et al. 2007; Matis et al. 1997; Singh et al. 1996). Dixit and Hering (2003) investigated As(V) and As(III) sorption onto amorphous iron oxide (termed hydrous ferric oxide (HFO) or ferrihydrite (5Fe 2 O. 3 9H 2 O), goethite, and magnetite under varying ph conditions. Experimental data were modeled using a diffuse double layer model and compared to previously reported data in the literature. Dixit and Hering (2003) found arsenate to sorb less to hydrous ferric oxides (HFO) and goethite at higher ph (7-8) and higher total As(V) concentrations than at lower phs (5-6). For arsenite sorption onto HFO, goethite, and magnetite, ph had a much smaller effect. Similar to arsenate, as the total As(III) concentration increased, As(III) sorption decreased. Dixit and Hering (2003) state that in ph ranges of 6-9, As (III) sorbed similarly or greater than As(V) onto HFO and goethite. Manning and Goldberg (1997) found As(V) had a higher affinity for soil than As(III) based on adsorption isotherms (ph ), contrary to sorption behavior described by Dixit and Herring (2003). Manning and Goldberg (1997) also found As(III) oxidation and increased As(III) sorption at ph>8. It is important to note that Manning and Goldberg (1997) studied sorption onto actual California soils, not pure minerals, as analyzed in the other sorption studies. These later results indicate that it is important for future studies to look at real soil samples since sorption onto them can be very different compared to a pure mineral. Smith et al. (1999) found increasing ph to have a minor effect on the concentration of As(V) sorbed in Australian soils with low concentrations of oxidic

27 9 minerals and found a decrease in sorbed As(V) at increasing ph in soils with a high concentration of oxidic minerals. These observations were attributed to an increasing negative surface potential and an increased concentration of negatively charged As(V) species with increasing oxidic mineral content. As(III) sorbed to a greater extent at increasing ph (6-7) than at lower ph (2-5) in two of the soils analyzed (high concentration of oxidic minerals). It is important to note again that sorption onto actual Australian soil samples were analyzed, not pure minerals. A summary of these studies is shown in Table 2.

28 10 Table 2: Summary of Arsenic Sorption to Iron and Manganese Oxides Studies Author(s) Dixit and Hering (2003) Pierce and Moore (1982) Raven et al. (1998) Manning and Goldberg (1997) Gimenez et al. (2007) Grossl and Sparks (1995) Matis et al. (1997) Smith et al. (1999) Sorption Media HFO, goethite, and magnetite ph Effect on As(III) Sorption As(III) sorbed similarly or greater than As(V) onto HFO & goethite at ph 6-9. ph Effect on As(V) Sorption As(V) sorbed less to HFO & goethite at ph 7-8 than at ph 5-6. HFO Maximum sorption at ph 7. Maximum sorption at ph 4. Ferrihydrite HFO Hematite, magnetite, and goethite Goethite Goethite Four Australian Soils Increasing As(III) sorption with increasing ph from ph 4-9. Sorption crosses between ph As(III) sorbed similarly to both natural and synthetic oxides. Constant sorption onto goethite and magnetite from acid to neutral ph and decreasing sorption at ph>9. As(III) and As(V) sorption decreased at neutral to alkaline phs. As(III) sorbed greater at increasing ph (6-7) than at lower ph (2-5) in two of the soils analyzed (high concentration of oxidic minerals). Decreasing As(V) sorption with increasing ph from ph 4-9. Sorption crossed between ph As(V) reduction promotes As mobility since As(III) is not as strongly adsorbed. As(III) and As(V) sorb more efficiently on hematite than magnetite or goethite. Sorption of both As(III) and As(V) decreased at ph>9. As(V) sorption decreased with increasing ph between ph Increasing ph has minor effect on amount of As(V) sorbed in soils with low concentration of oxidic minerals. Decrease in sorbed As(V) at increasing ph in soils with high concentration of oxidic minerals.

29 11 Manning et al. (2002) studied As(III) and As(V) interactions with birnessite (α- MnO 2 ) where they found evidence of As(V) sorption onto MnO 2 surfaces. They also found that the reductive dissolution of MnO 2 resulted in increased As(V) sorption as the As(III) was oxidized to As(V), creating new reaction sites on MnO 2 surfaces. 2.7 Arsenic Sorption to Carbonate Minerals Several studies (Dixit and Hering 2003; Gimenez et al. 2007; Goldberg and Glaubig 1988; Grossl et al. 1997; Wilkie and Hering 1996) have looked into As sorption onto clay and iron oxides, but few have addressed the importance of As sorption onto calcite (Sø et al. 2008). Since Utah soils are calcareous, As may readily sorb onto these carbonate minerals. Sø et al. (2008) performed a series of batch experiments to investigate As(III) and As(V) sorption onto calcite, as a function of time and solution chemistry. Sø et al. (2008) analyzed the sorption capacity of As(III) by varying the solid:solution ratio and keeping the initial As(III) concentration the same. After 24 hours, it was shown that the solid:solution ratio did not affect the aqueous As(III) concentration, even if the ph and calcium concentrations were altered. Then to understand how As(III) concentration affects sorption, As(III) concentrations between 0.14 and 4.7 μm were investigated. Regardless of the As(III) concentration used, little to no As(III) sorbed onto the calcite. As(V) was also studied under the same conditions. With a varying solid:solution ratio, the amount of As(V) in solution decreased with increasing calcite surface area, indicating As(V) sorption or surface precipitation. The As(V) concentrations were also

30 12 altered and a Langmuir isotherm fit the data. The results showed that As(V) sorption took place on a limited fraction of the available sorption sites. The effect of ionic strength on sorption was also investigated. Results showed that as the ionic strength increased, the amount of As(V) sorbed decreased. Sø et al. (2008) used Pokrovsky and Schott (2002)'s surface complexation equations to develop a surface complexation model. A calcite surface area of 44 m 2 /L and a density of 8.22 μmol/m 2 for Ca + 2- and CO 3 were used. Sø et al. (2008) claimed that by using a surface complexation model with weak and strong sites, As(V) sorption on to the calcite surface can be modeled successfully. Yokoyama et al. (2012) found similar results to Sø et al. (2008). In coprecipitation experiments with As(III) and As(V) onto calcite, As(V) was preferred over As(III) at ph Yokoyama et al. (2012) also found that over time, the As(III)/As total ratio decreased, ultimately resulting in only As(V) sorbed on calcite. Goldberg and Glaubig (1988) investigated As(V) sorption onto calcareous soil and concluded As(V) sorption increased with increasing ph (maximum near ph 10.5). They used a constant capacitance model to describe As sorption onto the soil. 2.8 Desorption of Arsenic from Mineral Surfaces: Bicarbonate Kim et al. (2000) analyzed the role of bicarbonate in As solubilization into groundwater by investigating the effects of ph and redox conditions. Core samples of Marshall Sandstone, the primary potable water aquifer in southeast Michigan, had elevated As levels in groundwater under reduced conditions. A variety of solutions, including KCl, Na 2 SO 4, MgSO 4, CaSO 4, NaHCO 3, KHCO 3, and FeCl 3 were used to determine the major ions that leached As most efficiently. It was found that bicarbonate

31 13 (NaHCO 3 ) leached As the greatest in air-saturated distilled water, and the smallest amount of As was released in anoxic groundwater. Additional work also looked at As leaching at varying phs. Kim et al. (2000) demonstrated a strong relationship exists between the release of As and the bicarbonate concentration in the leaching solution. 2.9 Natural Organic Matter (NOM) Some literature suggests natural organic matter (NOM) plays a significant role in influencing As speciation and mobility (Anawar et al. 2003; Tseng et al. 1968). For example, Anawar et al. (2003) suggested sedimentary organic matter and peat soils were related to high levels of As in Bengal sediments, while Tseng et al. (1968) found that water from organic black sediments from a lagoon in Taiwan had high levels of As as well. Redman et al. (2002) and Welch and Lico (1998) both found evidence that through redox reactions, NOM can change As speciation and form NOM-As complexes. In addition, NOM has been found to reduce As(V) when associated with metal hydroxides and form complexes with As(III) and As(V) (Redman et al. 2002). One of the most important relationships between NOM and As is the competitive sorption onto mineral surfaces. Humic and fulvic acids may compete with As(III) and As(V) for sorption sites which can influence As mobility (Wang and Mulligan 2006). Table 3 provides a brief summary of NOM-As related research and the main conclusions of each study.

32 14 Table 3: Summary of NOM-As Related Research (Adapted from Wang and Mulligan (2006); includes Cornu et al. (1999), Grafe et al. (2001), Redman et al. Organic Matter influence on As(V) sorption to clay minerals NOM effects on As sorption to hematite (2002), and Thanabalasingam and Pickering (1986). Subject Main Conclusions References As(V) sorption increased in the presence of a humic acid Cornu et al. (1999) coated in kaolinite. NOM competes with As for active sorption sites in Redman et al. (2002) natural waters. Effect of humic and fulvic acids on As(III) and As(V) sorption onto FeOOH As(III) and As(V) sorption by two humic acids Decreased As sorption to FeOOH due to organic acids Concentration greatly influenced As sorption to humic acids. Grafe et al. (2001) Thanabalasingam and Pickering (1986) 2.10 Arsenic Reductive Microbial Processes One of the main As release mechanisms from soils and aquiver material is microbial reductive dissolution of Fe(III) oxides (Bennett and Dudas 2003; Cummings et al. 1999; McArthur et al. 2001; Nickson et al. 2000) with the concurrent release of As(V) and Fe(II). Smedley and Kinniburgh (2002) and Islam et al. (2004) however, state that As(V) reduction could occur after Fe(III) reduction, not simultaneously. Horneman et al. (2004) and van Geen et al. (2004) also found As and Fe release did no co-occur. After 2 months of incubation, Horneman et al. (2004) and van Geen et al. (2004) found most of the leachable As and a small percentage of the Fe were released.

33 Reductive Dissolution of Iron Oxides Many arsenic studies in Bangladesh and West Bengal (McArthur et al. 2001; Nickson et al. 2000) concluded that microbially induced reductive dissolution of iron oxides results in arsenic release. Under anoxic conditions at field sites and in laboratory studies there is an observed dependency of As release on the microbial reductive dissolution of Fe(III) oxide minerals (Cummings et al. 1999; Langner and Inskeep 2000; Masscheleyn et al. 1991). This mechanism is proposed as the process that drives As solubilization in groundwater in SE Asia (Fendorf et al. 2010; Islam et al. 2004). The released As(V) may be resorbed onto minerals, precipitate as a new solid phase, or dissimilatory arsenic reducing bacteria (DARB) may reduce the As(V) in solution to As(III) Direct Arsenic Reduction DARB can directly reduce As(V) without the dissolution of Fe minerals (Ahmann et al. 1997; Islam et al. 2004). Native bacteria can directly reduce As(V) to As(III) for energy production (Krafft and Macy 1998; Oremland and Stolz 2005) through the arsenate respiratory reductase (ArrAB) enzyme using As(V) as an electron acceptor (Ahmann et al. 1997; Laverman et al. 1995). Campbell et al. (2006) found As(V) reduction to occur simultaneously or before Fe(III) reduction occurred. Islam et al. (2004) used a microcosm study to analyze the role of bacteria in arsenic release. They found arsenic release did not occur simultaneously with Fe(III) reduction. Instead, the arsenic release took place after Fe(III) reduction. The sequence of terminal electron acceptor usage reported in the literature is dependent on source of test materials (laboratory

34 16 generated minerals, sediments from different locations), bacteria used (pure cultures vs native), and experimental design (carbon source) Abiotic Redox Processes Bose and Sharm (2002) examined the effect of ph and redox potential on As associated with synthetic iron hydroxides. They found that in an oxidized environment, As(V) was the dominant species and mobilization was relatively low. In moderately reducing conditions, there was a partial reduction of As(V) to As(III) and mobilization increased. In the most reduced setting, As(III) was the dominant form and mobilization was the greatest. Though without abiotic controls, they claimed the As and Fe in this study were not exclusively reduced microbially but state that abiotic processes are responsible for some of the Fe and As mobilization Alternating Redox Conditions Alternating reducing/oxidizing conditions may play a major role in As dissolution. Jung et al. (2012) describe the importance of sediment redox state in arsenic partitioning through sorption/desorption experiments and partition coefficient (Kd) calculations. Fendorf et al. (2010) explain how destabilized As on iron oxides is a main factor in As release, including how reductive dissolution of Fe(III) oxides releases As, as well as As(V) reduction to more mobile As(III) (Islam et al. 2004). Sorption and desorption are also directly influenced by the oxidation state of As and Fe species. For example, Stollenwerk et al. (2007) studied how the oxidation state influences As sorption. They found the sorption capacity to be a function of the oxidation state and concentration of solutes competing for sorption sites. Stollenwerk et al. (2007) found

35 17 As(III) often sorbs less than As(V) in oxidized environments. MacKay et al. (2013) analyzed the impact seasonal changes have on As and Fe accumulation. Due to redox cycling, they found that during the summer (July-October), the greatest accumulation of As and Fe occurred and during the spring months (March-June) when water levels were the highest, the lowest rates occurred. Munk et al. (2011) analyzed eight private drinking water wells with As levels over 10 μg/l in Anchorage, Alaska to determine the As(III) and As(V) concentrations and the impact of seasonal groundwater fluctuations on measured As concentrations in the impacted wells. The drinking water wells had the highest water levels during the Spring and Fall seasons due to increased precipitation and snowmelt. The groundwater is classified as Ca-Mg-HCO 3, with a ph ranging from 7.6 to 8.8. Due to reducing conditions, the dominant As state was As(III) and dominant iron oxidation state was Fe(II). It was determined that increasing As concentrations were a function of increasing water levels, implying the highest As levels occurred during groundwater recharge events. Munk et al. (2011) also demonstrated a positive relationship between As and dissolved iron, as well as between As and SO 2-4, indicating As may be associated with - Fe-As-S minerals at this site. An additional relationship was seen between As and HCO 3 and the groundwater was found to be supersaturated with respect to HCO - 3. Beauchemin and Kwong (2006) also looked at the influence of alternating redox conditions on As release. Using soils contaminated by mine tailings, they found that a portion of the As(V) species transformed into As(III) after 30 days under reducing conditions. Though As(V) species dominated on day 0, by day 30, a portion is it reduced

36 18 into a more mobile species, As(III). In addition, the initial surface concentration of oxalate-extractable As, as well as the As in a glucose-amended layer, significantly decreased after reoxidation. These results were interpreted as meaning a small portion of As in the solid phase is irreversibly converted to a soluble solid phase after a full redox cycle Geochemical Equilibrium Modeling Environmental geochemical equilibrium modeling can be used to address various types of environmental problems, including waste disposal, mining drainage, and groundwater contaminant transport. Equilibrium modeling includes precipitation and solution speciation, as well as sorption models. Sorption models help describe the interactions between solutes and particulate surfaces and explain the solute concentrations. Variations in sorption models include the numbers of sorption layers and types of species sorbed in each layer (Schecher and McAvoy 1992). The models used in this study MINTEQ (Gustafsson 2011) and PHREEQC (Parkhurst and Appelo 1999) are two of the many equilibrium models available Modeling Inorganic Species Precipitation and Dissolution Reactions In general, geochemical models use the tableau approach to represent chemical formation reactions. A tableau includes each chemical component of interest, total concentrations, and corresponding thermodynamic constants. The constants are used as a starting point for defining the system. Given chemical concentrations, the geochemical model can predict which solids will form, which compounds will remain as dissolved solids, and the corresponding concentration of each.

37 Double Layer FeOH Model Wilkie and Hering (1996) investigated As(III) and As(V) sorption onto HFO using different As:total Fe ratios over a ph range of 4-9. The FeOH sorption model worked well when using total As:total Fe ratios previously studied in the literature (1.33 μm total As: 50 μm total Fe). There were, however, discrepancies between observed and predicted As(III) sorption when using a lower total As concentration Uncertainty in Equilibrium Models Equilibrium models do not include any data or information regarding the rate of a reaction. MINTEQ (Gustafsson 2011), similar to MINEQL+ (Schecher and McAvoy 2003), is limited by only allowing a system to include 25 components, including surface sorption and columbic components. A decision must be made whether or not to include a particular species or process, increasing model uncertainty if an important input is neglected. Additional uncertainty stems from mineral phase thermodynamic constants (Sracek et al. 2004). Uncertainty may include a model's thermodynamic equilibrium data, ionic strength adjustments, or component interactions, especially if a system is not completely at equilibrium. Since uncertainty can never be completely eliminated, the goal of any model is to minimize it Successful Geochemical Modeling Applications in Arsenic Research Hussam et al. (2003) employed MINEQL+ to determine the fate of As and other heavy metals in Bangladesh groundwater. As(III), As(V), metals, anions, and basic groundwater parameters from naturally attenuated groundwater and fresh groundwater were input into MINEQL+. Sorption, removal amounts, and speciation were determined.

38 20 Hussam et al. (2003) determined that 65% of the As was in a soluble form and 35% was removed through sorption. Dixit and Hering (2003) compared As(III) and As(V) sorption onto HFO, goethite, and magnetite using a diffuse double layer model. FITEQL (Westall 1982), a program used to determine equilibrium constants from experimental data, was utilized to estimate As(III) and As(V) surface complexation constants. MINEQL+ was used to model the As(III) and As(V) sorption. Dixit and Hering (2003) concluded that under typical groundwater ph (6-9), As(III) sorbs to a similar extent, if not greater, onto HFO and goethite to that of As(V). Several articles (Goldberg and Glaubig 1988; Sø et al. 2008) have also used surface complexation modeling to describe As sorption onto calcite. Romero et al. (2004) researched arsenic retention and processes controlling As release in a carbonate rich aquifer. MINTEQA2, an equilibrium modeling program used to determine aqueous chemical compositions, was used to model As sorption onto HFO. A constant capacitance model, instead of the diffuse double layer model, was used since it was a better fit to the experimental data. Romero et al. (2004) concluded that since MINTEQA2 modeled the As retention and sorption so well, HFO may be the main mineral species responsible for As sorption. Sø et al. (2012) successfully modeled the sorption of As(V) to calcite using the triple plane model within CD-MUSIC, including modeling the competitive sorption between As(V) and phosphate. The sorption of As(III) onto calcite was not examined in this paper since their previous work displayed that As(III) does not interact with calcite surfaces (Sø et al. 2008).

39 21 CHAPTER 3 STUDY BACKGROUND Anning et al. (2012) used a statistical approach to predict arsenic levels across the basin-fill aquifers in the Southwestern U.S., and estimated that approximately 43% of the area studied has As levels equal to or above the MCL (Figure 1). Many of the areas with elevated As levels are rural and sparsely populated. These areas, however, are located next to major metropolitan areas, including Salt Lake City, UT. This means that with growing populations and an increased drinking water demand, future groundwater development may be affected (Anning et al. 2012). This implication is significant and demonstrates the importance of being able to predict arsenic retention and release into groundwater in this semi-arid region.

40 22 Figure 1: Arsenic (μg/l) Prediction at 200 Foot Aquifer Penetration Depth Southwestern Basin (Anning et al. 2012) Cache Valley Basin (41 42' N / ' W) is located in Northern Utah and the southeastern part of Idaho. The Wellsville Mountains lie to the west and the Bear River Mountains, while an extension of the Wasatch Mountain Range lie to the east. The Cache Valley Basin, as well as a portion of the Great Basin Region (Figure 2) was once covered by Lake Bonneville (Currey 1990), a Pluvial lake, which receded at the end of the Pleistocene and onset of the Holocene era, about 14,500 years ago. The benches and deposits remaining from Lake Bonneville left Quaternary freshwater lake deposits. The parent material is dominated by lacustrine and alluvial sediments. Cache Valley has a semi-arid, continental climate, with a xeric soil moisture regime and mesic soil temperature regime. The Salt Lake Formation underlies the Lake Bonneville deposits. This formation consists of tuffaceous sandstones and siltstones, limestone, and

41 23 conglomerates, and is exposed at high elevation in the surrounding mountains. The volcanic origin of this Salt Lake Formation, including the presence of As minerals, enargite (Cu 3 AsS 4 ) and tennanite (Cu 12 As 4 S 13 ), may be the source of As in the basin. Several surveys of groundwater wells within the Cache Valley Basin (41 42' N / ' W) revealed that about 17% of the groundwater exceeds the As MCL (Sanderson and Lowe 2002). Well water collected from the area north of the Logan City Municipal Landfill also had concentrations of As above the MCL. This area was selected to study for ease of access since the property is owned by the State of Utah and Logan City. In addition, it is important to note that the groundwater is not influenced by the landfill (UWRL 2009) (Figure 3). Figure 2: Southwestern U.S. Principal Aquifers and Basin Locations (Anning et al. 2012)

42 24 Figure 3: Sampling Sites, NP 9 and NP 13, Near the Logan Landfill In 2008, 13 cores were collected from 60 cm above to 90 cm below the groundwater table (approximately 2.5 to 4.5 m below ground surface, transition to depletion zones) to determine As concentrations in aquifer solids up-gradient of the landfill. In 2010, five surficial cores were collected from the soil surface to 1.7 m below ground surface. The major finding of this study showed that As was distributed throughout the soil column, from the soil surface to the depth of groundwater (Meng 2015). Aquifer material within each core was divided into two to up to four sections depending on observed redoximorphic features, including color and texture. All sample processing was performed in a nitrogen-filled anaerobic chamber. To determine the solid phases associated with Fe and As, a six-step sequential extraction procedure was performed along with a separate HCl extraction, with analysis of Fe and As oxidation

43 25 states in each extract solution. Surface soils had total As concentrations ranging from 2,000 µg/kg to 18,000 µg/kg, as determined by a total digestion using nitric acid, indicating As was not secluded in the deep sediments. The majority of the arsenic in the aquifer solids was associated with the crystalline iron oxides and the residual fractions. However, carbonate minerals also had associated As. Carbonates are common minerals in semi-arid and arid environments and are thought to provide sorption sites for As, as they do for phosphate, which behaves similarly (Salingar and Kochva 1994). Based on these preliminary findings, further studies were conducted on continuous cores at two sites by Meng (2015). These studies considered pore water chemistry, mineral description by sequential extractions, Quantitative Evaluation of Minerals by Scanning Electron Microscope (QEMSCAN), and X-ray Absorption Spectroscopy (XAS). Sampling, analysis, and results in Meng (2015) resulted in four distinct sections (Figure 4), which can be generally described as follows: a. Vadose Zone: carbonates present, unsaturated, shallow, oxidized region with relatively low concentrations of As. b. Water Table Zone: carbonate nodules, water table level fluctuation between oxidizing and reducing conditions. c. Transition Zone: redoximorphic features present, fluctuation between oxidizing and reducing conditions. Under slightly reduced conditions, iron oxides and hydroxides undergo reductive dissolution, releasing As (Sracek et al. 2004).

44 26 d. Depletion Zone: mineral depletion, greenish color, a highly reduced region, where As co-precipitates with sulfides, forming secondary minerals like pyrite and orpiment in regions with moderately high concentrations of H 2 S. A B Figure 4: Schematic of NP 9 (A) NP 13 (B) Soil Profile Depths and Zones As described by Meng (2015), the near-surface soils (vadose zone) contain labile As due to the oxidation of primary As minerals associated with volcanic formations. In deeper sediments (transition zone), As transports downward and associates with metal oxide minerals, and in the permanently depleted zone, reduced As is present as As

45 27 sulfides. The processes controlling As in the water table zone, however, remain unknown. This study investigated these water table zone processes. 3.1 Hypotheses and Objectives Overall hypothesis: Arsenic, under strong oxidizing or reducing conditions, is associated with stable solid phases, whereas As within zones of seasonally altering redox conditions, due to fluctuating groundwater levels, is unstable and may be the source of As in the groundwater (Figure 5). The following objectives were used to test the hypothesis: 1. Determine the mineral phases that control As solution chemistry down the profile using geochemical modeling and data from sequential extracts of the solids. Aqueous specie inputs were determined based on pore water and solid phase As and Fe concentrations. 2. Determine the effect of altering redox conditions within the water table zone on As mineralogy using a laboratory microcosm study and verify model predictions.

46 28 Figure 5: Soil Profile Sketch. Larger Ellipses Indicate Higher Pore Water and Solid Phase Concentrations

47 29 CHAPTER 4 MATERIALS AND METHODS 4.1 Objective 1- Geochemical Modeling Characterization of Collected Soil A full description of sample collection, processing and analysis is given in Meng (2015). Briefly, two continuous cores, from the soil surface to the depth of groundwater, were collected at NP 9 and NP 13 (Figure 3) in September Meng (2015) sectioned one core in the field from each site for general division of zones based on color and texture. These cores were discarded. Upon returning to the lab, the undisturbed cores were divided into sections based on field observations, then further divided into sections observable (based on color and texture) in the lab. All laboratory sample handling and storage was in an anaerobic glovebag under nitrogen (Coy Lab Automatic Air Lock Anaerobic Chamber). NP 9 was sectioned into 16 layers: three layers in the vadose zone, two in the water table zone, six layers in the transition zone, and five layers in the depletion zone (Figure 4 (A)). NP 13 was sectioned into 14 total layers; two in the vadose zone, two in the water table zone, six in the transition zone, and four in the depletion zone (Figure 4 (B)). To facilitate homogenization of each section, the soil was placed into a plastic bag and double deionized water (DDW) was added to saturate the soil. Soil moisture content was determined for the field soil and after mixing. Due to the high clay percentage in NP 9 and NP 13 (Table 4), the pore water was not extractable through centrifugation alone even when water was added for mixing. To extract the water from the solids, 25 grams of wetted soil and an additional 10 ml of DDW were mixed together. The slurry was put

48 30 into spin filters (0.2 µm filter, 50 ml UltraClean Maxi Plasmid Prep Kit) and centrifuged for 20 min at 10,000 g. The extracts were then filtered through a 0.2 µm filter inside the glovebag and analyzed for As and Fe redox species, major cations and anions and trace elements (analytical methods are provided in Appendix A, Table A. 1). Hydrogen sulfide (HS) was determined from total sulfur determined by Inductively Coupled Plasma Mass Spectrometry (ICP-MS) and sulfate determined by ion chromatography (IC), the difference being assumed to be all HS. The analyzed concentrations were then back calculated to pore water concentrations by taking into account the field moisture, amount of water added in the lab, and the estimated amount of water the soil can hold at saturation based on soil texture (Schroeder et al. 1994). To determine the fraction of As and Fe associated with: 1) ligand/cation exchangeable, 2) carbonates, 3) acid volatile sulfides, Mn oxides, and very amorphous Fe oxides, 4) amorphous Fe oxides, 5) sulfides, 6) crystalline Fe oxides, and 7) residual phase minerals, a seven step sequential extraction method was performed (Amacher 1988; Huang and Kretzschmar 2010; Keon et al. 2001). The very amorphous Fe oxides from the third step differ from the amorphous Fe oxides in the fourth step in that they are the portion of amorphous Fe oxides with a loose structure and higher solubility. This is the portion that is more bioavailable than the amorphous Fe oxides extracted in the fourth sequential extraction step. The extraction method was specifically developed to preserve As speciation. The extraction method is detailed in Appendix A (Table A. 2). In addition, a separate 0.5 M HCl extraction was performed to determine the As and Fe speciation in acid soluble

49 31 mineral phases. The oxidation state of Fe and As were determined in Steps 2-3 and Steps 1-4, respectively. In the other steps, only the total Fe and As were measured. 4.2 Model Selection PHREEQC was selected over MINTEQ to predict the solid phases and oxidationreduction potential (Eh) throughout the profile due to the larger database of solids and aqueous species stored in PHREEQC. MINTEQ was selected to model sorption onto FeOH and CaCO 3 due to the ease of modeling within the program and the ability to add CaCO 3 sorption parameters into the model. 4.3 Objective 1 Data Analysis PHREEQC Modeling Using measured major anion, major cation, and trace element pore water concentrations, soil ph, soil electrical conductivity, and groundwater alkalinity data, PHREEQC was used to determine precipitates, dissolved species, and Eh for each sediment layer to understand which minerals play an important role in As release and retention. The pore water concentration input values for NP 9 and NP 13 are located in Appendix B (Table B. 1 Table B. 4). Pore water alkalinity was not available, so groundwater alkalinity data were used instead and assumed representative. The soil ph and groundwater alkalinity were used to calculate the carbonate concentration. Eh values were not available, so they were estimated based on the input parameters. Stability constants used in PHREEQC were from the MINTEQ v.4 database. Alteration to the database is described in the 4.5 Model Alterations section.

50 32 In the same run used to predict solid phases, the Eh, in volts, was calculated using As(III)/As(V), NO 2 /NO 3, Fe(II)/Fe(III), and HS - /SO 4 redox pairs in PHREEQC. For example, each pair was uniquely calculated using the ratio between the species concentrations and the solution ph. There are limitations in modeling and using redox pairs in calculating Eh (Cherry et al. 1979). Some redox reactions involve N, C, or S, which are biologically driven and cannot be measured using redox pairs. Furthermore, in natural waters, which include groundwater, often one of the redox species is below a detection limit and unsuitable for the redox pair method (Cherry et al. 1979). For example, the Fe 2+ concentration may be below the method detection limit (MDL) when using the Fe 2+ /Fe 3+ redox pair. 4.4 MINTEQ Modeling MINTEQ was used to model As specie distribution and the amount of As sorbed onto FeOH and CaCO 3 in the water table zone using a two-layer HFO sorption model and 1-pK Three Plane Model (TPM). Input values for NP 9 and NP 13 are in Appendix B. A surface area of 600 m 2 /g for HFO (Dzombak and Morel 1990) and a solids concentration (g L -1 ) calculated from the total Fe and solids concentration (Equations 1, 2 and 3) were used as model input (Dzombak and Morel 1990). To model CaCO 3, the calcite surface area was assumed to be between 10 and 70 m 2 /g (Wei et al. 1997). Using a more conservative area, a surface area of 15 m 2 /g was used in MINTEQ. The solids concentration was determined from the percent of CaCO 3 in each layer and converted to moles of Ca.

51 33 MINTEQ was used to determine the percent distribution of As species and predict As sorption using a two-layer HFO model and 1-pk TPM. The concentrations of As(III) and As(V) associated with the very amorphous Fe oxides, as defined by sequential extractions, were used as the input concentrations. MINTEQ also calculated the precipitated solids throughout the profile. All solids were allowed to precipitate. Inputs included pore water concentrations for the aqueous species, as well as the concentration of Fe extracted from the very amorphous iron and amorphous minerals (TOTFe) for Fe(III). Using terminology from Dzombak and Morel (1990) to describe the two-layer model, the total concentration of strong (st) HFO binding sites is calculated from the following equation: [Fe(st)OH] T = mol/mol Fe x TOTFe (mol/l) (1) The total concentration of weak (wk) HFO binding sites is calculated from the following equation: [Fe(wk)OH] T = 0.2 mol/mol Fe x TOTFe (mol/l) (2) The solids concentration (g/l) is defined as: TOTFe (mol/l) x 89 g/mol (3) The FeOH sorption model, CaCO 3 sorption model, and precipitation were all run simultaneously. The 1-pk Triple Plane Model (TPM) is the default sorption model for arsenate used in MINTEQ. Parameters for the model are given in Sø et al. (2012). The model accounts for the extent of protonation of surfaces of clays, oxides and carbonates. The

52 34 model considers three planes and a diffuse double layer to describe the charge distribution of the sorbed ions. 4.5 Model Alterations Barium arsenate was suggested as one of the solid phases controlling aqueous arsenic concentrations (Wagemann 1978). Barium arsenate is insoluble, with a solubility product constant of approximately 7.7 x in fresh water at 25 C (Frankenthal 1963). Numerous geochemical equilibrium programs (including MINEQL+, USGS Geochemical Software, and PHREEQC) use the above barium arsenate solubility product. It is assumed the MINTEQ database does not include this barium arsenate solubility product since it did not precipitate in any layers. There are, however, two more recent articles (Essington 1988; Zhu et al. 2005) which suggest the barium arsenate solubility product 7.7 x is incorrect. Zhu et al. (2005) reported the barium arsenate solubility product to be and Essington (1988) found the solubility product constant (Ksp) to be These values vary from the historical Ksp by about 28 log units and suggest that barium arsenate is less stable than previously thought. Abul Hussam (personal communication, August 16, 2012) suggested that since H 2 AsO - 4 and HAsO 2-4 are the dominant As species at typical groundwater ph, barium will precipitate with these species instead of AsO 3-4. Hussam used MINEQL+ to determine As speciation and sorption onto HFO in groundwater from Bangladesh. He found the Ba(H 2 AsO 4 ) 2 solid was one of the precipitated forms of As removing As from the groundwater (5%-15% depending on the As concentration), in addition to Fe(wk)OH-

53 35 AsO4. After further communication with Hussam (Abul Hussam, personal communication, August 17, 2012), he explained that he altered the Ba-As species in MINEQL+ to Ba(H 2 AsO 4 ) 2, but still used the Ksp for Ba 3 AsO 4 (Hussam et al. 2003). Since an incorrect Ksp value was used, Hussam et al. s (2003) claim that arsenate is removed as Ba(H 2 AsO 4 ) 2 instead of Fe(III), cannot be validated. For these reasons, the Ksp of Ba 3 AsO 4 was adjusted in PHREEQC to (Essington 1988) and used throughout the study. 4.6 Objective 2 Redox Experiment Sediments from NP 9-5 and NP 13-3 were used in the following experiments. These sediments were selected because they were located at the fluctuating water table levels in each core and had the highest As pore water concentrations. Similar mechanisms but different concentrations were expected between the two sediments due to the difference in location and soil properties (Table 4). For example, NP 13-3 is more alkaline and has a higher soil electrical conductivity (EC) than NP 9-5. However, the aqueous and 0.05 HCl extractable total As and As redox species concentrations in these layers are similar. Properties of the selected aquifer solids and groundwater are given in Table 4 and Table 5 Sequential extraction results are shown in Table 6 and Table 7.

54 36 Table 4: Summary of NP 9-5 and NP 13-3 Parameters NP 9-5 NP 13-3 ph EC (µs/cm) As(III), µg/l, pore water As(V), µg/l, pore water Total As, µg/l, pore water Fe(II) mg/l, pore water Fe(III) mg/l, pore water Total Fe, mg/l, pore water HCl Ex. As (III), µg/kg HCl Ex. As (V), µg/kg HCl Ex. Total As, µg/kg HCl Ex. Fe(II), mg/kg HCl Ex. Fe(III), mg/kg < HCl Ex. Total Fe, mg/kg Sand % 3% 0% Clay % 57% 65% CaCO 3 % 32.8% 29.7%

55 37 Table 5: NP 9 and NP 13 Groundwater Properties ph EC, As, Fe Total Alkalinity, µs/cm µg/l µg/l mg/l CaCO 3 NP NP Table 6: Sequential Extraction Results, As associated minerals NP 9-5 As(III) µg/kg NP 9-5 As(V) µg/kg NP 9-5 % distribution NP 13-3 As(III) µg/kg NP 13-3 As(V) µg/kg NP 13-3 % distribution Ligand Exchangeable As % % Carbonate Minerals % % Acid volatile sulfides, Mn oxides and % % very amorphous Fe oxides Amorphous Fe oxides % < % Sulfides % % Crystalline Fe oxides 907 8% 639 8% Residual % % Table 7: Sequential Extraction Results, Fe associated minerals NP 9-5 Fe(II) mg/kg NP 9-5 Fe(III) mg/kg NP 9-5 % distribution NP 13-3 Fe(II) mg/kg NP 13-3 Fe(III) mg/kg NP 13-3 % distribution Exchangeable Fe % 3 0% 1 0% Carbonate Minerals % % % Acid volatile sulfides, Mn oxides % % and very amorphous Fe oxides % Amorphous Fe oxides % 288 1% 359 1% Sulfides % % % Crystalline Fe oxides % % % Residual % % % A short-term preliminary Eh study was conducted to determine if As release would occur over a short period (7 days) under reduced conditions and to establish appropriate sample intervals for the main study. The preliminary study was performed

56 38 with two sacrifice times, Days 0 and 7. A summary of the results is located in Appendix C (Table C.1- Table C. 18). The redox experiment was designed to investigate two treatment factors: (1) redox environment (oxidized or reduced), and (2) time (length of exposure). To estimate experimental error, the experiment was performed in triplicate. Soil from cores taken at NP 9-5 and NP 13-3 was stored, since the time of collection, in an anaerobic glovebag filled with nitrogen (Coy Lab Automatic Air Lock Anaerobic Chamber). Inside the nitrogen filled glovebag, 5 gram samples (+/ g) from NP 9-5 and NP 13-3 were placed into individual 50 ml Fisherbrand Higher-Speed disposable plastic centrifuge tubes. Thirty milliliters of filter-sterilized groundwater were added to each tube, capped tightly, and shaken gently. Groundwater was collected at Monitoring Well 1 (MW 1), which is located at the Logan City Landfill, near NP 9-5 and NP Groundwater from MW 1 was selected due to its proximity to NP 9 and NP 13, and low concentration of As. MW 1 has an average (from 2008 and 2009 sampling events) ph of 7.69, alkalinity of 315 mg CaCO 3 /L, 2.3 µg/l total As (all of which is As(III)), 502 mg/l TDS, and EC of 530 µs/cm. MW 1 has lower alkalinity, EC, and As concentrations than NP 9 and NP 13 groundwater. The lower EC and alkalinity may have contributed to initial mineral solubilization. Additional centrifuge tubes were also filled with 30 ml of DDW to serve as sample blanks. The blanks were sampled three times throughout the experiment and no contamination was noted. All sample containers and solutions were stored in the glovebag for 2 weeks prior to the experiment to remove oxygen. These solutions included the groundwater, 0.5 M

57 39 HCl, and acetate extraction solutions. Both the reduced and oxidized samples began in the glovebag with deoxygenated solutions and containers. Half of the samples were kept in the glovebag as the reduced group and the remaining half were kept on the lab bench as the oxidized group. Samples from each environment were sacrificed to establish initial conditions (Day 0). The remaining samples in the glovebag were bubbled with a 95% nitrogen and 5% hydrogen gas mixture for 20 minutes to purge any oxygen in the sample tube. The samples under oxidized conditions were bubbled with compressed air. The gas mix and compressed air gas cylinders were both fit with a 0.45 µm gamma irradiated Sterivex-HV in-line filters to prevent any non-native microbes from entering the samples. In addition, all components downstream of the filter, including tubing, glass pipettes, and fittings, were autoclaved for 20 minutes at 121ºC for sterilization. All samples were tightly capped, placed horizontally, and stored in a constant temperature room (15ºC) in the dark. Roughly, every 2 weeks, each microcosm was opened and bubbled with the corresponding gas for 20 minutes to re-expose sediments to initial treatment conditions (Figure 6). On days where sampling and bubbling both occurred, the sampling was performed prior to bubbling. Samples were sacrificed at t = 0, 2, 7, 15, 28, 63, 77, and 113 days (Figure 6). At each sampling period, triplicate samples from the reduced and oxidized treatments were removed and centrifuged at 7,000 rpm for 20 minutes to separate the aqueous and solid phases. Then, inside the glovebag (Coy Lab Automatic Air Lock Anaerobic Chamber) filled with 95% nitrogen and 5% hydrogen, the samples were filtered through a 0.2 µm

58 40 nylon filter (Environmental Express Hydrophobic Nylon Syringe Filter) and Fe(II) was determined using the ferrozine analysis procedure (Lovley and Phillips 1987) using a Thermo Fisher Scientific Genesys 10 Vis spectrophotometer at 562 nm. Arsenic speciation was also preserved in the glovebag with 5% v/v of 0.25 M EDTA inside 500 µl Thermo Scientific National Target DP vials (Bednar et al. 2002). Arsenic species were analyzed within 8 hours. Outside the glovebag, the Eh, ph, and EC were measured. Subsamples were taken for anion (sulfate and nitrate) analysis and the remaining sample acidified with 1% HNO 3 (trace metal grade) for macro cation and trace element analysis by ICP-MS. = Sampling = Bubbling Figure 6: Sampling and Bubbling Dates. Red circles indicate sampling and yellow stars represent bubbling.

59 41 Anion samples were filtered through a 0.2 µm nylon filter and placed in 0.5 ml Dionex PolyVials with filter caps. Analysis was carried out using a high-capacity, hydroxide-selective, anion-exchange column (IonPac AS18) on a Dionex Reagent-Free Ion Chromatography (RFIC) System (Jones 2001). Eh was measured using a Fisher Scientific metallic combination electrode (Ag/AgCl), following Water Environmental Federation Standard Methods (APHA 1998). ph was determined using a Fisher Scientific Accument Excel XL25 Dual Channel ph/ion Meter (Kopp et al. 1979). EC was measured using a Fisher Scientific Accument Model 30 Conductivity Meter, standardized in 0.01 M KCl solution (Kopp et al. 1979). As speciation was determined by separating As(V) and As(III) on a C-18 column with an isocratic elution with 5 mm tetrabutylammonium hydroxide, 3 mm malonic acid in 5% methanol at ph 5.9 with an Agilent 1200 Series high performance liquid chromatograph (HPLC), followed by detection using an Agilent 7700 ICP-MS. Major cations and trace elements, including total Fe (Fe(II)+Fe(III)) and total As (As(III)+As(V)), were determined following U.S. EPA Method 6020 using an Agilent ICP-MS (US-EPA 1986). After preparation of the aqueous samples, the oxidized and reduced samples were returned to the N 2 glovebag for HCl and acetate extractions. The samples were returned to the N 2 glovebag to avoid altering As and Fe redox species due to the presence of hydrogen in the other glovebag. The sediments were thoroughly mixed and 0.5 M HCl and 1 M NH 4 OAc extractions were performed on the solid phase to determine the oxidation state of Fe and As acid soluble minerals (FeCO 3, FeS, and very amorphous Fe oxides) and carbonate associated Fe and As minerals. For this procedure, 20 ml of 0.5

60 42 M trace metal grade HCl (Thermo Fisher Scientific) and 20 ml of 1 M NH 4 OAc at ph 5 were each added independently to approximately 0.7 g (dry weight) of soil and shaken for 2 hours in the glovebag (both the oxidized and reduced samples). Afterwards, the samples were centrifuged for 20 min at 7,000 rpm and the supernatants filtered through 0.2 µm filters into snap-cap vials inside the H 2 /N 2 glovebag. Ferrozine analysis was performed immediately and As was preserved with 5% v/v of 0.25 M EDTA. The remaining extracts were preserved for major cation and trace elements analysis by ICP- MS analysis. 4.7 Poisoned Controls Although abundant research (Cummings et al. 1999; McArthur et al. 2001; Nickson et al. 2000) describes microbial reductive dissolution as a major mechanism for As release, Bose and Sharm (2002) suggest abiotic process also play significant roles in As release. To determine if microbiology played a role in As release, a set of poisoned controls were set-up. The controls were prepared identical to the biotic samples, but were dosed with 500 mg HgCl 2 /kg soil mercuric chloride (HgCl 2 ), similar to the dose used in Wolf et al. (1989). Previous research (Lee et al. 1992; Tuominen et al. 1994; Wolf et al. 1989) found significant reduction in microbial populations using mercuric chloride. The controls were set-up and analyzed independent of the regular samples. An initial control experiment using autoclave-sterilized sediments was not successful in stopping all microbial activity. For this reason, HgCl 2 was used instead. NP 9-5 controls were sampled on Days 0, 7, and 21. Due to a shortage in aquifer materials, NP 13-3 controls were only sampled on Days 0 and 21. The control experiment was shorter in

61 43 length because it was assumed that after 21 days, a difference between As release in poisoned and microbially active samples would be present. In hindsight, the control experiment would have served a greater purpose by describing and confirming As behavior had the amount of days and sampling periods been increased. Day 0 samples were immediately shaken on the shaker for 15 minutes after HgCl 2 addition and then analyzed. The same analyses were performed on the controls as the biologically active samples, with the exception of the 1 M NH 4 OAc extraction and anion analyses. On Day 21, samples were analyzed for cellular adenosine triphosphate (catp) from living microorganisms using a Quench Gone Aqueous ATP test kit (LuminUltra Ltd.) and Luminometer (PhotonMaster ). The catp results were used to confirm the absence of microorganisms (Appendix I). 4.8 Quality Control and Statistics To monitor, control, and quantify analytical error, triplicate samples, calibration checks, procedural blanks, reagent blanks, and matrix spikes were analyzed. Acceptable levels for all the methods were: instrument calibration R 2 > 0.995, calibration check 10%, spike recoveries 20%, relative percent difference (rpd) 10%, and blanks MDL. 4.9 Objective 2 Data Analysis Using JMP 8.0 Statistical Software by SAS, analysis of variance (ANOVA) analyses were carried out to determine if significant differences exist between condition (oxidized and reduced), time (each sampling period) and the interaction between redox condition and time. Aqueous As species, HCl As and Fe species, acetate extractable As and Fe species, Eh, ph, and EC were all analyzed. Tukey's Honestly Significant

62 44 Difference (HSD) multiple comparison test (α = 0.05) was used to determine which results affected by which parameters were significantly different. A Student's t-test and 95% confidence intervals were also used where appropriate. In all figures, error bars are expressed as Tukey s HSD, unless otherwise stated. Significant differences were selected for p < 0.05.

63 45 CHAPTER 5 RESULTS 5.1 NP 9 and NP 13 Pore Water Data Arsenic concentrations in NP 9 pore water increased with depth from the soil surface through the redox transition zone, with As(V) dominating in surface, water table, and transition zones (Figure 7 (A)). Arsenic levels are highest in the water table zone, the area of seasonal fluctuating groundwater, with concentrations up to 355 µg/l total As. Arsenic (III) accounts for 31% of the total arsenic in the vadose zone, 1% in the redox transition zone, and 55% in the depletion zone. The groundwater can fluctuate up to 3 meters depending on the season and precipitation each year in the middle of the valley (Site Code , (USGS 2013)). The lowest As concentration is in the depletion zone, where As(III) is the dominant species. Fe(III) is the dominate oxidation state of iron in the NP 9 surface soil and Fe(II) dominates in the transition and depletion zone soils (Figure 7 (B)). In deeper sediments, reduced conditions were expected due to the redoximorphic features seen during sampling (Meng 2015). Interestingly, while Fe(II) dominates in the transition zone, the As is oxidized, as As(V). Similar to NP 9, the highest As concentrations in NP 13 were in the water table zone and contain both As(III) and As(V) species (maximum 231 µg/l total As) (Figure 7 (C)). The transition zone has primarily As(V) present but in the depletion zone, both oxidation states are present. Unlike NP 9, Fe(III) dominates in the transition zone,

64 46 without the presence of Fe(II). In addition, NP 13 had higher concentrations of Fe (<MDL mg/l) compared to NP 9 ( mg/l) (Figure 7 (D)). Interestingly, the majority of Fe in the depletion zone is Fe(III), where the most reduced sediments are supposed to occur.

65 Depth, cm Depth, cm As, ug/l Water Table As(III) As(V) A NP 9-1 NP 9-2 NP 9-3 NP 9-4 NP 9-5 NP 9-6 NP 9-7 NP 9-8 NP 9-9 NP 9-10 NP 9-11 NP 9-12 NP 9-13 NP 9-14 NP 9-15 NP 9-16 Fe, mg/l Fe(II) Fe(III) B As, µg/l Fe, mg/l Water Table NP 13-1 Np NP NP NP NP NP As(III) As(V) C NP 13-8 NP 13-9 NP NP NP NP NP Fe(II) Fe(III) D Figure 7: Pore Water NP 9As, μg/l (A), NP 9 Fe, mg/l (B) NP 13 As, µg/l (C) and NP 13 Fe mg/l (D). The shaded region indicates the transition zone.

66 NP 9 and NP 13 Solid Phase Extractions Fe oxides are often identified in the literature as the major source of As release (Anawar et al. 2003; Jung et al. 2012). Above the depletion zone in both cores, the largest pool for As is associated with very amorphous (step 3) and amorphous (step 4) Fe oxides (Figure 8). However, arsenic-carbonate minerals in the environment may also be a sink for As. For the NP 9 and NP 13 soil cores, approximately 7-18% of the As is associated with carbonate minerals in these upper zones (Figure 8). These results demonstrate the significant role carbonate minerals may play in explaining As mobility in these semi-arid basin fill aquifer solids. Work by Brannon and Patrick (1987) and Goldberg and Glaubig (1988) concludes that As sorption onto carbonate surfaces in soil is described by ligand exchange (maximum sorption at ph 10.5) (Goldberg and Glaubig 1988), the fraction of As extractable in the first extraction step. Similarly, in soils with ph<9, carbonate surfaces have a positive charge and therefore provide sites for As sorption in calcareous soils (Sadiq 1997). The greatest percentage of ligand exchangeable As and As associated with carbonate minerals is in the water table zone, Figure 8 (A) and (B). Down the profiles, the percent of As associated with less soluble minerals increase, in general. Arsenic is associated with sulfide minerals in the depletion zone. The amount of As(III) and As(V) (µg/kg) extracted in NP 9 and NP 13 at each step is shown in Figure 9 (A), (B), (C), and (D), respectively. Per the sequential extraction procedure, the extractants used for crystalline Fe oxides oxidized As(III) to As(V) (Huang and Kretzschmar 2010) are reducing agents thus the oxidation state of As associated with these minerals cannot be distinguished. For this reason, the As associated

67 49 with crystalline Fe oxides is not displayed on Figure 9. Throughout the cores, As(V) is the dominant species in the solid phase above the depletion zone (Figure 9 (B) and (D)). As(III) was near the reporting limit throughout most of the NP 13 profile. In NP 9, the majority of the As(III) is associated with carbonates or with very amorphous Fe oxides and As(V) with very amorphous Fe oxides and the more insoluble Fe oxides and sulfide minerals. In the depletion zone of both cores, the majority of As is associated with sulfide minerals. This is important because As-pyrite or orpiment minerals may be a source of As in solution (Chowdhury et al. 1999). Less than 10% of the Fe is exchangeable, associated with carbonates, or Mn oxides and very amorphous Fe-O minerals in both cores. The greatest amount of Fe is associated with sulfide and crystalline Fe-O minerals (Figure 10). Geochemical modeling was used in this study to determine if pyrite or crystalline Fe-O minerals were present, particularly in the depletion zone where pyrite precipitation is favored.

68 50 NP 9-1 NP 9-2 NP9-3 NP 9-4 NP 9-5 NP 9-6 NP 9-7 NP 9-8 NP 9-9 NP 9-10 NP 9-11 NP9-12 NP 9-13 NP 9-14 NP 9-15 NP % 10% 20% 30% 40% 50% 60% 70% 80% 90% 100% Ligand Exchangeable Mn oxides, very amorphous Fe oxides, AVS Sulfides Residual Carbonate Minerals Amorphous Fe oxyhydroxides Crystalline Fe-O A 0% 10% 20% 30% 40% 50% 60% 70% 80% 90% 100% NP 13-1 NP 13-2 NP 13-3 NP 13-4 NP 13-5 NP 13-6 NP 13-7 NP 13-8 NP 13-9 NP NP NP NP NP Ligand exchangeable Mn oxides, very amorphous Fe oxides Sulfides Residual Carbonate Minerals Amorphorus Fe oxides Crystalline Fe-O B Figure 8: Percent of As extracted at each extraction step down the NP 9(A) and NP 13 (B) Profiles

69 51 µg/kg As(III) µg/kg As(V) NP 9-1 NP 9-2 NP 9-3 NP 9-4 NP 9-5 NP 9-6 NP 9-7 NP 9-8 NP 9-9 NP 9-10 NP 9-11 NP 9-12 NP 9-13 NP 9-14 NP 9-15 NP 9-16 A NP 9-1 NP 9-2 NP 9-3 NP 9-4 NP 9-5 NP 9-6 NP 9-7 NP 9-8 NP 9-9 NP 9-10 NP 9-11 NP 9-12 NP 9-13 NP 9-14 NP 9-15 NP 9-16 Ligand exchangeable Carbonate Minerals Mn oxides, very amorphous Fe oxides, AVS Amorphorus Fe oxides Sulfides B NP 13-1 NP 13-2 NP 13-3 NP 13-4 NP 13-5 NP 13-6 NP 13-7 NP 13-8 NP 13-9 NP NP NP NP NP As(III) ug/kg C NP 13-1 NP 13-2 NP 13-3 NP 13-4 NP 13-5 NP 13-6 NP 13-7 NP 13-8 NP 13-9 NP NP NP NP NP As(V) ug/kg D Figure 9: The Concentration of As(III) (A) and As(V) (B) Extracted in NP 9, µg/kg and the Concentration As(III) (C) and As(V) (D) Extracted in NP 13, µg/kg

70 NP 9-1 NP 9-2 NP 9-3 NP 9-4 NP 9-5 NP 9-6 NP 9-7 NP 9-8 NP 9-9 NP 9-10 NP 9-11 NP 9-12 NP 9-13 NP 9-14 NP 9-15 NP 9-16 NP13-1 NP13-2 NP13-3 NP13-4 NP13-5 NP13-6 NP13-7 NP13-8 NP13-9 NP13-10 NP13-11 NP13-12 NP13-13 NP % 10% 20% 30% 40% 50% 60% 70% 80% 90% 100% Exchangeable Mn oxides, very amorphous Fe oxides, AVS Sulfides Residual Carbonate Minerals Amorphous Fe oxides Crystalline Fe-O 0% 20% 40% 60% 80% 100% 52 A Ligand exchangeable Mn oxides, very amorphous Fe oxides, AVS Sulfides Residual Carbonate Minerals Amorphorus Fe oxides Crystalline Fe oxides B Figure 10: Percent of Fe extracted at each extraction step down the NP 9 (A) and NP 13 (B) Profiles

71 NP 9 and NP 13 Geochemical Modeling Results - Solid Phases Controlling Arsenic Solution Chemistry Using PHREEQC, the predicted solid phases were generated for each layer based on pore water chemistry and ph. Eh was not specified in the model, but instead was calculated based on the other input values. The following As minerals were stored in the database and used while running the model: orpiment (As 2 S 3 ), realgar (AsS), arsenolite (As 2 O 3 ), claudetite (As 2 O 3 ), As 2 O 5, AsI 3, AlAsO 4 :2H 2 O, Cu 3 (AsO 4 ) 2 :2H 2 O, Zn 3 (AsO 4 ) 2 :2.5H 2 O, FeAsO 4 :2H 2 O, Mn 3 (AsO 4 ) 2 :8H 2 O, Ca 3 (AsO 4 ) 2 :4H 2 O, and Ba 3 (AsO 4 ) 2. Other Cu-As minerals, including enargite (Cu 3 AsS 4 ) and tennanite (Cu 12 As 4 S 13 ), were unable to be added to the PHREEQC database due to limited K value information and the inability to add minerals to the database. The Ksp of Ba 3 AsO 4 was adjusted to and used throughout the study in PHREEQC. If the default Ksp stored in the database was used, all As(V) would precipitate with Ba 3 AsO 4. All model inputs are in Appendix B, Table B. 1-Table B. 4. Throughout the soil profile, As(V) was not predicted to form any As(V) minerals in NP-9 or NP-13. Though not predicted to form in every individual layer, orpiment was the controlling As(III) solid phase in each zone in NP 9 and NP 13, with the exception of the vadose zone for NP 13. The solids formed in each layer throughout the NP 9 and NP 13 profile are in Appendix D: (Table D. 1-Table D. 8). Due to the highest concentrations of As in the pore water in the water table zone, it is important to understand solid phases controlling release and retention in this zone. Orpiment controlled the As(III) chemistry in NP 9-5 and NP 13-3 and 13-4, but not NP 9-

72 54 4 (Figure 11 (A) and (B)). It was determined that the limiting constituent in NP 9-4 for the formation of orpiment is HS -1. A concentration of 1E-7 M or 3.21E-3 mg L -1 HS -1 is required for the precipitation of orpiment. This was determined through HS -1 titrations using PHREEQC. The HS -1 in the system precipitates as pyrite and Cu-S minerals prior to orpiment formation. This was observed by varying the HS -1 concentration and observing the changes in precipitates formed. Similar to the vadose zone, several Fe and Cu solids also form in these layers. Calcite and various Fe and Mn oxide minerals were predicted in various layers down both profiles. Orpiment is the solids phase predicted to control As(III) solubility throughout the rest of the NP-13 profile (Appendix D). This is not the case for NP-9 where orpiment only forms in the layer right below the water table zone, NP 9-6 and in layers 12 and 13 in the depletion zone. Other mineral predicted to form down the profiles are calcite, dolomite and various Fe and Mn oxides. The sequential extraction data demonstrated significant amounts of As(V) associated with ligand exchangeable minerals, very amorphous Fe oxides, manganese oxides, and amorphous Fe oxides. Further modeling was performed to include sorption onto FeOH and CaCO 3 to determine the location of As(V).

73 55 Anilite Aragonite Barite Blaublei I Blaublei II Calcite Chalcocite Chalcopyrite Covellite Cupricferrite Cuprousferrite Diaspore Djurleite Dolomite(disordered) Fe(OH)2.7Cl.3 Fe3(OH)8 Ferrihydrite Gibbsite Goethite Greigite Hematite Hercynite Hydroxylapatite Lepidocrocite Maghemite Magnesioferrite Magnetite MnHPO4 MnS(grn) Orpiment Pyrite Rhodochrosite Sphalerite Sulfur Dolomite(ordered) NP 9-4 NP 9-5 A Anilite Aragonite Barite Blaublei I Blaublei II Boehmite Calcite Chalcopyrite Covellite Cupricferrite Cuprousferrite Diaspore Dolomite(disordered) Fe(OH)2.7Cl.3 Fe3(OH)8 Ferrihydrite Gibbsite Goethite Greigite Hematite Hercynite Huntite Hydroxylapatite K-Jarosite Lepidocrocite Maghemite Magnesioferrite Magnesite Magnetite Orpiment Pyrite Sphalerite Sulfur Wurtzite ZnS(am) Dolomite(ordered) NP 13-3 NP 13-4 B Figure 11: Predicted Solids in NP 9-4 and NP 9-5 (A) and Predicted Solids in NP 13-3 and NP 13-4 (B)

74 Sorption onto FeOH and CaCO 3 Using each soil layer's pore water concentrations, Fe(III) associated with very amorphous and amorphous Fe oxides minerals, and the Sø et al. (2008) database stored in MINTEQ, sorption onto HFO and CaCO3 was modeled simultaneously. The predicted As(III) distribution throughout the NP 9 soil profile was as FeHAsO3-. (over 70%) with the remaining fraction as FeH2AsO3 (Table 8). As(III) is not significantly sorbed by high affinity sites on HFO. Sø et al. (2008) reported As(III) was not significantly sorbed to calcite and therefore did not report log K values. The database within MINTEQ for As sorption to calcite was taken form Sø et al. (2008) and therefore does not have log K values for As(III) sorption to calcite.

75 57 Table 9 shows the predicted As(V) distribution and sorption onto HFO and calcite down the NP-9 profile. Modeling results indicated approximately 99.4% of As(V) sorbs onto hydrous ferric oxides (HFO) and <1% sorb onto calcite throughout the NP 9 profile, with the exception of NP 9-3 and 9-4. These layers are in the vadose and water table zones, the layers with the highest As concentration in the pore water (Table 9). It is important to note that the Fe(III) solids concentration input into the model includes several underlying assumptions. The Fe(III) extracted from the very amorphous Fe oxides may also include Fe extracted from acid volatile sulfides and Mn-O minerals. In this model, it is assumed that 100% of the Fe(III) extracted at that step is HFO. In addition, the Fe extracted from amorphous Fe oxides minerals is assumed to be entirely Fe(III). The sum of these two extractions provide a "best case scenario" for As(V) removal through sorption processes. Similar to NP 9, the model predicted As(III) was sorbed throughout the NP 13 profile as FeHAsO - 3 (80-99%) (

76 58 Table 10). Unlike NP-9, a significant percentage of As(V) was sorbed to calcite (Table 11). An average of 9% of the As(V) sorbed onto calcite. In layers NP 13-1 and 13-2, 59% and 19%, respectively, of the As(V) sorbed onto calcite. These layers are from the surface soils. Table 8: NP 9 As(III) Distribution and Sorption onto HFO as Predicted by MINTEQ (small h denotes high affinity sites on HFO) H3AsO3 >FehHAsO3(-) >FeHAsO3(-) >FehH2AsO3 >FeH2AsO3 NP % 0.20% 71.7% 0.08% 28.0% NP % 0.41% 82.8% 0.08% 16.7% NP % 0.60% 84.4% 0.11% 14.9% NP % 0.87% 84.0% 0.16% 14.9% NP % 1.71% 79.5% 0.40% 18.4% NP % 1.47% 80.2% 0.33% 17.9% NP % 1.33% 80.6% 0.29% 17.7% NP % 1.26% 79.0% 0.31% 19.4% NP % 1.34% 78.8% 0.33% 19.5% NP % 1.40% 78.6% 0.35% 19.7% NP % 1.12% 76.9% 0.32% 21.6% NP % 0.98% 81.9% 0.20% 16.9% NP % 0.94% 75.6% 0.29% 23.1% NP % 0.74% 73.1% 0.26% 25.8% NP % 1.17% 79.1% 0.29% 19.4% NP % 0.79% 83.3% 0.15% 15.8%

77 Table 9: NP 9 As(V) Distribution and Sorption onto HFO and CaCO3 as Predicted by MINTEQ (small h denotes high affinity sites on HFO) (=SO)2A so2ca >FeHAsO 4(-) >FehAsO4( -2) >FeAsO4( -2) >FehOHAsO4 (-3) 59 >FeOHAsO4 (-3) NP % 0.99% 0.09% 33.5% 0.18% 65.2% NP % 0.32% 0.10% 20.7% 0.39% 78.0% NP % 0.25% 0.13% 18.4% 0.56% 79.0% NP % 0.25% 0.19% 18.1% 0.80% 77.4% NP % 0.40% 0.49% 22.6% 1.60% 74.2% NP % 0.38% 0.41% 22.2% 1.39% 75.5% NP % 0.37% 0.36% 21.9% 1.25% 75.6% NP % 0.44% 0.38% 23.8% 1.18% 73.5% NP % 0.45% 0.41% 24.0% 1.25% 73.7% NP % 0.46% 0.43% 24.2% 1.31% 73.5% NP % 0.56% 0.39% 26.4% 1.04% 71.4% NP % 0.33% 0.25% 20.9% 0.91% 76.8% NP % 0.65% 0.35% 28.1% 0.87% 69.7% NP % 0.83% 0.32% 31.1% 0.68% 66.8% NP % 0.45% 0.35% 23.9% 1.10% 74.1% NP % 0.28% 0.19% 19.7% 0.75% 78.7%

78 60 Table 10: NP 13 As(III) Distribution and Sorption onto HFO as Predicted by MINTEQ (small h denotes high affinity sites on HFO) H3AsO3 >FehHAsO3(-) >FeHAsO3(-) >FehH2AsO3 >FeH2AsO3 NP % 0.15% 99.4% 0.00% 0.5% NP % 0.51% 88.9% 0.06% 10.5% NP % 1.04% 91.1% 0.09% 7.7% NP % 1.14% 86.4% 0.16% 12.0% NP % 1.20% 86.6% 0.17% 12.1% NP % 1.16% 86.2% 0.17% 12.5% NP % 1.47% 87.3% 0.19% 11.0% NP % 1.29% 84.2% 0.22% 14.2% NP % 0.67% 83.6% 0.13% 15.6% NP % 1.20% 84.2% 0.20% 14.4% NP % 1.43% 80.4% 0.32% 17.9% NP % 1.21% 82.9% 0.23% 15.6% NP % 1.45% 81.8% 0.29% 16.5% NP % 1.54% 80.96% 0.33% 17.2%

79 61 Table 11: NP 13 As(V) Distribution and Sorption onto HFO and CaCO 3 as Predicted by MINTEQ (small h denotes high affinity sites on HFO) (=SO)2AsO 2Ca (1) >FeHAsO 4(-) >FehAsO4 (-2) >FeAsO4 (-2) >FehOHAsO4 (-3) >FeOHAsO4 (-3) NP % 0.00% 0.00% 0.2% 0.06% 41.2% NP % 0.10% 0.06% 10.9% 0.40% 69.6% NP % 0.06% 0.10% 9.1% 0.93% 81.8% NP % 0.15% 0.19% 14.4% 1.04% 78.3% NP % 0.15% 0.20% 14.6% 1.10% 79.5% NP % 0.17% 0.21% 15.3% 1.08% 80.2% NP % 0.13% 0.22% 13.2% 1.34% 79.4% NP % 0.21% 0.25% 16.6% 1.14% 74.5% NP % 0.25% 0.14% 17.7% 0.58% 72.1% NP % 0.23% 0.25% 17.8% 1.12% 78.9% NP % 0.37% 0.39% 22.1% 1.34% 75.4% NP % 0.28% 0.28% 19.3% 1.14% 77.9% NP % 0.31% 0.36% 20.4% 1.36% 76.7% NP % 0.34% 0.41% 21.3% 1.45% 76.2% 5.5 NP 9 and NP 13 Modeled Eh To understand how redox conditions change throughout the soil profile and since it is likely that redox conditions play a significant role in As release, PHREEQC was used to model the reduction potential (pe) (which was converted to Eh, in volts) down the soil profile using Fe(III)/Fe(II), As(V)/As(III) and sulfate/sulfide redox pairs. The Eh was predicted simultaneously with precipitate prediction. For layers with concentrations below the MDL for Fe(II) (12 µg/l), the MDL was used as the input. However, with the input of the MDL, PHREEQC models a higher proportion of reduced Fe than is actually present in the pore water. The Eh, down the NP 9 and NP 13 soil profile, is shown in Figure 12 (A,B). Depending on which redox pair was used, the Eh varies greatly down the two profiles. The reduction of Fe occurs in the Eh range of 300 to 90 mv and As

80 62 reduction is in the range of 24 to -150 mv (Mukherjee et al. 2011). Fe and As reduction is observed down the profile based on HCl extractable redox species (Figure 13, Figure 14). Only the depletion zone showed gleying-indicating reduced conditions. This is also the only zone with detectable acid volatile sulfide (AVS) and higher % of Fe(II). The model prediction for Eh was based on solution phase redox species. However the main redox potential is in the solid phase. Observing the ratio of As(III) to total As and Fe(II) to total Fe in the 0.5M HCl extract and other collected data is a check with unmodeled data (Figure 13 and Figure 14). There is a zone under Fe /As reduction at 100 cm and at depths below 375 cm (NP 9) and 100 cm 200 cm then below 350 cm for NP 13. AVS was only detected in solids in the depletion zone. There are several zones under Fe reduction, but only the depletion zone under sulfide reduction.

81 Depth, cm Depth, cm 63 Eh, V Eh, V As Pair S Pair Fe Pair A 500 As Pair S Pair Fe Pair B Figure 12: Modeled Eh Calculated Using As, N, Fe, and S Redox Pairs Down the NP 9 (A) and NP 13 (B) Profiles

82 Depth, cm Depth, cm 64 As(III)/As 0% 50% 100% Fe(II)/Fe 0% 50% 100% A B Figure 13: Measured Redox down the NP 9 Profile As(III)/As 0% 50% 100% Fe(II)/Fe 0% 50% 100% A B Figure 14: Measured Redox down the NP 13 Profile

83 Redox Experiment Results -As Solubilization and Fe Reduction In both aquifer solids, As(III) was released to solution by Day 7 (NP 9) and Day 15 (NP 13) at concentration that exceeded the MCL. Desorption or dissolution of As(V) does not contribute to As in solution in either sediment. The extent of Fe reduction was monitored in the solution phase and Fe(II) associated with HCl soluble minerals. There was no discernible Fe reduction until Day 63 in either sediment. Arsenic solubilization was independent of Fe reduction, contrary to many studies reported in the literature that emphasize reductive dissolution of Fe(III) minerals with release of the associated As as the major mechanism for As contamination of aquifers. Several studies that have reported decoupling of Fe and As reactivity have only measured Fe(II) in solution, but as displayed in this study the reduced Fe may remain associated with the solid phase. Without analysis of the solid phase the rate of extent of Fe reduction would have been greatly underestimated as also pointed out in the study by Weber et al. (2009). In this study, however, there was no detectable Fe(II) in either the solution or HCl extractable solid phase until Day 63. As(III) in solution may be due to desorption from or dissolutions of As(III) minerals. The concentration of As(III) associated with ligand exchange sites, as define by sequential extractions, was 139 µg/kg and 1 µg/kg for NP 9 and NP 13 respectively. By Day 7 and Day 15, the time period when As(III) was first measured in solution, the amount of As(III) in solution (on a mass bases (240 µg/kg for NP 9 and 390 µg/kg NP 13)) exceeded the exchangeable concentration. Over time, the maximum As(III) was 912 µg/kg for NP 9 and 624 µg/kg for NP 13 which was more As(III) than extracted with the

84 66 first three extraction steps (853 µg/kg for NP 9 and 374 µg/kg for NP 13). The As(III) in solution is a result of As(V) reduction (Figure 15). In NP 9, both As(III) and Fe(II) decreased with time. This may indicate precipitation of FeAsS, an association not extractable with HCl. This could also be attributed to problems with maintaining a low Eh (Figure F. 1). However, in NP 13 the Eh remained relatively low and As(III) and Fe(II) reached steady state conditions.

85 Aqueous As, µg/l HCl Fe, mg/kg Aqueous As, µg/l HCl Fe, mg/kg As(III) mg/l As(V) mg/l Fe(II) mg/kg A Days As(III) ug/l As(V) ug/l Fe(II) mg/kg B Days Figure 15: HCl Extractable Fe(II) and Aqueous As(III) and As(V) in NP 9-5 (A) and NP 13-3 (B) Reduced Samples. Red circles indicate point at which reduction occurred. Error bars represent ± Tukey's HSD.

86 68 Arsenic release was not unique to the reduced samples. As release occurred in the oxidized samples as As(V), and reached a maximum of 99 ± 22.1 µg/l on Day 63 in NP 9-5 (Figure 16 (A)). Similarly in NP 13-3, all the As was As(V) and reached a maximum of 71±23.1 µg/l on Day 63 (Figure 16 (B). These significant releases of As(V) imply that even under oxidized conditions, a significant amount of As(V) goes into solution. This may be attributed to the dissolution of As(V) minerals or the oxidation of As(III) to As(V), as discussed in the section 5.7 Arsenic Associated with the Solid Phase. It should be noted that though As released occurred under both conditions, larger concentrations of total As were released in the reduced samples. HCl extractable Fe(II) was present in both NP 9-5 and NP 13-3 oxidized samples, but the concentrations decreased from Day 0 to Day 7 then remained constant over the remainder of the study (Figure 16 (A) and (B)).

87 Aqueous As, µg/l HCl Fe, mg/kg Aqueous As, µg/l HCl Fe, mg/kg Days A As(V) ug/l 2500 As(III) ug/l 2000 Fe(II) mg/kg As(III) ug/l As(V) ug/l Fe(II) mg/kg Days B Figure 16: HCl Extractable Fe(II) and Aqueous As(III) and As(V) in NP 9-5 (A) and NP 13-3 (B) Oxidized Samples. Error bars represent ± Tukey's HSD.

88 70 In NP 9-5 reduced samples, HCl extractable Fe(III) was below the MDL, but was present in NP 13-3 reduced samples (Appendix F (Figure F. 4 and Figure F. 5). In NP 9-5 oxidized samples, from Day 63 to 113, HCl extractable Fe(III) was present at concentrations ranging from 122 ± 79 to 94 ± 23 mg/kg. The Fe(III) in the 13-3 oxidized samples, was 100 mg/kg greater than the NP 13-3 reduced samples on Days 63 and 113. The release of Fe(III) in NP 9-5 and NP 13-3 oxidized samples after Day 63 may be the result of Fe(II) oxidation or the dissolution of Fe(III) HCl extractable minerals (Appendix F). In addition, it is important to note that the lack of HCl extractable Fe(III) minerals on Day 0 (Table 4) in NP 9-5 and NP These results are supported by Meng s (2015) seven step sequential extractions (Table 7). Meng (2015) also found the amorphous Fe(III) was not extractable by 0.5 M HCl, but instead was extracted in 0.2 M NH oxalate buffer (ph 3.25) + 1 mm HgCl 2 (Huang and Kretzschmar 2010). As expected, the fraction of Fe(III) extracted by HCl in NP 13-3 demonstrates that only a portion of the amorphous Fe(III) is HCl extractable. 5.7 Arsenic Associated with the Solid Phase Contrary to literature that describes the release of As and Fe simultaneously through reductive dissolution, results from NP 9-5 and NP 13-3 indicated these reactions, in fact, do not co-occur. To further determine what may control As release and retention, the amount of As and Fe associated with carbonate minerals was determined through a 1M ammonium acetate extraction at ph 5. This extraction solution was selected due to the high percentage of CaCO 3 in both NP 9-5 and NP 13-3 (32.8% and 29.7%, respectively) and modeling results that indicated carbonate minerals account for some of

89 71 the sorb As(V) in these soils. The acetate buffer is specifically designed to determine the concentrations and speciation of As and Fe associated with carbonate soluble minerals (Amacher 1988). Fe(II) extracted from carbonate minerals in NP 9-5 reduced sediments went from 38 mg/kg on Day 0 to 116 mg/kg on Day 113 (Figure 17(A)). Similarly, the Fe(II) associated with carbonate minerals in NP 13-3 increased 64 mg/kg over 113 days (Figure 18 (A)). The increase in Fe(II) can be described by siderite (FeCO 3 ) formation, which is known to both sorb and co-precipitate with As (Bardelli et al. 2011; Sø et al. 2008). The Fe(II), however, from the oxidized samples, did not significantly change throughout the experiment. An average 17 mg/kg and 13 mg/kg Fe(II) was extracted from NP 9-5 and 13-3 oxidized samples, respectively (Figure 17 (A) and Figure 18 (A)). The release of As(III) from the ammonium acetate extractions in the NP 9-5 and NP 13-3 reduced samples was significantly greater than the oxidized samples by Day 15 and Day 7 (Figure 17 (B), Figure 18 (B)), respectively. In both reduced soils, the amount of As(III) increased and As(V) decreased. NP 9-5 reduced samples began with a 2%:98% ratio of As(III) to As(V), but by Day 28, the ratio changed to 13%:87%, respectively. Similarly, NP 13-3 reduced samples began with a 4%:96% ratio of As(III) to As(V), but by Day 7, the ratio changed to 21%:79%, respectively. There was an initial release of As(III) in both sets of oxidized samples, but by Day 15 and Day 28, all the As was As(V) in NP 9-5 and NP 13-3, respectively (Figure 17(A) and Figure 18 (A)). There were significant differences between oxidized and reduced As(V) concentrations in NP 13-3 extractable As(V), but not NP 9-5.

90 As(III), µg/kg As(V), µg/kg Fe(II), mg/kg Fe(II) Oxidized Fe(II) Reduced A 400 As(III) Oxidized As(V) Oxidized As(III) Reduced As(V) Reduced Days B Figure 17: Acetate Extractable Fe(II) (mg/kg) (A) and As (III) and As(V) (µg/kg) (B) in NP 9-5 Under Oxidized (solid line) and Reduced (dashed line) Conditions. Error bars represent ± Tukey's HSD.

91 As(III), µg/kg As(V), µg/kg Fe(II), mg/kg Fe(II) Oxidized Fe(II) Reduced A As(III) Oxidized As(V) Oxidized As(III) Reduced As(V) Reduced Days B Figure 18: Acetate Extractable Fe(II) (mg/kg) (A) and As (III) and As(V) (µg/kg) (B) in NP 13-3 Under Oxidized (solid line) and Reduced (dashed line) Conditions. Error bars represent ± Tukey's HSD. The decrease in acetate extractable As(V) in the reduced samples can be explained when analyzing both the acetate extractable As(III) and aqueous phase As(III). The gain and loss of aqueous and acetate extractable As, relative to time 0, was analyzed over time for NP 9-5 and NP 13-3 under reducing conditions (Figure 19 and Figure 20). The significant As(III) and As(V) gain and loss over 133 days was determined by a Student's t-test between Day 0 and each subsequent time period. The lack of balance

92 74 between gains and losses may be attributed to only analyzing aqueous and acetate extractable As. As(III) and As(V) may have been redistributed onto mineral surface not extractable with acetate. In NP 9-5, the As, all as As(V), was removed from mineral phases soluble in 1 M acetate buffer (ph 5). The dissolution of carbonate minerals with the acetate buffer would remove As associated with the surface of carbonate minerals and As incorporated into the structure of carbonate minerals. Acetate is not an effective ligand for removal of As from ligand exchange sites; As(V) surface bound to non-acetate soluble minerals, such as Fe oxides, would not be removed by this extractant. These observations defined the major portion of As that is subject to solubilization in these sediments as carbonate-associated As. As(V) was reduced and the produced As(III) was associated with the aqueous phase and also associated with carbonate minerals, or other minerals solubilized by acetate buffer at ph 5. Sø et al. (2008) did not observed sorption of As(III) to calcite in their batch sorption studies. Roman-Ross et al. (2006), however, observed As(III) associated with calcite surfaces. Recent findings suggest that arsenite is trapped in the crystalline lattice of calcite (Bardelli et al. 2011; Castagliola et al. 2007). The lack of mass balance between the loss of As(V) and gain of As(III) for most time intervals was due to either As(V) or As(III) (from the reduction of As(V)) sorbing onto or being incorporated into non-acetate extractable minerals. In order to track the fate of As(V) and As(III) across all potential mineral phases, a full sequential extraction would be needed, as described by Meng (2015).

93 75 In NP 9-5, there was a continuous gain of aqueous As(III) from Day 7 to 77 (Figure 19). In addition there was a loss of acetate extractable As(V) and gain of As(III) from Day 28 to 113. The loss of As(V) associated with carbonates (Day 28-Day 113) was most likely due to desorption from carbonate surfaces. Reduced to As(III) in solution, the As then both resorbed onto carbonate surfaces and remained in solution, as seen with the gain in As(III) until Day 113. It is important to note that the As associated with carbonates refers to As associated with carbonate minerals, ligand exchangeable, and unknown minerals. As(III) incorporated in the calcite lattice structure and As(V) loosely bound to carbonate surfaces is supported by Sø et al. (2008), who found the rate of As(V) sorption and desorption from CaCO 3 to be relatively quick, indicating the As(V) is not incorporated in the calcite lattice structure. In NP 13-3 reduced samples, there is a continuous gain of As(III) associated with carbonate minerals throughout the experiment (Figure 20). There was an initial gain of aqueous As(V) on Day 2, but that is assumed to be from initial dissolution or desorption of As(V). With the exception of Day 15 and 28, there was a loss of As(V) associated with carbonate minerals from Day 7 to 113, similar to NP 9-5. As(III) is continuously gained and As(V) lost between Days 63 and 113. Similar to NP 9-5 reduced samples, the loss of As(V) associated with carbonates was most likely due to desorption from carbonate surfaces. Reduced in solution, the As(III) resorbed onto carbonate surfaces and/or remained in solution.

94 76 Figure 19: NP 9-5 Distribution of As in the Carbonate Mineral and Aqueous Phases Over 113 Days Under Reduced Conditions Figure 20: NP 13-3 Distribution of As in the Carbonate Mineral and Aqueous Phases Over 113 Days Under Reduced Conditions