UBC Theses and Dissertations

UBC Theses Logo

UBC Theses and Dissertations

Application of molybdenum (and zinc) stable isotopes to trace geochemical attenuation in mine waste Skierszkan, Elliott Karl 2018

Your browser doesn't seem to have a PDF viewer, please download the PDF to view this item.

Item Metadata

Download

Media
24-ubc_2018_september_skierszkan_elliott.pdf [ 11.71MB ]
Metadata
JSON: 24-1.0368919.json
JSON-LD: 24-1.0368919-ld.json
RDF/XML (Pretty): 24-1.0368919-rdf.xml
RDF/JSON: 24-1.0368919-rdf.json
Turtle: 24-1.0368919-turtle.txt
N-Triples: 24-1.0368919-rdf-ntriples.txt
Original Record: 24-1.0368919-source.json
Full Text
24-1.0368919-fulltext.txt
Citation
24-1.0368919.ris

Full Text

   APPLICATION OF MOLYBDENUM (AND ZINC) STABLE ISOTOPES TO TRACE GEOCHEMICAL ATTENUATION IN MINE WASTE  by Elliott Karl Skierszkan B.Sc., University of Ottawa, 2012    A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF  THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY in THE FACULTY OF GRADUATE AND POSTDOCTORAL STUDIES (Geological Sciences)     THE UNIVERSITY OF BRITISH COLUMBIA (Vancouver) July 2018    © Elliott Karl Skierszkan, 2018   ii   The following individuals certify that they have read, and recommend to the Faculty of Graduate and Postdoctoral Studies for acceptance, the dissertation entitled Application of Molybdenum (and Zinc) Stable Isotopes to Trace Geochemical Attenuation in Mine Wastes, submitted by   Elliott Skierszkan in partial fulfillment of the requirements for the degree of   Doctor of Philosophy in    Geological Sciences. Examining Committee: Prof. K. Ulrich Mayer, Geological Sciences Co-supervisor Prof. Roger Beckie, Geological Sciences Co-supervisor Prof. Dominique Weis, Geological Sciences Supervisory Committee Member Prof. Susan Baldwin, Chemical and Biological Engineering University Examiner  Prof. Lee Groat, Geological Sciences University (Departmental) Examiner  Prof. Dominik Weiss External Examiner   Additional Supervisory Committee Members Dr. Marghaleray Amini, Geological Sciences, Supervisory Committee Member    iii   Abstract Mining activities generate tremendous quantities of waste rock and tailings that must be carefully managed to prevent contamination of water resources by metal-leaching. Proper environmental management of mine drainage requires a detailed understanding of the mechanisms that control the mobility of metals in mine waste. This thesis applied stable-isotope analyses of molybdenum (Mo) and zinc (Zn) to constrain the geochemical attenuation processes controlling transport of these metals in mine waste. The key outcomes of this work are: (1) the establishment of a robust protocol for determining high-precision Mo isotope ratios in mine-waste samples using double-spike multi-collector inductively coupled plasma mass-spectrometry; (2) the demonstration that mine drainage at field sites becomes enriched in heavy Mo isotopes because Mo attenuation preferentially removes light isotopes; the predominant Mo attenuation mechanisms considered being sorption onto (oxyhydr)oxides and precipitation of molybdate minerals; (3) the demonstration that mine drainage becomes depleted in heavy Zn isotopes under alkaline pH conditions because of preferential removal of heavy Zn isotopes during Zn adsorption and/or precipitation of secondary minerals; and (4) the determination of new Mo isotopic fractionation factors for the precipitation of powellite (CaMoO4) and wulfenite (PbMoO4)—important sinks of Mo in mine waste environments. Overall, this thesis demonstrates that metal stable isotope analyses are an informative new tool now available to trace the processes that control metal transport in the environment. Further improvements in the   iv   quantification of metal removal using stable isotopic analyses should become possible with ongoing research into the causes of metal stable isotope fractionation.     v   Lay Summary Mining generates tremendous quantities of waste rock and tailings that must be carefully managed to prevent contamination of water resources by potentially harmful trace metals. Proper environmental management of mine waste requires a detailed understanding of the processes that control the mobility of trace metals. This thesis applied a new technique of chemical analysis—metal stable-isotope measurements—to track the processes that control the leaching of metals from mine waste. Most trace metals comprise multiple stable isotopes, which have subtle variations in mass. The heavy and light isotopes of a given chemical element are distributed unevenly during chemical reactions, including those reactions that can remove metals from water. Metal removal from mine drainage causes changes in the stable isotopic composition of metals in water; isotopic analyses therefore constitute a new tool available to indicate removal of metals from contaminated environments such as mine waste facilities.     vi   Preface This thesis comprises an introduction, a conclusion, and five research chapters. The latter were prepared in manuscript format for publication in international geoscience journals. I am the lead author for all manuscripts. A list of contributions from each of my co-authors and colleagues is provided below for each of these manuscripts.  Chapter 2 A version of Chapter 2 has been published as: Skierszkan E. K., Amini M. and Weis D. (2015) A practical guide for the design and implementation of the double-spike technique for precise determination of molybdenum isotope compositions of environmental samples. Anal. Bioanal. Chem. 407, 1925–1935. I conducted the majority of ion-exchange chemistry and geochemical analyses during the development of the analytical protocol described in this chapter with assistance from Dr. Marghaleray Amini. Kathy Gordon also assisted with operation of the Nu Plasma MC-ICP-MS used in this study. Dr. Amini and Vivian Lai assisted with operation of the Agilent 7700x ICP-MS used to measure sample concentrations. Drs. Amini and Dominique Weis provided guidance regarding the interpretation of the data and editorial support during the writing of the manuscript. Discussions with Drs. Uli Mayer, Roger Beckie, Nicolas Estrade, Jan Fietzke, and M. Choi improved the writing of the manuscript. It was reviewed by Dr. Klaus Heumann and two   vii   anonymous reviewers prior to publication. Reference materials and isotope standards used to confirm the accuracy of the Mo isotope analyses were provided courtesy of Maureen Soon and Drs. Christopher Siebert, Mark Rehkämper, and Jane Barling.  Chapter 3 A version of Chapter 3 has been published as: Skierszkan E.K., Mayer K.U., Weis D. and Beckie R.D. (2016) Molybdenum and zinc stable isotope variation in mining waste rock drainage and waste rock at the Antamina mine, Peru. Sci. Total Environ. 550, 103–113. Mine drainage and waste rock samples were collected at the Antamina mine by UBC graduate students Laura Laurenzi, Mélanie St-Arnault, and María Lorca-Ugalde, and Dr. Roger Beckie with support from Antamina mine staff. Ore samples were provided by Bevin Harrison. I conducted ICP-MS and MC-ICP-MS analyses of samples, with analytical assistance from Kathy Gordon (MC-ICP-MS), Vivian Lai (ICP-MS), and Dr. Amini (MC-ICP-MS and ICP-MS). I also conducted the crushing and digestion of solid-phase samples analyzed for Mo and Zn isotopes in this study. Solid-phase mineralogy by XRD was compiled from the theses of J. Dockrey (M.Sc) and Dr. H. Peterson (PhD). I was responsible for interpretation of the data and the writing of the manuscript, with intellectual input from Drs. Beckie, Mayer, and Weis. Dr. Ryan Mathur and two anonymous peer reviewers also suggested improvements to the manuscript during the publication process.   viii   Chapter 4  A version of Chapter 4 has been published as: Skierszkan, Elliott K., Stockwell J.S., Dockrey J.W., Weis D., Beckie R.D. and Mayer K.U. (2017) Molybdenum (Mo) stable isotopic variations as indicators of Mo attenuation in mine waste-rock drainage. Appl. Geochemistry 87, 71–83. I was responsible for the collection of all samples. I conducted most laboratory analyses, with the exception of anions analyses in water samples, which were conducted by ALS Laboratories in Burnaby, Canada. Rock samples were selected with support from Thompson Creek Mine staff, including mine geologist Ray Cheff and environmental technician Richard Giampedraglia. Analytical support was provided by Dr. Elisabetta Pani, Lan Kato, and Jenny Lai for XRD; Maureen Soon for ICP-OES; Vivian Lai and Dr. Amini for ICP-MS; and Kathy Gordon and Dr. Amini for MC-ICP-MS. The Raman spectrum sample PH Toe-S1 was collected by Dr. Soumya Das at the University of Saskatchewan and interpreted by me with Dr. Das’ assistance. Justin Stockwell, Laura Laurenzi, Dr. Gregor Lucic, Chris Natoli, Bert Doughty, and Richard Giampedraglia assisted with the collection of samples in the field, which I oversaw. I was responsible for writing the manuscript and interpretating the data, with intellectual input from my co-authors Justin Stockwell, John Dockrey, and Drs. Beckie, Mayer, and Weis. An anonymous peer-reviewer and Dr. Michael Kersten provided input prior to publication.     ix   Chapter 5 A version of Chapter 5 has been submitted to an international geochemistry journal. I was responsible for the design of experiments, collection and analysis of samples, interpretation of data, and writing of this manuscript. My co-authors Drs. Beckie, Mayer, and Weis provided guidance throughout this project. Dr. Laura Bilenker provided comments on the manuscript. Kathy Gordon, Maureen Soon, Dr. Amini, Elisabetta Pani, Lan Kato, and Jenny Lai provided support with analytical instrumentation, which comprised ICP-OES, MC-ICP-MS, and XRD. Chapter 6 A version of Chapter 6 is in preparation for submission to an international geochemistry journal. I was responsible for collection of samples, with field support from Justin Stockwell, Laura Laurenzi, Dr. Gregor Lucic, Chris Natoli, Bert Doughty, and Richard Giampedraglia. I conducted the majority of laboratory analyses used in this study. Dr. Marg Amini and Vivian Lai assisted with ICP-MS analyses, Maureen Soon assisted with ICP-OES and Elemental Analysis (Total Organic/Inorganic Carbon, Total Sulfur), Dr. Amini, and Kathy Gordon assisted with MC-ICP-MS analyses. Anions analyses were conducted by ALS Laboratories in Burnaby, BC. I conducted X-ray diffraction analyses with support from Dr. Elisabetta Pani, Jenni Lai, and Lan Kato. Dr. Soumya Das helped collect Raman spectra of sludge samples and provided support in the interpretation of these analyses. Transmission-electron microscopy was conducted with assistance from Dr. Jared Robertson. Dr. Matt Lindsay collected X-ray absorption spectra (XAS)   x   of samples at the Canadian Light Source in Saskatoon, Canada and Dr. Robertson conducted the modelling of the XAS results with input from me. Data interpretation was guided by conversations with my co-authors Justin Stockwell and John Dockrey and Drs. Jared Robertson, Uli Mayer, Roger Beckie, Dominique Weis, Soumya Das, and Matt Lindsay. Financial support for the research conducted in this dissertation was provided through: NSERC Alexander Graham Bell Canada Graduate Scholarships that I received (CGS-M, CGS-D); the Multidisciplinary Applied Geochemistry Network (MAGNET) grant, which is part of the NSERC Collaborative Research and Training Experience (CREATE) program; the Compania Miñera Antamina; Lorax Environmental Services Ltd.; a Graduate Scholarship from the Canadian Institute for Mining, Metallurgy and Petroleum; the W.H. Mathews Memorial Scholarship; and a UBC Four-Year Fellowship. Additional equipment and/or analytical contributions came from the Thompson Creek Mine and Dr. Matthew Lindsay’s Environmental Geochemistry Laboratory at the University of Saskatchewan.    xi   Table of Contents Abstract .......................................................................................................................................... iii Lay Summary ...................................................................................................................................v Preface............................................................................................................................................ vi Table of Contents ........................................................................................................................... xi List of Tables ...............................................................................................................................xxv List of Figures ............................................................................................................................ xxxi List of Abbreviations and Symbols................................................................................................ xl Acknowledgements ...................................................................................................................... xlii Dedication ................................................................................................................................... xliv CHAPTER 1. Introduction .......................................................................................................... 1 1.1 Environmental challenges posed by mining waste ..................................................... 1 1.2 Analytical techniques available to study metal attenuation ........................................ 2 1.3 Mechanisms of stable-isotope fractionation ............................................................... 9   xii   1.4 Studies of metal stable-isotope fractionation applied to mine waste ......................................................................................................................... 11 1.5 Molybdenum elemental and isotopic geochemistry ................................................. 13 1.6 Zinc elemental and isotopic geochemistry ................................................................ 18 1.7 Analytical considerations for metal stable-isotope analyses by MC-ICP-MS ......................................................................................................... 22 1.8 Thesis objectives and organization ........................................................................... 24 1.8.1 Thesis objectives ................................................................................................. 24 1.8.2 Selection of field sites ......................................................................................... 25 1.8.3 Thesis organization ............................................................................................. 26 CHAPTER 2. A practical guide for the design and implementation of the double-spike technique for precise determination of molybdenum isotope compositions of environmental samples .................................................................. 31 2.1 Introduction ............................................................................................................... 31 2.2 Methods .................................................................................................................... 36   xiii   2.2.1 Considerations for the double spike design ........................................................ 36 2.2.2 Double spike and standard preparation and calibration ...................................... 37 2.2.3 Validation of the double spike method’s accuracy ............................................. 43 2.3 Results and Discussion ............................................................................................. 46 2.3.1 Verification of the accuracy of the double spike and standard calibration ...................................................................................................... 46 2.3.1.1 Double-spiked Mo(UBC) standards ............................................................. 46 2.3.1.2 Analyses of other Mo isotope standards and natural reference materials ........................................................................................ 50 2.3.2 Importance of spike-to-sample ratios ................................................................. 51 2.3.3 Evaluation of column chemistry artefacts on isotope measurements ............................................................................................................... 53 2.4 Conclusions ............................................................................................................... 55 CHAPTER 3. Molybdenum and zinc stable-isotope variations in mining waste-rock drainage and waste rock at the Antamina mine, Peru .................................... 57   xiv   3.1 Introduction ............................................................................................................... 57 3.2 Study site ................................................................................................................... 61 3.3 Methods .................................................................................................................... 64 3.3.1 Sample selection and preparation ....................................................................... 64 3.3.2 Analytical methods ............................................................................................. 66 3.4 Results and Discussion ............................................................................................. 70 3.4.1 Molybdenum isotopic composition of waste rock, ores and mine drainage ........................................................................................................ 70 3.4.2 Zinc stable-isotope composition of waste rock, sphalerites, and mine drainage ..................................................................................... 82 3.4.3 Implications for use of metal stable-isotope ratios to monitor mine drainage geochemistry ........................................................................... 90 CHAPTER 4. Molybdenum stable isotopic variations as indicators of Mo attenuation in mine drainage from full-scale waste-rock storage facilities .................................................................................................................................... 93 4.1 Introduction ............................................................................................................... 93   xv   4.2 Study site ................................................................................................................... 96 4.2.1 Local climate and geology .................................................................................. 96 4.2.2 Site hydrogeology ............................................................................................... 98 4.3 Sample collection and preparation ............................................................................ 99 4.3.1 Water sampling and storage ................................................................................ 99 4.3.2 Rock sampling and preparation ........................................................................ 102 4.3.3 Characterization of weathered waste-rock surfaces ......................................... 103 4.4 Analytical methods ................................................................................................. 105 4.5 Results ..................................................................................................................... 107 4.5.1 Source characterization: Rocks ......................................................................... 107 4.5.2 Source characterization: Process water and runoff from the mine’s open pit ..................................................................................................... 111 4.5.3 Groundwater, surface water, and waste-rock storage facility drainage .......................................................................................................... 117 4.5.3.1 Buckskin Creek Catchment ........................................................................ 118   xvi   4.5.3.2 Pat Hughes Creek Catchment ..................................................................... 122 4.5.4 Characterization of weathered waste-rock surfaces ......................................... 123 4.6 Discussion ............................................................................................................... 124 4.6.1 Absence of isotopic fractionation during molybdenite oxidative dissolution .................................................................................................. 125 4.6.2 Evidence for the role of adsorption in controlling Mo mobility in waste rock ................................................................................................ 126 4.6.3 Molybdenum attenuation in WRSF drainage ................................................... 130 4.6.4 Use of δ98Mo and Mo/SO42- to trace Mo sources in shallow groundwater (well BW4) .............................................................................. 133 4.7 Conclusions ............................................................................................................. 134 CHAPTER 5. Molybdenum stable-isotope fractionation during the precipitation of powellite (CaMoO4) and wulfenite (PbMoO4) ........................................ 137 5.1 Introduction ............................................................................................................. 137 5.2 Methods .................................................................................................................. 141   xvii   5.2.1 Experimental methods ...................................................................................... 141 5.2.2 Powellite precipitation experiments ................................................................. 142 5.2.3 Wulfenite precipitation experiments ................................................................. 145 5.2.4 Analytical methods ........................................................................................... 146 5.2.4.1 Elemental and isotopic analyses ................................................................. 146 5.2.4.2 Mineralogy by X-ray diffraction................................................................. 148 5.2.4.3 Aqueous speciation modelling .................................................................... 148 5.3 Results ..................................................................................................................... 149 5.3.1 Powellite precipitation ...................................................................................... 149 5.3.1.1 CaCl2 series ................................................................................................. 149 5.3.1.2 CaCO3/SO42- series ..................................................................................... 156 5.3.2 Wulfenite precipitation ..................................................................................... 158 5.4 Discussion ............................................................................................................... 162   xviii   5.4.1 Molybdenum isotopic fractionation during powellite precipitation ................................................................................................................ 162 5.4.1.1 Kinetic isotope effect (KIE) ........................................................................ 162 5.4.1.2 Equilibrium fractionation after advanced reaction progress ....................................................................................................... 167 5.4.1.3 Effect of sulfate and/or DIC inhibition on powellite precipitation ................................................................................................ 170 5.4.2 Molybdenum isotopic fractionation during wulfenite precipitation ................................................................................................................ 171 5.4.3 Molybdate removal mechanisms and environmental implications ................................................................................................................ 175 5.5 Summary and conclusions ...................................................................................... 180 CHAPTER 6. Molybdenum attenuation mechanisms in variably oxic mine tailings drainage: insights from Mo stable isotopes and X-ray spectroscopy .......................................................................................................................... 183 6.1 Introduction ............................................................................................................. 183   xix   6.2 Study site ................................................................................................................. 189 6.2.1 Bruno Creek Tailings Management Facility ..................................................... 189 6.2.2 Hydrology and geochemistry of the BCTMF ................................................... 189 6.3 Sample collection and analyses .............................................................................. 192 6.3.1 Field sampling .................................................................................................. 192 6.3.2 Molybdenum sorption experiment .................................................................... 193 6.3.3 Analytical methods ........................................................................................... 194 6.3.3.1 Elemental abundance and Mo isotopic composition .................................. 194 6.3.3.2 Ochreous precipitate mineralogical characterization .................................. 195 6.4 Results ..................................................................................................................... 197 6.4.1 Aqueous geochemistry ...................................................................................... 197 6.4.2 Molybdenum sorption experiment .................................................................... 203 6.4.3 Ochreous precipitate sample characterization .................................................. 204 6.4.3.1 Geochemistry of ochreous precipitates ....................................................... 204   xx   6.4.3.2 Mineralogical characterization of ochreous precipitates .................................................................................................. 208 6.4.3.2.1 X-ray absorption spectroscopy .................................................. 211 6.5 Discussion ............................................................................................................... 218 6.5.1 Processes of Mo attenuation in the BCTMF..................................................... 218 6.5.2 Comparing molybdenum mobility in anoxic mine tailings to waste-rock environments ........................................................................... 225 6.5.3 Molybdenum attenuation during oxidation of Fe(II)-rich tailings drainage in the SRD ...................................................................................... 227 6.5.3.1 Determination of isotopic separation factor for sorption to Fe-(oxyhydr)oxides .................................................................. 228 6.5.3.2 Molecular mechanism of Mo isotopic fractionation and sorption to Fe-(oxyhydr)oxides ............................................................ 229 6.5.3.3 Multiple bonding configurations in sample SRD-S1 .................................. 234 6.6 Summary and conclusions: Mo attenuation behavior in mine waste ....................................................................................................................... 235   xxi   CHAPTER 7. Conclusion ......................................................................................................... 238 7.1 Overview of research objective .............................................................................. 238 7.2 Summary of main findings ..................................................................................... 238 7.2.1 Analytical protocol development ...................................................................... 238 7.2.2 Application of the stable-isotope fractionation of Mo (and Zn) in mine waste ............................................................................................... 239 7.2.2.1 Zinc stable-isotope fractionation and attenuation in mine waste .................................................................................................. 240 7.2.2.2 Molybdenum stable-isotope fractionation and attenuation in mine waste ........................................................................... 241 7.3 Future research directions ....................................................................................... 245 References ....................................................................................................................................249 Appendices ...................................................................................................................................287 APPENDIX A. Summary δ98Mo values measured for reference materials and Mo isotope standards analyzed during this thesis .................................... 288   xxii   A.1 Introduction .............................................................................................................. 288 APPENDIX B. Development of an ion-exchange protocol for the separation of Mo and Zn from environmental samples for isotopic analyses…………………. ..................................................................................................... 291 B.1 Introduction .............................................................................................................. 291 B.2 Column Test PEA1: Evaluation of the method of Pearce et al. (2009) ......................................................................................................................... 292 B.3 Column Tests PEA2A and PEA2B: Modifications to the method of Pearce et al. (2009) to remove Zn from the Mo elution peak ................................................................................................................ 295 B.4 Column Test asma3a ................................................................................................ 297 B.5 Validation of ion-exchange protocol using BCR-2 and mine waste samples from Antamina ................................................................................... 299 B.6 Final ion-exchange chromatography method and Mo and Zn column blank compilation .......................................................................................... 301 APPENDIX C. Supplementary information to Chapter 3 .................................................... 308   xxiii   APPENDIX D. Supplementary information to Chapter 4 .................................................... 312 APPENDIX E. Supplementary tables to Chapter 5. ............................................................. 321 APPENDIX F. Determination of powellite mineral saturation indices using PHREEQC for CaCO3/SO42- series .......................................................................... 324 F.1 Problem Formulation ................................................................................................ 324 F.2 Determination of HCO3- and DIC from charge balance ........................................... 326 F.3 Bracketing the HCO3- and DIC concentrations using a degassing simulation in PHREEQC ........................................................................... 328 F.3.1 Degassing model conceptual description .......................................................... 329 F.3.2 Degassing model results .................................................................................... 330 F.4 Determination of powellite SIs using estimated DIC concentrations ............................................................................................................ 331 F.4.3 Description of modelling approach ................................................................... 331 F.4.4 Results ............................................................................................................... 332 F.5 Conclusion ................................................................................................................ 333   xxiv   APPENDIX G. Conversion table between molar and mass-based concentration units for molybdenum ................................................................................. 335 APPENDIX H. Detailed descriptions of sampling and analytical methods used in for Chapter 6 ............................................................................................ 336 H.1 Field water-sampling protocols ................................................................................ 336 H.2 Laboratory analyses of elemental abundance and Mo isotopic compositions of water and ochreous precipitate samples .......................................... 337 H.3 Ochreous precipitate mineralogical characterization ............................................... 338 H.3.5 X-ray diffraction ............................................................................................... 338 H.3.6 Raman spectroscopy ......................................................................................... 339 H.3.7 Transmission electron microscopy ................................................................... 340 H.3.8 Synchrotron-based X-ray absorption spectroscopy .......................................... 340 APPENDIX I. Supplementary photos, tables, and figures accompanying Chapter 6 ..................................................................................................... 342   xxv   List of Tables Table 2.1. Faraday cup configuration for Pd-doped analysis of double spike and standard during calibration session. ................................................................. 41 Table 2.2. Faraday cup configuration for Mo analysis of pure double spike and standard during calibration session. ................................................................. 43 Table 2.3. Typical instrument settings for Mo isotope analysis on the Nu MC-ICP-MS. .................................................................................................................... 46 Table 2.4. Calibrated double spike and standard values. .............................................................. 47 Table 2.5. Comparison of δ98/95Mo values obtained in this study relative to published values. .......................................................................................................... 51 Table 2.6. Values of δ98/95Mo for Mo(UBC) standards double spiked before and after column chromatography. ........................................................................ 54 Table 2.7. Comparison of δ98/95Mo measurements using double spike, SSB, and SSB + Pd for mass bias corrections. ................................................................. 55 Table 3.1. Molybdenum isotope data of mine drainage and rock and mineral samples from Antamina ....................................................................................... 72   xxvi   Table 3.2. Zinc isotopic compositions and Zn content in mine drainage, sphalerites, and waste rock from the Antamina mine ....................................................... 86 Table 4.1. Description of water sampling locations ................................................................... 101 Table 4.2. Molybdenum abundance and δ98Mo in waste rock, ore, and tailings. ........................................................................................................................... 110 Table 4.3. Water-sample chemistry. ........................................................................................... 113 Table 5.1. Initial concentrations and conditions for powellite and wulfenite precipitation experiments. .............................................................................. 144 Table 5.2. Geochemical data for samples drawn from powellite precipitation experiments. .............................................................................................. 153 Table 5.3. Geochemical data for samples drawn from wulfenite precipitation experiments ............................................................................................... 161 Table 5.4. Summary of kinetic and equilibrium Mo isotopic fractionation factors in powellite precipitation experiments. ............................................................... 164 Table 5.5. Equilibrium Mo isotope fractionation factors in wulfenite precipitation experiments. .............................................................................................. 174   xxvii   Table 5.6. Summary of known Mo isotopic fractionation factors for Mo removal processes from aqueous solutions .................................................................... 177 Table 6.1. Water-sample chemistry: Field parameters and anions concentrations ................................................................................................................. 199 Table 6.2. Water-sample chemistry: Metal concentrations and Mo isotope ratios ............................................................................................................................... 200 Table 6.3. Water-sample chemistry: Mineral saturation indices (SI) ......................................... 201 Table 6.4. Elemental abundances and Mo isotope ratios in ochreous precipitate samples ......................................................................................................... 207 Table 6.5. Linear combination fitting results for Mo K-edge XANES spectra of ochreous precipitate samples. ........................................................................ 215 Table 6.6. Summary of Mo K-edge EXAFS fitting parameters for ochreous precipitate samples. ......................................................................................... 215 Table A.1. Values of δ98Mo for Mo isotope standards and geological reference materials analyzed as part of this thesis. ..................................................................... 290 Table B.1. Ion-exchange column test “PEA1”. .......................................................................... 293   xxviii   Table B.2. Ion-exchange column test “PEA2A”. ....................................................................... 296 Table B.3. Ion-exchange column test “PEA2B”. ........................................................................ 296 Table B.4. Ion-exchange column test “asma3a”. ........................................................................ 298 Table B.5. Recovery of Mo and Zn during ion-exchange chromatography of test samples using the separation scheme used in this thesis. .................................... 300 Table B.6. Final anion-exchange chromatography protocol used for Mo and Zn separations in this thesis ..................................................................................... 302 Table B.7. Compilation of Mo and Zn column blanks analyzed during this thesis ........................................................................................................................ 303 Table C.1. Mineralogical abundances (in wt. %) in Antamina waste-rock samples of different lithologies, as determined by XRD. ............................................... 309 Table D.1. Comparison of measured and expected concentrations for a secondary multi-element standard† analyzed on the ICP-OES....................................... 312 Table D.2. Replicate analyses of ICP standards at different concentrations showing the ICP-OES instrumental precision. ....................................... 313   xxix   Table D.3. Relative standard deviation for replicate analyses of calibration standards in ICP-MS analytical sessions. ..................................................... 315 Table D.4. Trace element concentrations in BCR-2 determined by ICP-MS in this study, compared with previously published values. ..................................... 316 Table D.5.a. Water sample duplicates by ICP-OES (cations) and ion chromatography (anions) ................................................................................................ 317 Table D.6.a. Elemental abundances and δ98Mo in 2-step sequential chemical extractions of weathered waste-rock samples ................................................. 318 Table D.7. Elemental abundances and Mo isotopic compositions in the secondary Al-(hydroxy)sulfate precipitate forming at the base of the Pat Hughes WRSF .................................................................................................... 319 Table D.8. Occurrence of selected minerals in weathered waste rock surfaces as identified by X-ray diffraction ..................................................................... 319 Table E.1. Summary of replicate and duplicate MC-ICP-MS sample analyses ............................................................................................................................321 Table E.2. Values δ98Mo for reference materials and standards passed through ion-exchange chemistry along with samples in Chapter 5. ............................... 322   xxx   Table E.3. Calculation of the equilibrium fractionation factor by isotopic mass balance for powellite precipitation in experiment POW3 from samples collected close to chemical equilibrium. .................................................. 323 Table F.1. DIC and HCO3- concentrations determined in PHREEQC from charge balance, and corresponding powellite saturation indices. ....................................328 Table I.1. Calculation of the isotopic separation factor (Δ98Mo) between   aqueous Mo and Mo sorbed on ferrihydrite.................................……………………..348 Table I.2. Summary of replicate sample analyses by MC-ICP-MS ............................................ 349    xxxi   List of Figures Figure 1.1. Examples of waste-rock storage facilities showcasing their size. ..................................................................................................................................... 6 Figure 1.2. The mine waste-rock environment. .............................................................................. 7 Figure 1.3. Hydrogeochemical processes impacting mine waste drainage. ................................... 8 Figure 1.4. A Nu multi-collector inductively coupled plasma mass spectrometer (Nu 021). ....................................................................................................... 9 Figure 1.5. Molybdenum Eh-pH diagram for a Mo–H2O–H2S system. ....................................... 15 Figure 1.6. Molybdenum stable-isotope compositions of various Earth materials. ........................................................................................................................... 17 Figure 1.7. Zinc stable-isotope compositions for natural materials found at the surface of the Earth. ................................................................................................ 21 Figure 1.8. Flow-chart linking the different research chapters composing this thesis. ......................................................................................................................... 27 Figure 2.1. Flow chart for the design, calibration, and implementation of a double spike for stable-isotope analysis by MC-ICP-MS. ............................................ 34   xxxii   Figure 2.2. Flow chart for the design, calibration, and implementation of a double spike for stable-isotope analysis by MC-ICP-MS. ............................................ 35 Figure 2.3. ln-ln plots of measured Pd and Mo isotope ratios during the Pd-doped Mo(UBC) standard calibration. ........................................................................ 40 Figure 2.4. Analyses of δ98/95Mo for the Mo(UBC) standard over a range of double-spike-to-standard ratios. ................................................................................... 48 Figure 2.5. Variation of δ98/95Mo values for Mo(UBC) standard over an analytical session. ............................................................................................................. 50 Figure 2.6. Values of δ98/95Mo for Mo(UBC) standards over a range in spike-to-sample ratios after normalization to the daily average value of the standard. ........................................................................................................ 53 Figure 3.1. Location of the Antamina Mine, Peru. ....................................................................... 62 Figure 3.2. Molybdenum isotopic compositions of mine drainage from experimental field barrels, and waste rock piles, and of molybdenites, waste-rock and ore concentrate. ................................................................ 73 Figure 3.3. Molybdenum isotopic compositions against Mo concentrations and pH in Antamina mine drainage. ......................................................... 79   xxxiii   Figure 3.4. Seasonal variation in powellite and wulfenite mineral saturation indicies (SI) in Antamina field-barrel drainage. .............................................. 81 Figure 3.5. Zinc isotopic compositions in mine drainage and sphalerites at Antamina. ...................................................................................................................... 84 Figure 3.6. Zinc isotopic compositions against Zn concentrations and pH in Antamina mine drainage. .............................................................................................. 85 Figure 3.7. Seasonal variation in ZnCO3•H2O, hydrozincite, and Zn(OH)2 mineral saturation indices (SI) in Antamina field-barrel drainage. .................................. 88 Figure 4.1. Map of the Thompson Creek Mine waste-rock storage facilities showing sampling locations and surficial geology. ........................................... 98 Figure 4.2. Molybdenum isotopic composition and Mo contents of waste- rock, ore, and mine tailings samples. .............................................................................. 112 Figure 4.3. Molybdenum isotopic compositions of water and solid-phase samples from the Thompson Creek Mine. ...................................................................... 118 Figure 4.4. Cross-sectional views along the Pat Hughes Creek Catchment showing concentrations of Mo and SO42- and δ98Mo in water samples. .......................................................................................................................... 120   xxxiv   Figure 4.5. Cross-sectional views along the Buckskin Creek Catchment showing concentrations of Mo and SO42- and δ98Mo in water samples. .......................................................................................................................... 121 Figure 4.6. Molybdenum and Fe contents and δ98Mo from sequential chemical extractions of weathered waste rock. .............................................................. 124 Figure 4.7. Molybdenum isotopic compositions against Mo/SO42- in samples from the Buckskin Creek and Pat Hughes Creek Catchments. .................................................................................................................... 128 Figure 4.8. Molybdenum isotopic compositions against Mo/SO42- for all water samples. ................................................................................................................ 132 Figure 5.1. Aqueous speciation at the start of powellite experiments. ....................................... 151 Figure 5.2. Decreasing Mo concentrations and powellite saturation indices over time during powellite precipitation experiments. ....................................... 152 Figure 5.3. Molybdenum concentrations, δ98Mo, and powellite SI vs. time during powellite precipitation experiments. ........................................................... 157 Figure 5.4. Molybdenum removal from solutions with the addition of Pb in wulfenite precipitation experiments. .......................................................................... 159   xxxv   Figure 5.5. Increasing δ98Mo during Mo removal via wulfenite precipitation. ................................................................................................................... 160 Figure 5.6. Isotopic enrichment factors (ε) determined for kinetic isotope fractionation during powellite precipitation under supersaturated conditions. ....................................................................................................................... 164 Figure 5.7. Best-fit curves for Mo removal via powellite precipitation using the Rayleigh model. .............................................................................................. 166 Figure 5.8. Molybdenum isotope fractionation during wulfenite precipitation. ................................................................................................................... 173 Figure 5.9. Isotopic separation factors (∆98Mo) for equilibrium isotope fractionation during wulfenite precipitation. .................................................................. 174 Figure 5.10. Fractionation factors for Mo removal processes from aqueous solutions. ........................................................................................................... 178 Figure 6.1. Conceptualized cross-section of the Bruno Creek Tailings Management Facility showing the major flowpaths and design features of the facility (modified from Lorax, 2011b). ................................................... 187 Figure 6.2. Photographic overview of the study site. ................................................................. 188   xxxvi   Figure 6.3. SO42-, Fe, Mo, and δ98Mo in basal tailings drainage at the Rock Toe in comparison to the SRD pond. .................................................................... 202 Figure 6.4. Concentrations of dissolved Fe, Mo, and Mn during the oxidation of a 20-L pail of Rock Toe water. .................................................................. 204 Figure 6.5. Comparison of δ98Mo values in the SRD, Rock Toe, and Fe-(oxyhydr)oxides, shown along with TCM mine tailings, process water, waste rock, and ore δ98Mo from Skierszkan et al. (2017). .................................. 206 Figure 6.6. Select raman spectra for ochreous precipitate samples. ........................................... 210 Figure 6.7. Selected transmission-electron microscopy images of Fe-(oxyhydr)oxide ochreous precipitates collected in the SRD. ......................................... 211 Figure 6.8. Normalized Mo K-edge X-ray absorption near-edge spectra of reference compounds and ochreous precipitate samples. ........................................... 216 Figure 6.9. Molybdenum K-edge k3-weighted χ(k) background-subtracted spectra (left) and the corresponding Fourier transforms (right). .................................... 217 Figure 6.10. Molybdenum isotopic compositions against Mo concentrations in the flowpath from the tailings pond to the Rock Toe and SRD pond. ........................................................................................................ 221   xxxvii   Figure 6.11. Simulations of the effect of Mo removal from mine process water on δ98Mo as a function of Mo/Cl ratios. ............................................................... 221 Figure 6.12. Molybdenum isotope ratios against Fe/Mo molar ratios in Fe-(oxyhydr)oxide-rich ochreous precipitate samples collected in the SRD pond. ................................................................................................................. 231 Figure 7.1. Summary of Mo attenuation processes and the direction of associated isotopic fractionation (α). .............................................................................. 243 Figure 7.2. Conceptual flow-chart demonstrating the implementation of metal stable-isotope analyses to trace attenuation along with other techniques. ...................................................................................................................... 246 Figure A.1. Compilation δ98Mo analyses of Mo isotope standards (panels a to d) and geological reference materials (panel e) analyzed during this thesis. ............................................................................................................ 289 Figure B.1. Elution profile for column test PEA1. ..................................................................... 294 Figure B.2. Elution profile for column test PEA2A. .................................................................. 304 Figure B.3. Elution profile for column test PEA2B. ................................................................... 305   xxxviii   Figure B.4. Elution profile for column test asma3a. ................................................................... 306 Figure B.5. Column calibration for BCR-2 and mine waste samples. ........................................ 307 Figure C.1. Photos of waste rock weathering field experiments conducted to evaluate drainage water quality at the Antamina mine, Peru. .................................... 308 Figure C.2. Three-isotope plot of δ68Zn against δ66Zn for all samples analyzed in Chapter 3 showing an absence of mass-independent fractionation and isobaric interferences on Zn isotope analyses. ................................... 308 Figure F.1. Modelled geochemical evolution of experiment POW2B2 as a result of the introduction of atmospheric headspace and CO2 degassing. 331 Figure F.2. Powellite saturation indices calculated in PHREEQC for all samples in experiment POW2B2, using different estimates of DIC. ................................................................................................................................ 333 Figure I.1. Photos of the BCTMF. .............................................................................................. 342 Figure I.2. Close-up of the SRD ochreous precipitate sampling locations. ................................ 343 Figure I.3. Photographic log showing the progressive oxidation of Rock Toe water in the Mo sorption experiment. ..................................................................... 344   xxxix   Figure I.4. Additional transmission-electron microscopy images of SRD ochreous precipitate samples with selected area electron diffractograms (SAED). .................................................................................................. 345 Figure I.5. X-ray diffractograms for ochreous precipitate samples. ........................................... 346 Figure I.6. Raman spectrum from sample SRD-S2 showing the presence of quartz. ......................................................................................................................... 347    xl   List of Abbreviations and Symbols BCTMF   Bruno Creek Tailings Management Facility DIC   Dissolved inorganic carbon EXAFS    Extended X-ray absorption fine structure ICP-MS   Inductively coupled plasma mass spectrometry ICP-OES   Inductively coupled plasma optical-emission spectroscopy LCF   Linear-combination fitting MC-ICP-MS   Multi-collector inductively coupled plasma mass spectrometry Mo    Molybdenum RSD   Relative standard deviation SAED   Selected area electron diffractogram SE   Standard error SD   Standard deviation mineral SI  Mineral saturation index    SSB    Standard-sample bracketing TCM   Thompson Creek Mine TIC   Total inorganic carbon vol. %   Volumetric percent. WRSF   Waste rock storage facility wt. %    Weight percent XANES   X-ray absorption near-edge structure XAS   X-ray absorption spectroscopy XRD   X-ray diffraction Zn   Zinc αx-y Isotopic fractionation factor; e.g. the ratio of 98Mo/95Mo in a given phase x over the ratio in another phase y ‰ per-mil unit notation ε Isotopic enrichment factor; ε = 1000 x (α – 1) (in per mil units) Δ98Mo Isotopic separation factor; the difference in δ98Mo between reactant and product (in per mil units)   xli   δ66Zn   Zinc stable-isotope ratio (in per mil units) δ98Mo and δ98/95Mo Molybdenum stable-isotope ratio (in per mil units)    xlii   Acknowledgements Producing this thesis has provided me with tremendous learning and growth, and I am thankful to everyone that supported me along the way. First and foremost, thank you to my parents, Simon and Martine, who nurtured my curiosity and taught me to give my best effort all along. My decision to purse graduate studies in geoscience is in large part due to inspiring mentorship towards the end of my BSc studies in Ottawa. Thank you to: Ian Clark for sparking my interest in aqueous geochemistry; Jules Blais for encouraging me to publish my BSc thesis and setting me off on the right track; and Krista Trounce for making me feel like a competent and valuable geoscientist as a coop student at Intera (Geofirma). Thank you all for your mentorship. I benefited from a wonderful advisory committee throughout my PhD. Marghaleray Amini patiently taught me the basics of working in a clean laboratory and was a stalwart supporter during my struggles with analytical method development. Dominique Weis always kept her door open and provided me with honest advice and sharp technical feedback. My advisors, Roger Beckie and Uli Mayer, provided constant support and the right balance of freedom and input to help me grow as a researcher. I commend my committee for their unwavering dedication to student well-being and academic growth, for offering their time and assistance without hesitation, and for being a constant source of support and inspiration. I could not have asked for a better cast of academic mentors. In addition to my committee members, I was surrounded by an incredible team of research scientists. From the PCIGR team, I thank Vivian Lai, Bruno Kieffer, and Kathy Gordon in particular for training me in hands-on geochemistry. Thanks also to Dr. Richard Friedman for constantly ensuring that the lab runs smoothly. I also thank other research staff in EOAS, in particular Elisabetta Pani, Jenny Lai, Lan Kato, Matthijs Smit, Hai Lin, Taylor Ockerman, and Maureen Soon, for training me on various scientific instruments and equipment. My research collaboration with Matt Lindsay, Jared Robertson, and Soumya Das at the University of Saskatchewan’s Environmental Geochemistry Lab provided me with a wonderful academic exchange towards the end of my thesis. It was a pleasure to work with you and I hope that we will put our minds together again in the future.   xliii   This thesis would not have been possible without industrial support. The applicability of my research to challenges faced by industry provided a steady source of motivation. I thank the Antamina Mine staff for their tremendous commitment to the Antamina waste-rock project for so many years. It was a pleasure to meet many of you in person and I wish you all the best. The support from the staff at the Thompson Creek Mine for my project—especially Richard Giampedraglia and Bert Doughty— is deeply appreciated. Thank you. The opportunity to work at both the TCM and Antamina provided me with a lot of perspective on the mining industry in addition to allowing me to meet some of the very best people. While not officially part of my advisory committee, this thesis also benefitted from a great deal of input from two great guys and great geochemists at Lorax Environmental Services, Justin Stockwell and John Dockrey. Thanks for finding a way to include me as a MAGNET intern with Lorax and supporting my involvement with the TCM. That project provided me with a great deal of learning and was a big part of my decision to transition from MSc to PhD studies.  With my involvement in both the PCIGR and Hydrogeology research groups, there is a very long list of fellow students and post-doctoral fellows that provided laughs, peer-support, empathy, and feedback along the way. Thank you all! I wish to each of you successful and fulfilling careers beyond EOAS. I wish to thank Laura Laurenzi, Anaïs Fourny, Lauren Harrison, and Mehrnoush Javadi in particular for always openly answering my questions. Laura Laurenzi, Gregor Lucic, Mélanie St-Arnault, and María Lorca-Ugalde all put in a solid effort to help me secure the samples I needed to complete this research—overcoming soroche, fatigue, and food poisoning all in the name of science! This thesis also benefitted from a great deal of financial support, without which none of this would have been possible. Thank you to NSERC, to Lorax Environmental Services, to the Antamina and Thompson Creek mines, to UBC EOAS, and to the Canadian Institute for Mining, Metallurgy and Petroleum for supporting this research.  Thank you to my friends, for all of the fun time and emotional support along the way. En particulier, merci à Alex Bevington, Marie-Pierre Varin, Stéphane Leahy, Ellorie McKnight, Ashlin Kelly et Victoria Ho: c'est un plaisir d'avoir retenu votre amitié au fil du temps, malgré les distances.  Finally, my biggest thank you goes to my partner Emily Rossnagel. Your unwavering support, trust, and cheer give me the confidence and strength to tackle any challenges that may come my way. You have taught me a great deal about life. I can’t wait for our next adventures together!   xliv   Dedication I wish to dedicate this work to my fellow humans around the world: May we continue to work together towards sustainable interaction with our planet!  1   CHAPTER 1. INTRODUCTION 1.1 Environmental challenges posed by mining waste Human societies are inextricably reliant on minerals; these are fundamental components in the majority of the goods that we use daily, from vehicles and electronics to tools and building materials. Products processed from minerals in 2009 accounted for 3 % of the U.S. national GDP—for a total of $US 454 billion (Keller, 2012). Insatiable global demand for consumer products which are increasingly dependent on metals is driving growth in mineral demand and extraction (Langkau and Tercero-Espinoza, in press).  A byproduct of such large-scale resource mineral extraction is the production of a tremendous volume of waste rock and tailings which contain residual metals at abundances uneconomical for processing. Waste rock and tailings compose the “largest volume of materials handled in the world” (Blowes et al., 2014); the U.S. alone adds 1 billion to 2 billion tons of mine waste annually to the 50 billion tons that they already produced over the last century (Keller, 2012). This waste accumulates in stockpiles at mine sites and can contain tens to hundreds of millions of tons of material and reach hundreds of meters in height (Figure 1.1). It is also a critical environmental concern because of the release of potentially toxic metals into water resources. Metal leaching can be caused by the weathering of metal-bearing sulfide minerals, which are chemically unstable in the presence of oxygen and water and therefore dissolve in conditions typical of mine waste facilities (Figure 1.2).   2   This metal release often requires water treatment—at substantial cost—to mitigate environmental and human health risks and adhere to regulatory guidelines. For example, in Canada, billions of dollars have been spent to remediate sites impacted by mine runoff (Nordstrom et al., 2000). In general, metal concentrations in mine drainage are controlled by the dissolution and precipitation of minerals, surface adsorption, and hydrological transport. Environmental practices at mine sites would be improved by a better understanding of these processes in mine waste, which would lead to better strategies to mitigate water contamination and better predictions of water treatment requirements in the decades following waste disposal.  1.2 Analytical techniques available to study metal attenuation A variety of techniques are currently available to characterize the fate of metals of potential environmental concern analyses of solid and aqueous mine waste samples. Examples of solid-phase analyses include: (1) determination of elemental abundance using mass spectrometry and/or X-ray fluorescence; (2) microscopic imaging and elemental mapping using scanning-electron microscopy; (3) mineral identification by X-ray diffraction (XRD) and Raman spectroscopy; (4) chemical sequential extractions to indicate under which environmental conditions (e.g. pH, redox) metals are mobile; and (5) X-ray absorption spectroscopy (XAS) to constrain the chemical coordination environment of metals (Jamieson et al., 2015). Nevertheless, solid-phase analyses require weathered mine tailings or waste rock samples, which are difficult to access due to the high cost of drilling into waste dumps for sample collection. Alternatively,  3   monitoring the concentration of metals in mine drainage allows inferences to be made to constrain the processes that control metal transport in waste facilities (Figure 1.3). Mine drainage samples are readily available and therefore provide important windows into processes internal to waste impoundments. In addition to these aforementioned approaches, the fate of metals in mine waste may be further constrained by a promising new technique: analyses of metal stable isotopes. Stable isotopes are atoms of a given chemical element that differ in the number of neutrons in their nuclei. Because the thermodynamic properties of chemical bonds are partially related to the mass of the atoms involved, changes in the number of neutrons (and therefore atomic masses) cause differences in the chemical reactivity among isotopes of a given element. Consequently, stable isotopes can become unevenly distributed during chemical reactions; this process is termed “isotopic fractionation” (Urey, 1931; Urey and Grieiff, 1935). Isotopic fractionation can be used to trace the occurrence of certain chemical reactions in the environment on the basis that they cause a shift isotopic ratios (e.g. Wiederhold et al., 2015).  Until recently, stable-isotope analyses were generally restricted to the lighter non-metallic elements C, H, N, O, and S that are suitable for gas-source mass spectrometry. Isotopic analyses of those elements since the 1930s have proven extremely useful in tracing their biogeochemical cycles (e.g. Bigeleisen and Mayer, 1947; Nier, 1940, 1939, Urey, 1947, 1931). The environmental fate of heavier elements—i.e. most metals of potential environmental concern  4   in mine waste—might also be possible to constrain by stable-isotope analyses.  However, stable-isotope variations for heavier elements are much smaller than those of the lighter elements (Wiederhold et al., 2015). In addition, heavier elements are not easily converted to gaseous form and are therefore poorly suited to gas-source mass spectrometry. It is only recently that the development of a new analytical instrument—the multi-collector inductively coupled plasma mass spectrometer (MC-ICP-MS)—has overcome analytical barriers to metal-stable isotope analyses [Figure 1.4; see also Walder and Freedman (1992)].  Defining features of the MC-ICP-MS include: (1) the inductively coupled plasma, whose elevated temperatures (6,000 to 10,000 °K) efficiently ionize elements throughout the periodic table to provide positively charged ions to the mass spectrometer (Wolf, 2005); (2) the mass analyzer, comprising an electrostatic analyzer and a magnetic sector that focus the ion beam in the mass spectrometer and separate isotopes on the basis of their mass-to-charge ratio (Wolf, 2005); and (3) an array of ion detectors that enable simultaneous counting of several isotopes for high-precision data acquisition. Consequently, the high ionization efficiency and unprecedented precision and accuracy of MC-ICP-MS enables inter-sample resolution of metal stable-isotope fractionations despite their comparably small magnitude. It turns out that most metals of potential environmental concern at mine sites (e.g. Mo, Zn, Cr, Hg, Cu) comprise many stable isotopes. This observation, along with the emerging capability of MC-ICP-MS, has opened the door to applying metal isotope analyses to track geochemical attenuation at mine sites (e.g. Smith et al., 2015).  5   This dissertation is therefore centered upon the hypothesis that interpretation of metal stable-isotope variations can indicate attenuation processes in mine waste. In the following sections, additional background about this technique and its environmental applications is provided.    6    Figure 1.1. Examples of waste-rock storage facilities showcasing their size. (A) The East Dump at the Antamina Mine, Peru is close to 300 m in height and many kilometers wide. (B) Buckskin waste-rock storage facility at the Thompson Creek Mine, Idaho. Photo (A) by Roger Beckie; photo (B) by Elliott Skierszkan.  7    Figure 1.2. The mine waste-rock environment. (A) Waste-rock storage facility typical of a modern large-scale open-pit operation. These facilities can be hundreds of meters in height and contain hundreds of millions of tons of sulfide-bearing waste rock. (B) Waste rock contains variable amounts of sulfide minerals; such as the chalcopyrite (Cu2S)-bearing sample shown here. (C) Oxidation of sulfides leads to leaching of metals and precipitation of a variety of secondary minerals such as sulfosalts (white minerals in this photo), oxides (red/brown minerals in this photo), and (hydroxy)carbonates (turquoise minerals in this photo). (D) Water percolating through mine waste rock can become enriched in dissolved sulfate and metals and must be contained or treated to protect the surrounding environment. All photos by Elliott Skierszkan.  8   Figure 1.3. Hydrogeochemical processes impacting mine waste drainage. Atmospheric oxygen and water enter a mine-waste storage facility, causing the oxidation of sulfide minerals and liberation of metals and sulfate. Released metals can be attenuated via adsorption and secondary mineral precipitation. Chemical reactions shown are examples of possible attenuation reactions relevant to molybdenum (further explained in section 1.5), including pyrite (FeS2) and molybdenite (MoS2) oxidative dissolution, adsorption of molybdate (MoO42-) on metal (oxyhydr)oxide surfaces (SOH), and precipitation of the secondary mineral powellite (CaMoO4). Seepage at the base of the waste facility is contaminated with sulfate and metals. Gray-shading represents compaction surfaces and tipping phases, with coarser grain size occurring at the bottom of tipping phases as a result of sorting during waste-rock dumping. Compaction surfaces, tipping phases and granulometry impact transport processes. The scale is approximately representative of mine-waste facilities at many modern large-scale operations.    9    Figure 1.4. A Nu multi-collector inductively coupled plasma mass spectrometer (Nu 021). Labels show a selection of key components in the instrument. The sample solution is introduced into the plasma where high temperatures ionize its constituents. Turbo pumps create a vacuum and draw the sample into a flight path through the mass spectrometer, where the electrostatic analyzer and the magnetic sector are used to focus the sample ion beam and separate individual isotopes based on their charge/mass. Positively charged ion beams create a current that is measured by the Faraday Cup detectors, from which isotope ratios are determined. Photo by Dominique Weis. 1.3 Mechanisms of stable-isotope fractionation Stable-isotope fractionation of metals comprises mass-dependent and mass-independent fractionation types (MDF and MIF, respectively). MIF is most applicable to the heaviest elements and those susceptible to photochemical reactions (e.g. U, Hg; Buchachenko, 2013), and therefore falls outside of the scope of this thesis. On the other hand, MDF is widespread among all elements comprising multiple stable isotopes (e.g. Wiederhold et al., 2015). Mass-dependent fractionation was discovered in the first half of the 20th Century when basic mass spectrometry  10   and isotopic fractionation were first explored (Bigeleisen and Mayer, 1947; Nier, 1940, 1939, Urey, 1947, 1931).  Mass-dependent fractionation originates from two main processes: (1) kinetic isotope effects (KIE), and (2) equilibrium isotope fractionation. Kinetic isotope effects occur in reactions far from chemical equilibrium: Lighter isotopes tend to have weaker chemical bonds that are therefore more easily broken during the dissociation of reactants, which causes relative enrichment of light isotopes in the product phase. Kinetic isotope effects can also result from diffusion processes, because lighter isotopes also have greater diffusive rates (Clark and Fritz, 1997; Rodushkin et al., 2004). When a reaction reaches chemical equilibrium (i.e. equal forward and backward reaction rates), the isotopes of an element are redistributed as a function of their respective thermodynamic stability in product and reactant phases. This redistribution is behind equilibrium isotope fractionation. Schauble (2004) summarized the general conditions controlling equilibrium isotope fractionation, namely: 1. Equilibrium isotope fractionation is temperature-dependent and increases roughly in proportion to 1/T or 1/T2. 2. Equilibrium isotopic fractionations are larger for lighter elements, which have greater relative mass differences between isotopes.  11   3. At chemical equilibrium, heavy isotopes tend to be enriched in the substance which has the "stiffest" bonds. Stiff bonding environments tent to be shorter in length and have lower coordination numbers. For redox-sensitive elements, stiffer bonds usually also involve the oxidized species. 1.4 Studies of metal stable-isotope fractionation applied to mine waste The development of MC-ICP-MS has initiated research into the application metal stable-isotope analyses at mine sites over the last two decades. Early efforts focused on identifying the stable isotopic signatures of various sulfide ore minerals such as molybdenite, sphalerite, and chalcopyrite for mineral exploration (e.g. Hannah et al., 2007; Kelley et al., 2009; Mathur et al., 2010, 2009). A few pioneering studies have characterized metal stable-isotope fractionation associated with weathering of sulfidic mine waste for a handful of chemical elements. Copper isotopic fractionation in mining environments results from oxidation-reduction reactions involving Cu(I)/Cu(II), such as oxidative dissolution of Cu(I)-bearing covellite (CuS) and chalcopyrite (CuFeS2) (Dótor-Almazán et al., 2017; Ehrlich et al., 2004; Fernandez and Borrok, 2009; Pérez Rodríguez et al., 2013; Song et al., 2016). Mine-impacted waters are enriched in heavy Cu isotopes relative to source minerals (Kimball et al., 2009). Copper adsorption onto Fe-(oxyhydr)oxides and precipitation of Cu-hydroxy-carbonate minerals—common processes in sulfidic mine waste—also causes isotopic fractionation (Balistrieri et al., 2008; Maréchal and Sheppard, 2002). Mercury isotopic signatures in sediments contaminated by mining of cinnabar  12   (HgS) enabled Foucher et al. (2009) to trace Hg sources in a mine-impacted fluvial setting in Slovenia. Large Hg isotopic fractionation occurs during the chemical processes involved in ore extraction (Stetson et al., 2009). Iron is redox- and biologically active, and so its isotopes also fractionate in mining environments due to biological and chemical oxidative dissolution of ore minerals and reactions of Fe(II)/Fe(III)+ speciation and precipitation (e.g. Beard et al., 1999; Egal et al., 2008; Pérez Rodríguez et al., 2013). Characterization has also begun of the isotopic composition of less-studied elements in mine-impacted waters such as antimony (Resongles et al., 2015).  These studies collectively show that there are multiple processes that can cause isotopic fractionation in mine waste, including adsorption, mineral dissolution-precipitation, aqueous complexation, changes in redox state, and biological reactions. Prior to the work presented in this thesis, no studies had examined Mo stable-isotope fractionation in mine drainage. Past studies of Zn stable isotopes in mine drainage found only limited Zn isotopic fractionation (Aranda et al., 2012; Borrok et al., 2008; Matthies et al., 2014; see also section 1.6). Both Mo and Zn have widespread industrial applications and serve as micronutrients—but can be toxic at elevated environmental concentrations (Barceloux, 1999; Canadian Council of Ministers of the Environment, 1999; Valko et al., 2005). They also both undergo a variety of mineral dissolution-precipitation and adsorption reactions in mine waste, which have the potential to cause isotopic fractionation. The objective of this dissertation is therefore to investigate the magnitude of Mo and Zn isotopic fractionation in mine-drainage settings and how it can be used to indicate  13   chemical attenuation of these elements. Both field and laboratory-based approaches were taken. A greater focus was given to Mo, because it is less well studied in mine waste and its complex geochemical behavior makes it well suited to isotopic study. The following section reviews the elemental and isotopic geochemistry of Mo and Zn to provide the background information relevant to the study of these elements. 1.5 Molybdenum elemental and isotopic geochemistry Molybdenum is used industrially in alloys, lubricants, printing inks, rubbers, fertilizers, and a variety of other compounds (CCME, 1999a). Its aqueous and isotopic geochemistry was recently reviewed by Kendall et al. (2017) and Smedley and Kinniburgh (2017). It is redox-sensitive, and occurs predominantly with oxidation states of Mo(IV), Mo(V), and Mo(VI) under Earth-surface conditions (Kendall et al., 2017). In the context of mine drainage the dominant source of Mo is the ore mineral molybdenite (MoS2). Silicates also contain Mo at trace levels (Greaney et al., 2016). Sulfide minerals of igneous origin (e.g. pyrite) are not substantially enriched in Mo (Greaney et al., 2016; Pichler and Mozaffari, 2015), but sedimentary pyrite can be (Smedley and Kinniburgh, 2017). Under most natural and oxic aqueous environments, Mo is dominantly present as the molybdate oxyanion MoVIO42-. At pH of 6 or lower, its protonated forms HMoO4- and H2MoO4 become increasingly important at the expense of MoO42- (Smedley and Kinniburgh, 2017) (Figure 1.5).  14   Adsorption is an important control of Mo transport in the environment, and is strongly pH-dependent. Under alkaline-pH conditions, Mo adsorbs minimally, whereas in moderately acidic solutions it adsorbs strongly to a variety of solid surfaces, including (oxyhydr)oxides of Fe, Mn, and Al; clays; pyrite; and organic matter (Bostick et al., 2003; Goldberg et al., 1996; Gustafsson, 2003; Wichard et al., 2009; Xu et al., 2006). In strongly acidic conditions (pH < 3), Mo does not adsorb (Frascoli and Hudson-Edwards, 2018). In addition, adsorption on calcite is relatively weak compared to other mineral surfaces (Goldberg et al., 1996). The precipitation of molybdate minerals can control its solubility over a wide pH range (Torres et al., 2014). These minerals include: powellite (CaMoO4); wulfenite (PbMoO4); nickel(II) molybdate [NiMoO4]; FeMoO4; and ferrimolybdite [Fe2(MoO42-)3] (Blanchard et al., 2015; Conlan et al., 2012; Essilfie-Dughan et al., 2011; Hayes et al., 2014; Hirsche et al., 2017). At pH < 6 and high Mo concentrations (> 96 mg/L), polymolybdate molecules such as Mo7O246- or Mo8O264- form via MoO42- polymerization (Torres et al., 2014), although such concentrations are rarely reached, even in the most Mo-enriched water sources in the environment (Smedley and Kinniburgh, 2017). Under reducing and sulfidic conditions, Mo is insoluble: MoO42- undergoes stepwise conversion to MoS42- via a series of intermediate (oxo)thiomolybdate species and co-precipitates with Fe sulfides and organic matter (Erickson and Helz, 2000; Helz et al., 2011, 1996; Vorlicek et al., 2004; Wang et al., 2011) (Figure 1.5). Eventual reduction of Mo(VI) to Mo(V) and  15   Mo(IV) in sulfidic sediments is possible (Wang et al., 2011) but direct precipitation of MoIVS2 does not occur under near-surface temperature conditions as a result of kinetic inhibitions (Vorlicek et al., 2004).  Figure 1.5. Molybdenum Eh-pH diagram for a Mo–H2O–H2S system. Concentrations are 10-6 M Mo and 10-4 M S, after Kendall et al. (2017). Note that the intermediate MoVO2+ species identified by Wang et al. (2011) is not shown. Intermediate MoOXS4-x2- species are metastable and not shown here.     16   Applications of Mo stable isotopes in environmental earth sciences are growing rapidly since the first such studies in the early 2000s (Barling et al., 2001; Siebert et al., 2001). There are seven stable isotopes of Mo: 92Mo (14.7 %), 94Mo (9.2 %), 95Mo (15.9 %), 96Mo (16.7 %), 97Mo (9.6 %), 98Mo (24.3 %), and 100Mo (9.7 %). Isotopic fractionations are reported in per mil units (‰) relative to a standard using the δ98Mo notation: δ98Mosample (‰) = ((Mo98Mo95)sample(Mo98Mo95)standard− 1) × 1000  (1.1) The most commonly used Mo isotope standard is the synthetic NIST-SRM-3134, whose δ98Mo is defined as +0.25 ‰ (Nägler et al., 2014). The δ98Mo values measured in natural reservoirs to date vary by a few per mil (Figure 1.6). Equilibrium isotopic fractionation occurs during Mo adsorption onto a variety of mineral surfaces and organic matter, which enriches aqueous solutions in heavy Mo isotopes (Barling and Anbar, 2004; Goldberg et al., 2009; King et al., 2017; Wasylenki et al., 2008). The magnitude of isotopic fractionation during adsorption is at least partially controlled by a coordination change from tetrahedral MoO42-(aq) to octahedral Mo on the adsorption surface (Arai, 2010; Gustafsson and Tiberg, 2015; Kashiwabara et al., 2011; Wasylenki et al., 2011). There also is evidence for preferential removal of light Mo isotopes during conversion of molybdate to (oxo)thiomolybdate and subsequent Mo precipitation under sulfidic aqueous conditions (Bura-Nakić et al., 2018; Dahl et al., 2010; Nägler et al., 2011).   17    Figure 1.6. Molybdenum stable-isotope compositions of various Earth materials. Aqueous reservoirs are in blue and solid-phase materials are in beige. Data are reported relative to δ98Mo of NIST-SRM-3134, which is by convention equal to +0.25 ‰. Aqueous Mo shows enrichment in heavy Mo isotopes relative to igneous and sulfide minerals, while ferro-manganese crusts are enriched in isotopically light Mo isotopes. Modified from Smedley and Kinniburgh (2017). Note that values reported for mine drainage are from Skierszkan et al. (2016; Chapter 3). To date, applications of Mo stable-isotope analyses have overwhelmingly focused on unravelling the history of Earth’s oxygenation, because the redox-sensitive mobility of Mo provides evidence for the appearance of oxygen in ancient sedimentary records (Kendall et al., 2017). A handful of studies have also examined isotopic fractionation of Mo during continental weathering (Archer and Vance, 2008; King et al., 2015; Pearce et al., 2010; Siebert et al., 2015) to constrain the global Mo cycle. Another branch of the literature has sought to apply variations of δ98Mo in ore bodies to understand mineralization processes (Breillat et al., 2016; Greber et al., 2014; Hannah et al., 2007; Mathur et al., 2010; Yao et al., 2016). However, relatively few studies have used Mo isotopes with a focus on environmental contamination (Kendall et al., 2017) and no prior studies have focused on Mo isotopic variations in mine-impacted waters. There have  18   also been no studies of Mo isotopic fractionation during the precipitation of molybdate minerals—a process known to attenuate Mo in mine waste (Bissonnette et al., 2016; Blanchard et al., 2015; Conlan et al., 2012; Essilfie-Dughan et al., 2011, 2010; Hayes et al., 2014). Overall, the widespread demonstration of Mo isotopic fractionation at Earth’s surface suggests that Mo isotopes may be applicable to trace geochemical processes relevant to mine waste. 1.6 Zinc elemental and isotopic geochemistry Zinc is widely used in galvanizing processes; in brass and bronze used for plumbing; heating, and cooling systems; as an additive in plastics and rubber products; and in multiple other industrial applications (CCME, 1999b). In contrast to Mo, Zn is redox-inactive at Earth's surface and is dominantly present as a cation (Zn2+). In mine waste, the principal source of Zn is sulfide minerals, namely sphalerite (ZnS). Because of its cationic form in aqueous solution, the pH-dependency of Zn adsorption is opposite that of MoO42-: Zn adsorption is strongest under alkaline pH conditions and decreases in acidic conditions (Stumm and Morgan, 1996). Zinc can also be removed from aqueous solutions via the precipitation of a variety of secondary minerals, including: Zn(OH)2; carbonates [e.g. hydrozincite – Zn5(OH)5(CO3)2]; silicates [e.g. hemimorphite – Zn4Si2O7(OH)2•H2O]; sulfates [e.g. zincosite – ZnSO4] and amorphous hydroxycarbonate and hydroxysulfates (Iavazzo et al., 2012; Jacquat et al., 2008; Laurenzi, 2016; Wanty et al., 2013a; 2013b).   19   Zinc comprises five stable isotopes: 64Zn (49.2 %), 66Zn (27.8 %), 67Zn (4.0 %), 68Zn (18.4 %) and 70Zn (0.6 %). Zinc isotope ratios are noted using the δ66Zn notation, analogous to δ98Mo: δ66Znsample (‰) = ((Zn66Zn64)sample(Zn66Zn64)standard− 1) × 1000 (1.2) δ66Zn data are most-often normalized to the isotope standards JMC 3-0749C (also called JMC-Lyon) and IRMM-3702. JMC-Lyon has been proposed as the international standard for reporting δ66Zn (Moynier et al., 2017). The stable isotope geochemistry of Zn has been periodically reviewed since δ66Zn measurements were first conducted with MC-ICP-MS in the late 1990s (Cloquet et al., 2008; Maréchal et al., 1999; Moynier et al., 2017; Wiederhold, 2015), and the range in δ66Zn reported in the literature to date spans approximately 2.5 ‰ (Figure 1.7). Zinc isotopes have been studied in mining environments in a handful of previous works. A 2-year monitoring study of an experimental waste-rock pile in northern Canada found relatively little variation in δ66Zn, which suggested limited fractionation from Zn attenuation processes (Matthies et al., 2014b). A subsequent laboratory study of oxidative weathering of sulfidic mine tailings found modest isotopic fractionation which was explained by release of Zn from isotopically distinct mineral pools contained within the tailings (Matthies et al., 2014a). Other studies have found limited Zn isotope fractionation in surface waters downstream of  20   historical acid-rock drainage (ARD) sites, leading to the proposition that Zn stable-isotope ratios were more likely to be tracers of Zn sources rather than attenuation processes, due to its conservative isotopic and geochemical behavior under acidic conditions (Aranda et al., 2012; Borrok et al., 2008; Fernandez and Borrok, 2009). However, precipitation of Zn-bearing sulfide, carbonate, and silicate minerals at near-surface temperature conditions can cause Zn isotopic fractionation (Jamieson-Hanes et al., 2017; Veeramani et al., 2015; Wanty et al., 2013; Wanty et al., 2013). Zinc adsorption onto mineral surfaces also causes Zn isotopic fractionation, but in the opposite direction relative to Mo isotopes: most studies to date report enrichments in heavy Zn isotopes on sorption surfaces (Balistrieri et al., 2008; Bryan et al., 2015; Dong and Wasylenki, 2016; Guinoiseau et al., 2016; Juillot et al., 2008; Kafantaris and Borrok, 2014; Pokrovsky et al., 2005; Nelson et al., 2017). This behavior is generally explained by a change in coordination from octahedral aqueous Zn to tetrahedral adsorbed Zn, which is opposite to that of Mo (Juillot et al., 2008; Wasylenki et al., 2011). Kinetic isotope fractionation is possible during Zn adsorption in experiments lasting less than 48 hours; longer timescales are sufficient to reach isotopic equilibrium (Bryan et al., 2015).   21    Figure 1.7. Zinc stable-isotope compositions for natural materials found at the surface of the Earth. Thicker lines represent the average and 1 SD for each type of sample; thinner lines represent the total range of values measured. For oceans the average shown is for depths > 600 m. For rivers, the average shown is weighted by total discharge. Modified from Moynier et al. (2017).    22   1.7 Analytical considerations for metal stable-isotope analyses by MC-ICP-MS Application of metal stable-isotope data to geoscience problems is only possible with highly precise and accurate data, which are usually collected by MC-ICP-MS. There are two important analytical challenges to overcome for MC-ICP-MS analyses. First, the element(s) of interest—present at parts-per-million levels or less in most samples—must be chemically separated from the other sample matrix elements. This extraction is typically accomplished using ion-exchange chromatography, and is necessary because other matrix constituents cause interferences that render inaccurate isotope-ratio measurements (e.g. Barling and Weis, 2012, 2008; Pietruszka and Reznik, 2008; Shiel et al., 2009). Second, the isotopic ratios in the samples are skewed by instrumental fractionation occurring in the MC-ICP-MS and also during any laboratory manipulations that yield incomplete sample recovery, including sample digestion and ion-exchange chromatography. This laboratory fractionation can be orders of magnitude larger than the natural fractionation of samples which is of interest. The instrumental fractionation, also known as mass bias, can be corrected by two approaches: (1) standard-sample bracketing (SSB) and/or (2) the use of a “double spike”. Standard-sample bracketing involves measurement of standards before and after each sample to determine mass bias. Mass-bias corrections are then applied to the sample isotope ratios by linear interpolation. A variation of the SSB technique involves the addition (“doping”) of another element of similar mass as the analyte to the standard and sample solutions. Variations in the  23   measured isotopic ratio of the dopant element are used as an additional monitor of mass bias in combination with SSB, for example using 65Cu/63Cu to correct 66Zn/64Zn (Maréchal et al., 1999).  Standard-sample bracketing requires quantitative analyte recovery during laboratory handling of samples (e.g. ion-exchange chromatography and sample digestion) to avoid artefactual laboratory fractionation. Furthermore, it relies on the assumption that mass bias impacts bracketing standards and dopant elements in the same way as samples, and that it varies linearly between sample and standard analyses within a mass spectrometer run. Both of these assumptions can be compromised by a variety of spectral and non-spectral matrix effects (Pietruszka and Reznik, 2008).  The “double-spike” approach for mass bias correction involves addition of a solution containing two artificially enriched isotopes of the element of interest to the samples prior to ion-exchange chromatography and isotopic analysis (Compston and Oversby, 1969; Eugster et al., 1969). The correction is achieved by comparing the deviation in the isotope ratios of the double spike relative to calibrated values (Rudge et al., 2009). Because the double spike inevitably contains isotopes that are also found in the samples, a mathematical inversion is required to determine the proportion of double-spike and sample isotopes prior to applying mass-bias corrections (Rudge et al., 2009). The double-spike mass-bias correction is truly internal, being applied to the same element as the analyte and also being applied during the same analysis time; it therefore presents considerable advantages over SSB. Furthermore, addition of the double  24   spike prior to ion-exchange chromatography circumvents the need for quantitative sample recovery because it provides adequate corrections from laboratory fractionation (Siebert et al., 2001).  Given that the analytical challenges associated to ion-exchange chromatography and mass-bias corrections are not trivial, they are discussed in more detail in Chapter 2, which describes the implementation of a Mo double spike for MC-ICP-MS analyses of δ98Mo. 1.8 Thesis objectives and organization 1.8.1 Thesis objectives The fundamental objective of this thesis is to evaluate the use of isotopic analyses of Mo and Zn to evaluate attenuation of these metals in mine waste environments. To enable this study, a first goal was to implement a suitable isotopic analytical protocol, which included the purification of Mo and Zn from environmental samples by ion-exchange chromatography (as adapted from previous workers: Pearce et al., 2009; Shiel, 2010) and the development of a Mo double-spike MC-ICP-MS method (Chapter 2). A second goal was to characterize the magnitude of Mo and Zn isotopic fractionation among solid mine waste, ore minerals, and mine drainage at field sites (Chapters 3, 4, and 6). This characterization set the stage for further investigation into the processes driving this fractionation. A third goal was to determine the magnitude of Mo isotopic fractionation during specific reactions known to attenuate Mo in mine waste, in  25   particular precipitation of powellite and wulfenite (Chapter 5) and Mo adsorption onto Fe-(oxyhydr)oxides (Chapters 4 and 6). Finally, current understanding of the processes causing Mo and Zn isotopic fractionation was applied to link variations in δ98Mo and δ66Zn and other geochemical and mineralogical data to processes of geochemical attenuation of these metals under field conditions at two mine sites (Chapters 3, 4, and 6). 1.8.2 Selection of field sites Field investigations conducted in Chapters 3, 4, and 6 involved two mine sites, which allows for broader conclusions to be drawn from the results of this work. The first field site, presented in more detail in Chapter 3, was the Antamina mine, Peru. This site was selected because: (1) the waste rock at Antamina includes substantial lithological heterogeneity and therefore results in highly variable drainage pH and metal concentrations, and creates an ideal opportunity to characterize metal stable-isotopic fractionation in mine drainage over a wide range of geochemical conditions at a single site; (2) both Mo and Zn are contaminants of potential environmental concern at Antamina; and (3) it is the locus of a multi-year study of waste-rock drainage chemistry and hydrology (Beckie et al., 2011) including well-characterized small- to intermediate-scale waste-rock “field barrel” and “constructed pile” experiments (Figure C.1) that provide well-constrained lithological and hydrogeological controls on drainage chemistry. The second field site was the Thompson Creek Mine (TCM), in Idaho, USA. The TCM was chosen because: (1) its Mo-bearing waste rock and mine tailings management  26   facilities exhibit large contrasts in pH (alkaline vs. acidic) and redox conditions (oxic vs. anoxic), both of which are key parameters in controlling Mo mobility; and (2) Mo is a contaminant of potential concern at the site (but not Zn; hence Zn isotopes were not analyzed at the TCM).  1.8.3 Thesis organization This thesis contains five research chapters (Chapters 2 to 6), each describing different aspects of the application of metal isotopic analyses as environmental tracers in mine waste. The inter-relation among these chapters is shown in (Figure 1.8). These research chapters are bounded by the present introduction, and a conclusion chapter (Chapter 7) that presents the integrated knowledge gained in this thesis. Some repetition is present between the chapters because they are written in manuscript format for submission to peer-reviewed scientific journals.  27    Figure 1.8. Flow-chart linking the different research chapters composing this thesis. Chapter 2 provides a description of the analytical protocol that was developed and used in this thesis for Mo stable-isotope analysis by double-spike MC-ICP-MS. This chapter provided the fundamental methodology for isotopic analyses that was used throughout the thesis and constitutes a detailed guide for double-spike analysis implementation for other users. The method includes three main steps: (1) extraction of Mo from sample matrices using ion-exchange chromatography; (2) the use of a 97Mo–100Mo double spike to correct for isotopic fractionation occurring during laboratory and instrumental manipulations; and (3) isotopic analysis using MC- 28   ICP-MS. This analytical protocol was validated by reproduction of δ98Mo values for a variety geological reference materials and Mo isotope standards (Appendix A).  The ensuing chapters include characterization of the extent of Mo and Zn isotopic fractionation in mine waste and the interpretation of stable-isotope, geochemical, and mineralogical data to track processes of metal attenuation. Beginning in Chapter 3, the occurrence and extent of Mo and Zn isotopic fractionation was documented in drainage and waste rock from the field-barrel and constructed-pile experiments at the Antamina mine. Adsorption and secondary mineral precipitation reactions were invoked as the probable causes for this fractionation on the basis of geochemical and isotopic analyses. The observation of isotopic fractionation for Mo and Zn in mine drainage, likely resulting from attenuation processes, opened the door to further study in this thesis. However, a lack of quantitative knowledge on Mo isotopic fractionation during precipitation of the minerals powellite and wulfenite—probable controls on Mo transport in Antamina waste-rock drainage (Conlan et al., 2012)—limited our ability to interpret δ98Mo data in mine waste, providing the motivation for studying isotope fractionation related to the formation of these minerals in Chapter 5. Chapter 5 therefore comprises a series of laboratory experiments quantifying Mo isotopic fractionation factors during the precipitation of the molybdate minerals powellite (CaMoO4) and wulfenite (PbMoO4); both of these minerals preferentially scavenge light Mo isotopes from aqueous solution. The solution chemistry used in the experiments of Chapter 5 was varied to  29   mimic conditions more similar to mine drainage, and therefore help validate the application of the fractionation factors that were obtained in the lab to field conditions. Meanwhile, Chapters 4 and 6 further evaluated the geochemical attenuation of Mo by analysis of its isotopes in mine drainage at the field scale, i.e. full-scale waste rock and tailings storage facilities, at the Thompson Creek Mine. These chapters utilize knowledge gained from previous chapters, in particular the new constraints on (1) the role of powellite precipitation (Chapter 5) and adsorption onto Fe-(oxyhydr)oxides (Chapter 3) in attenuating Mo and fractionating its isotopes in mine waste. Under oxic conditions, adsorption of Mo onto Fe-(oxyhydr)oxides was found to be the dominant attenuation mechanism, as revealed by geochemical analyses of secondary coatings on weathered waste-rock, Fe-(oxyhydr)oxide-rich sediment, and mine drainage samples. Precise molecular mechanisms for this sorption process were determined with X-ray absorption spectroscopy analyses in Chapter 6. In contrast, under the anoxic conditions found in the Bruno Creek Mine Tailings Management Facility (BCTMF) at the TCM, Mo stable-isotope analyses indicated that Mo is more mobile because Fe-(oxyhydr)oxides are reductively dissolved and therefore no longer available to adsorb Mo.  Finally, Chapter 7 summarizes the general conclusions from this thesis, including: (1) contributions to the application of metal stable-isotope analyses as a new tool now available to constrain metal transport and attenuation in mine waste; and (2) further insight on the geochemical mobility of Zn and especially Mo at mine sites. Limitations and future research  30   directions related to this study are also provided in this chapter. This discussion includes the suggestion that metal stable-isotope analyses can be included in the growing arsenal of techniques available to geochemists in the study of the mobility of many multi-isotopic chemical elements in the environment (in addition to Mo and Zn) now that that analytical hurdles are being overcome by MC-ICP-MS methods.  31   CHAPTER 2. A PRACTICAL GUIDE FOR THE DESIGN AND IMPLEMENTATION OF THE DOUBLE-SPIKE TECHNIQUE FOR PRECISE DETERMINATION OF MOLYBDENUM ISOTOPE COMPOSITIONS OF ENVIRONMENTAL SAMPLES 2.1 Introduction Metal stable isotopes have proven to be extremely useful in elucidating the environmental fate of many elements since the development of multi-collector inductively coupled-plasma mass spectrometry (MC-ICP-MS) has enabled high-precision isotope ratio measurements. However, these analyses require analytical precision and accuracy at the sub per-mil level, which can be challenging to achieve because natural isotope fractionation is easily masked by instrumental mass fractionation (or mass bias, also noted as ß) that is orders of magnitude larger than natural fractionation.  In order to correct for mass bias, standard-sample bracketing (SSB), external normalization (also called element-doping) and double-spike procedures have been applied (e.g. Albarède and Beard, 2004; Anbar et al., 2001; Pietruszka et al., 2006; Siebert et al., 2001; Télouk et al., 1999). Of these three methods, the double-spike approach provides the most robust correction for instrumental and laboratory mass fractionation (Siebert et al., 2001) and protocols are now in place for the isotopic analysis of several elements such as Ba, Ca, Cd, Hg, Ge, Mo, Ni, Pb, Pt, Se, U and Zn (e.g. Bermin et al., 2006; Bonnand et al., 2011; Compston and Oversby, 1969; Creech et al., 2013; Eugster et al., 1969; Heumann et al., 1998; Makishima et al., 2007; Mead and Johnson, 2010; Miyazaki et al., 2014; Ripperger and Rehkämper, 2007; Russell et al.,  32   1978; Siebert et al., 2006; Zhu et al., 2008). However, a more widespread adoption of double-spike protocols has been hindered by perceived difficulties in proper double-spike calibration, mathematical inversion (Rudge et al., 2009), and fear of memory effects (Albarède et al., 2004). While the mathematical challenges of the double-spike approach are now readily overcome (Rudge et al., 2009), successful double-spike calibration remains a challenging task, in part due to a paucity of detailed explanations of double-spike calibration procedures that have been established in various laboratories.  Therefore, in order to facilitate the adoption of the double-spike method, our objective here is to provide a detailed account of the development and implementation of a Mo double-spike protocol which yields excellent precision and accuracy. In this light, a description of the double-spike calibration is presented in greater detail than has been presented in the double-spike literature (e.g. Bermin et al., 2006; Rudge et al., 2009; Siebert et al., 2001). In addition to the details contained in this text, a graphical summary of the technique is provided for ease of use ( Figure 2.1. and Figure 2.2.). We also showcase the robustness of the double-spike method to laboratory and instrumental fractionation in comparison to SSB and external normalization, and explore some of the limitations of the double spike when samples are spiked at sub-optimal spike-to-sample ratios. Finally, the accuracy and precision of the method are evaluated by the analysis of Mo isotope standards and reference materials. It should be noted that the calibration approach described here invokes an exponential law normalization step to a dopant (Pd) to  33   determine a first isotopic ratio of the double spike and the standard. Two recent studies present Mo isotope calibration methods which make it possible to circumvent assumptions related to the element-doping step (Malinovsky et al., 2014; Mayer and Wieser, 2014). Nonetheless, the approach outlined here provides accurate and precise isotope measurements generally better than the <0.02 ‰ amu-1 level and remains a viable and straightforward approach for Mo and for analogous elements where absolute isotope compositions are not available. Several recent publications have recommended that the NIST-SRM-3134 be adopted as an international Mo isotope standard (Goldberg et al., 2013; Greber et al., 2012; Mayer and Wieser, 2014; Nägler et al., 2014; Wen et al., 2010). Early Mo isotope studies have shown that Mo isotopic fractionation occurs as a result of geochemical interactions such as adsorption, biological uptake or mineral dissolution and precipitation, making Mo isotope a promising tool for geological and environmental applications (Anbar et al., 2001; Goldberg et al., 2009; Greber et al., 2014; Hannah et al., 2007; Lane et al., 2013; Siebert et al., 2003; Voegelin et al., 2009; Wieser et al., 2007). Thus far the terrestrial span of δ98/95Mo is 6.4 ‰ (Goldberg et al., 2013).    34    Figure 2.1. Flow chart for the design, calibration, and implementation of a double spike for stable-isotope analysis by MC-ICP-MS. Continued in Figure 2.2. Boxes with dashed lines indicate examples for a Mo double spike calibrated using Pd-doping. α refers to the natural fractionation factor.   35     Figure 2.2. Flow chart for the design, calibration, and implementation of a double spike for stable-isotope analysis by MC-ICP-MS. Continued from Figure 2.1.    36   Molybdenum has seven stable isotopes which vary in abundance from ~ 9 % to ~ 24 %. Molybdenum isotope compositions are reported as per mil deviations of a given sample isotope ratios relative to a standard's isotope ratios, e.g.: δ98/95Mosample (‰) = ((Mo98Mo95)sample(Mo98Mo95)standard− 1) × 1000    (2.1) 2.2 Methods 2.2.1 Considerations for the double spike design The multitude of Mo stable isotopes (92Mo, 94Mo, 95Mo, 96Mo, 97Mo, 98Mo, and 100Mo) presents several options for the selection of spike isotopes. The best isotopes to select for a double spike should adhere to the following criteria: (1) They should have low relative abundances in nature, which reduces the mathematical error when deconvolving the spike-sample mixture; and (2), they should be free from isobaric interferences, unless these can entirely removed by ion-exchange chromatography. In addition, the error occurring during the double-spike inversion is sensitive to which isotopes are selected to make up the double spike (Rudge et al., 2009).  The double-spike approach is limited to elements with at least four isotopes, such that there are two available analyte isotopes and two spike isotopes. For Mo, most protocols to date use a 97Mo–100Mo double spike (e.g. Dahl et al., 2010b; Goldberg et al., 2009; Pearce et al.,  37   2009; Siebert et al., 2001; Wen et al., 2010) because these isotopes have low natural abundances (9.55 and 9.63 %, respectively); and 97Mo is free from isobaric interferences while the only interference on 100Mo is 100Ru. Ruthenium is typically found in much lower concentrations than Mo in most environmental samples and is removed by ion-exchange chromatography prior to sample analysis (Siebert et al., 2001).  In addition to selection of the spike isotopes, the double-spike deconvolution is sensitive to the proportion of individual spike isotopes in the double spike. The optimal proportion of each individual spike in the double spike can be simulated such that the deconvolution yields the minimal error in the natural fractionation factor (typically denoted "α") using an openly available MATLAB® code (Rudge et al., 2009). Assuming that pure individual isotope spikes are available, these simulations show that for the 97Mo–100Mo double spike a wide range in the proportion of 97Mo in the 97Mo–100Mo double spike from ~ 0.2 to 0.8 minimizes the error in α with an absolute minimum occurring at a proportion of 0.49. This corresponds to a 100Mo-97Mo ratio of 1.6, although most 97Mo–100Mo protocols developed to date have used a 100Mo-97Mo ratio closer to 1 (e.g. Archer, 2007; Pearce et al., 2009; Siebert et al., 2001).  2.2.2 Double spike and standard preparation and calibration 97Mo and 100Mo spikes (94.19 % and 92.16 % purity, respectively) were acquired from Oak Ridge National Laboratories (Oak Ridge, Tennessee, USA). Each individual spike, containing 1 mg of Mo, was dissolved in sub-boiled concentrated HNO3 in Savillex® PFA  38   beakers, dried, and re-constituted in ~ 10 g of 0.05 M HF + 0.05 M HNO3. Aliquots of each individual spike were then taken out and mixed together to form the double spike. This mixture was subsequently diluted to form a 250 mL double-spike solution in a 2 % HNO3 + 0.1 % HF matrix with a 100Mo/97Mo ratio of ~ 1, similar to previous 100Mo–97Mo double-spike protocols (Archer, 2007; Pearce et al., 2009; Siebert et al., 2001). The in-house isotope standard used throughout this thesis—henceforth named “Mo(UBC)”—was a pure Mo ICP standard solution (Specpure lot 214373D). Prior to the calibration session, an aliquot of Mo(UBC) stock solution was evaporated to dryness in a Savillex® beaker, and re-constituted in the same 2 % HNO3 + 0.1 % HF solvent as that of the double spike. The δ98/95Mo of all subsequent sample analyses was defined relative to this standard as per Equation (2.1). For inter-laboratory comparison of measurements, δ-values were then converted to δ98/95MoNIST-SRM-3134 = 0 ‰1 using Equation 2.2:  δ98/95MoSample-NIST =  δ98/95MoSample-Mo(UBC) + δ98/95MoMo(UBC)-NIST x 10-3(δ98/95MoSample-Mo(UBC) + δ98/95MoMo(UBC)-NIST)  (2.2) The double spike and standard calibration was based on the approach of Siebert et al., (2001) with some modifications. In the absence of standards with a known absolute isotope                                                  1 The δ98/95Mo of all data reported in all subsequent chapters (Chapters 3 to 7) in this thesis use the convention of NIST-SRM-3134 = +0.25 ‰, after Nägler et al. (2014).  39   composition to use in the calibration, one approach is to calibrate the double spike and standard relative to another element used as a dopant. In this element-doping approach, the calibrated double spike and standard compositions are dependent upon an assumed isotope ratio of the dopant, but accurate isotope measurements of samples remain possible so long as the double spike and standard are well-calibrated relative to each other (Siebert et al., 2001). Here, Pd was used as a dopant rather than other elements (Ru, Zr; c.f. Anbar et al., 2001) to eliminate the possibility of introducing isobaric interferences (e.g. 92Zr, 94Zr, 96Zr, 96Ru, 98Ru, 100Ru). A Pd dopant was prepared by re-constituting an aliquot of a pure Pd solution (Specpure Pd ICP standard, lot 225840B) into a 500 ppm solution with the same matrix as the double spike and standard (2 % HNO3 + 0.1 % HF).  An assumed 108Pd/105Pd of 1.184953 (de Laeter et al., 2003) was used as a first step to correct for instrumental mass bias and define a “true” 100Mo/97Mo in the spike and standard using the exponential law (Russell et al., 1978). The other Mo isotope ratios of the spike and standard were then calibrated relative to this “true” 100Mo/97Mo. This choice of Pd isotope ratio used as a mass bias monitor differed from the method presented by Siebert et al. (2001) in which the 104Pd/102Pd was used to calibrate 100Mo/97Mo in the double spike and standard. Despite the large spread in masses from 97Mo to 108Pd that required a dynamic mode analysis, the accuracy of the Pd-normalization scheme using 108Pd/105Pd was improved compared to using 104Pd/102Pd. This improvement was demonstrated by better mass bias correlations between measured 100Mo/97Mo ratios and 108Pd/105Pd ratios than when using the 104Pd/102Pd ratios (Figure 2.3.). The enhanced  40   mass-bias correlations seen with 108Pd/105Pd are thought to result from better matching of Pd and Mo isotope intensities due to the similarities in the natural 108Pd/105Pd and 100Mo/97Mo in the double spike and the standard, which diminished non-spectral matrix effects. However, it is also possible that the mass bias correlation of 104Pd/102Pd and 100Mo/97Mo could improve at higher Pd concentrations to enhance the signal of 102Pd, which was only ~ 130 mV in our calibration session, although this effect was not examined in this study.   Figure 2.3. ln-ln plots of measured Pd and Mo isotope ratios during the Pd-doped Mo(UBC) standard calibration. Improved correlation of mass bias of Pd and Mo isotopes were obtained when using 108Pd/105Pd (left) compared to 104Pd/102Pd (right). Similar relationships were present in the Pd-doped double-spike analyses (not shown). Measurements were made over a single analytical session. Error bars indicate ± 2 SE and are usually smaller than symbol size. The Pd-doped double spike and standard were analyzed in dynamic mode with isotope masses 96-108 monitored (Table 2.1), on a Nu Plasma MC-ICP-MS (Nu 021, Nu Instruments Ltd, Wrexham UK) housed at the Pacific Centre for Isotopic and Geochemical Research  41   (PCIGR, University of British Columbia, Vancouver, Canada). Solutions were introduced into the MC-ICP-MS using a wet nebulizing system (ESI PFA 200 µL) to avoid additional mass bias and memory effects occurring in desolvating units (e.g. DSN, Aridus). Three measurement batches of a Pd-doped standard solution were analyzed, followed by three batches of a Pd-doped double-spike solution. Pd isotope intensities were held constant in the Pd-doped double spike and Pd-doped standard analyses to ensure that the double spike and standard were well calibrated relative to one-another. In practice, this meant maintaining 105Pd and 108Pd intensities within 5% and maintaining the same 105Pd/97Mo (~ 1.2) in the doped double-spike and doped standard analyses.  Table 2.1. Faraday cup configuration for Pd-doped analysis of double spike and standard during calibration session. Underlined mass is used in peak-centering Faraday Cup H6 H5 H4 H3 H2 H1 Ax L1 L2 L3 L4 L5 Cycle 1 108Pd -- 106Pd -- 105Pd -- 104Pd -- -- 102Pd -- 100Mo Cycle 2 104Pd -- 102Pd -- -- -- 100Mo -- 99Ru 98Mo 97Mo 96Mo Each measurement batch consisted of 5 runs. A single run consisted of 30 measurements with a 5 s integration time and a 2 s magnet delay time. For this dynamic mode analysis, shorter integration times were used to reduce the effects of transient changes in mass bias when switching from measurements of 108Pd/105Pd in Cycle 1 to measurements of 100Mo/97Mo in Cycle 2. Prior to each run, a deflection of the electrostatic analyzer deviated the ion beam such that an instrumental electronic blank could be measured for 20 s and subtracted from the sample ion  42   beams. On-peak zero blank measurements were made for the Pd-doped analyses, although they were not found to improve the calibration results and thus not used in the data reduction.  Between each batch run, the instrument was washed for a minimum of 10 minutes in 2 % HNO3 + 0.1 % HF, and before switching from the standard to the double-spike solution, the instrument was washed for 1 hour to mitigate memory effects. Maintaining the same solvent for washout as for sample analysis (as opposed to using a stronger acid for washout) helps reduce plasma instabilities caused by changing solvents.  Once “true” 100Mo/97Mo values were defined for the double spike and standard, pure (Pd-free) double spike and standard solutions were analyzed and all other Mo isotope ratios were calibrated using the “true” 100Mo/97Mo to correct for mass bias via the exponential law. Molybdenum masses 92 to 100 were measured in static mode (Table 2.2.). Washout procedures were the same as in the Pd-doped analyses. In addition to the ESA blank correction, an on-peak zero blank correction was included to alleviate memory effects when switching from double spike to standard analyses due to their drastically different isotopic compositions. As with the Pd-doped analyses, the standard and double spike were analyzed in three batches consisting of 5 runs. One run consisted of 30 measurements with a 10 s integration time and a 2 s magnet settling time. In this analytical session, isotope intensities were maximized in order to improve signal intensities for 98Mo and 95Mo, which are present at trace levels in the 97Mo–100Mo double spike. In practice, this meant a ~ 7 V intensity on 100Mo (the most abundant isotope in the double  43   spike), and a 100 mV intensity on 95Mo. Because natural Mo isotope abundances are well distributed across the masses (9 to 24 %), high isotope intensities (> 2.5 V on 94Mo; the least abundant isotope in the standard) were easily achieved for the pure Mo standard calibration.  2.2.3 Validation of the double spike method’s accuracy In order to verify the accuracy of the calibration, a series of double-spiked standards were measured and the δ98/95Mo values for the standard, treated as an unknown, were calculated. An accurate calibration should yield a δ-value of 0 ‰ relative to Mo(UBC) for these standard measurements. These mixtures were first analyzed using a wet nebulizing sample introduction, within the same analytical session as the calibration. They were then analyzed in multiple separate sessions using a desolvator (DSN-100, Nu Instruments Ltd, UK), which is also used in the analyses of samples to benefit from the enhanced sensitivity. Aliquots of Mo isotope standard solutions (ICL-Mo, RochMo2, NIST-SRM-3134) obtained from other laboratories and with known isotopic compositions (Goldberg et al., 2013) were mixed with the double spike at molar proportions of 0.5 to 0.9 and analyzed to confirm the accuracy of the procedure.  Table 2.2. Faraday cup configuration for Mo analysis of pure double spike and standard during calibration session. Underlined mass is used in peak-centering Faraday  H6 H5 H4 H3 H2 H1 Ax L1 L2 L3 L4 L5 Cycle 1 100 -- 98 -- 97 -- 96 -- 95 94 -- 92 As a final confirmation of the protocol, natural samples requiring digestion and ion-exchange purification were analyzed. The natural samples selected for this study were seawater  44   from the Pacific Ocean and the USGS reference shale material SDO-1, both of which have well-characterized Mo isotopic compositions (Goldberg et al., 2013), and USGS reference basalt material BCR-2. Approximately 100 mg of powdered SDO-1 was digested on the hotplate in 2 mL of 3:1 concentrated HCl:HNO3 followed by 5 mL of 4:1 HF:HNO3 and then 2.5 mL of 4:1 HNO3:H2O2 (to break down organic molecules). The digested sample was then treated 4 times with a drop of concentrated HNO3 to further break down any organic matter prior to dissolution in 7 N HCl in preparation for ion-exchange chromatography. Similarly, approximately 100 mg of BCR-2 was digested on the hotplate in 1 mL of concentrated HNO3 and 9 mL of concentrated HF followed by 10 mL of 6 N HCl. For seawater analysis, a sample from the North Pacific open ocean was collected on February 2013 at approximately 5 m below sea level and about 1,100 km west of Vancouver Island, Canada (Station Papa, Line P Cruise, 50˚ N, 144˚ 15.9’ W). The sample was field-filtered through a 0.8-µm Supor® filter and acidified using sub-boiled HNO3. In the clean laboratory, double spike was added to a ~ 30 mL aliquot and dried down prior to re-dissolution for ion-exchange chromatography. The sample was then dissolved in ~ 10 mL > 18.2 MΩ water and converted to a 7 N HCl solution by addition of 12 N HCl. While the addition of HCl to the re-dissolved seawater led to the formation of some precipitates (presumably chloride salts), any Mo isotope fractionation that may have occurred was accounted for by adding the double spike prior to sample acidification in HCl, as is revealed by later δ98Mo analyses (section 2.3.1.2).  45   For these environmental samples, purification of the analyte by ion-exchange chromatography is necessary to reduce matrix effects. In this study, ion-exchange chromatography was completed using a modified version of the chemistries used by Shiel et al. (2009) and Pearce et al. (2009). The development of this ion-exchange protocol is described in Appendix B. Samples, typically containing approximately 1 µg of sample Mo, were loaded in 7 N HCl onto 2 mL of BioRad AG® MP-1M resin (100-200 mesh). The Mo fraction was collected in 10 to 14 mL of 8 N HF + 2 N HCl. Molybdenum recoveries for two separate replicates of the USGS reference material BCR-2 using this protocol yielded 94 % as measured using ICP-MS (Agilent 7700, Agilent Technologies, Santa Clara, CA, USA). In addition, this procedure provides quantitative separations of Zn and Cd in a single pass through the column (Shiel et al., 2013), allowing for rapid throughput of multiple isotope systems (Mo, Zn, Cd). Elution curves and further details regarding the anion-exchange chemistry protocol are available in Appendix B. Molybdenum column blanks through this thesis varied from 3 to 14 ng, averaging 5 ng (Table B.7). In order to reduce any impact from column blanks in isotope measurements, > 1 μg of sample Mo was processed through ion-exchange columns, except for seawater which has low Mo concentrations; in this case, ~ 300 ng of sample Mo were processed.  Molybdenum isotope analyses of all reference materials and standards following the calibration session were conducted on the Nu Plasma MC-ICP-MS at PCIGR. Typical instrument operating settings for these analyses are presented in Table 2.3. Samples were dissolved in 2 % HNO3 + 0.1 % HF and introduced to the plasma source using a DSN-100  46   desolvating nebulizer. Molybdenum isotope masses 92 to 100 were monitored in static mode using the Faraday cup configuration presented in Table 2.3. Each sample was measured in triplicate, with an individual analysis consisting of 30 measurement cycles using a 10 s integration time and a 2 s magnet settling time. Measurements of select samples and standards over multiple analytical sessions were made to further confirm the analytical reproducibility. Prior to each sample measurement, an on-peak zero blank and an ESA deflection blank were measured, and subtracted from the sample isotope intensities. One sample analysis in dry plasma consumed ~ 200 ng of sample-Mo.  Table 2.3. Typical instrument settings for Mo isotope analysis on the Nu MC-ICP-MS. RF power 1300 W Acceleration potential  6000 V Argon gas flow rate       Coolant 13 L min-1      Auxiliary 0.7 L min-1 Nebulizer pressure 44 psi Sensitivity for Mo ~ 100-180 V ppm-1 2.3 Results and Discussion 2.3.1 Verification of the accuracy of the double spike and standard calibration 2.3.1.1 Double-spiked Mo(UBC) standards Analyses of double-spiked Mo(UBC) standards with molar proportions of double spike in the spike-sample mixture that ranged from 0.5 to 0.9 in wet plasma gave an average δ98/95Mo value of -0.07 ± 0.07 ‰ (2 SD). However, it was possible to improve the accuracy of the  47   calibration result by iteratively modifying the standard’s calibrated 100Mo/97Mo (from the Pd-doped calibration step) and subsequently re-calculating the other standard ratios using the new 100Mo/97Mo to infer mass-bias. The final calibrated isotope ratios of the standard and double spike are presented in Table 2.4. These compositions gave an average δ98/95Mo value of 0.00 ± 0.03 ‰ for the measurement of double-spiked Mo(UBC) standards in wet plasma (Figure 2.4.a). Table 2.4. Calibrated double spike and standard values. Errors are 2 SD.  Errors are 2 SD (n = 15) a Note that the final 100Mo/97Mo in the standard was determined iteratively. See main text for details 100/97Mo ± 98/97Mo ± 96/97Mo ± 95/97Mo ± 94/97Mo ± 92/97Mo ± Double spike 0.94279 5E-5 0.060559 4E-6 0.026448 5E-6 0.017222 3E-6 0.009172 3E-6 0.016120 3E-6 Mo(UBC) standard 1.00420* 1E-4 2.52442 9E-5 1.74737 7E-5 1.67060   5E-5 0.97116 3E-5 1.5618 1E-4  48    Figure 2.4. Analyses of δ98/95Mo for the Mo(UBC) standard over a range of double-spike-to-standard ratios.  (A) Analysis in wet plasma in same analytical session as the spike and standard calibration. (B) Analysis in dry plasma, ESA blank correction only. (C) Analysis in dry plasma, ESA and on-peak zero blank corrections applied. Error bars indicate ± 2 SD for 3 to 5 replicate analyses. Note differences in y-axis scales. Analyses of the same double-spiked standards in a separate analytical session, this time using a desolvator unit (DSN-100) to increase sensitivity, yielded a consistent δ98/95Mo of -0.41 ± 0.05 ‰ (2 SD) for all measurements except for the mixture with the highest spike-to-sample ratio (Figure 2.3.c). The systematic shift of ~ 0.4 ‰ when using the DSN implies additional mass  49   bias effects which amplify uncertainties associated with the calibration session, in particular at extreme spike-to-sample ratios. This effect has been observed in other laboratories where the double spike technique is employed (Arnold et al., 2010; Bermin et al., 2006; Ripperger and Rehkämper, 2007; Siebert et al., 2006, 2001) and may be related to mass bias changes that do not follow the exponential fractionation law (Ripperger and Rehkämper, 2007). However, this effect is easily overcome by subtracting the sample δ-values by the daily average of spiked standards. Multiple analyses of double-spiked standards within each analytical session indicate that there is no systematic drift in these fluctuations, and that δ98/95Mo values always vary within the external 2 SD precision of the measurement (Figure 2.5.). This means that accurate δ-values are given simply by subtracting sample δ98/95Mo values by the daily average of the standard analyses to account for this offset. It was also necessary to apply an on-peak zero correction to overcome memory effects in the mass spectrometer which resulted in deviations of spiked-standard measurements by several tenths of a per mil (Figure 2.4.b). After daily normalization, the long-term average δ98/95Mo of spiked-standards is 0.00 ± 0.05‰ (2 SD) for 144 analyses in 11 separate sessions over an eight-month period, which is equal to, or better than previous Mo isotope protocols (Anbar et al., 2001; Goldberg et al., 2013; Pietruszka et al., 2006; Pietruszka and Reznik, 2008; Siebert et al., 2001; Wen et al., 2010).  50    Figure 2.5. Variation of δ98/95Mo values for Mo(UBC) standard over an analytical session. Data from February 27, 2014 at left and March 12, 2014 at right. The spike-to-standard ratio of each solution is indicated in the legend. Solid line is the average of day, dashed line indicates ± 2 SD for standards on that day 2.3.1.2 Analyses of other Mo isotope standards and natural reference materials A final verification of the double-spike protocol’s accuracy is to compare our results to published δ98/95Mo values for standards and reference materials. Repeated analyses of various double-spiked Mo isotope standards yielded δ98/95Mo which were in excellent agreement with the values compiled in Goldberg et al., (2013) (Table 2.5.). The accuracy of the full analytical protocol, including ion-exchange chromatography, was confirmed by measurements of SDO-1 shale, BCR-2 basalt, and seawater which had δ98/95Mo values of 0.79 ± 0.05 ‰, -0.04 ± 0.10‰ and 2.13 ± 0.04 ‰, respectively; these are in agreement with previous measurements (Goldberg et al. (2013; Li et al., 2014) (Table 2.5.Table 2.5.). While these Mo isotopic compositions are the  51   first to be published for BCR-2 in the literature except for a conference abstract (Li et al., 2014), further measurements in other laboratories will be required to confirm whether BCR-2 has a homogeneous Mo isotopic composition due to its possible contamination by crushing during the reference material’s preparation at the USGS (Weis et al., 2005). Table 2.5. Comparison of δ98/95Mo values obtained in this study relative to published values. Errors are ± 2 SD. Parentheses indicate number of analyses/analytical sessions   δ98/95Mo (‰) normalized to NIST-SRM-3134 = 0 ‰   This study   Published Mo isotope standards               ICL-Mo -0.14 ± 0.05 (13/5) -0.16 ± 0.05a Mo(UBC) -0.18 ± 0.05 (144/11) -- NIST-SRM-3134 0 ± 0.05 (18/7) 0.00     Roch-Mo2 -0.32 ± 0.06 (14/6) -0.34 ± 0.05a Reference Materials               Seawater 2.13 ± 0.04 (3/1) 2.09 ± 0.10a USGS SDO-1 0.79 ± 0.05 (11/5) 0.80 ± 0.14a USGS BCR-2 -0.04 ± 0.10 (3/1) -0.01  N/Ab a Goldberg et al. (2013) b Li et al. (2014) 2.3.2 Importance of spike-to-sample ratios Theoretical and empirical investigations suggest that at extreme spike-to-sample ratios error magnification can result in a degradation of accuracy and precision (Archer, 2007; John, 2012; Rudge et al., 2009). As seen in Figure 2.4.c, when the molar proportion of double spike in the spike-sample mixture was greater than 0.9, inaccurate δ98/95Mo values were obtained for the Mo(UBC) standard (after daily normalization). This phenomenon was further constrained by analyzing several Mo isotope standards in a wide range of spike-sample ratios. Results from  52   these analyses clearly demonstrated that within molar proportion of double spike of 0.4 to 0.8, δ98/95Mo was accurate but that underspiking or overspiking samples beyond this range resulted in significant deviations from the correct value by several tenths of a per mil (Figure 2.6.), similar to the results of Arnold et al. (2010). There was no significant correlation of spike-to-sample ratio and precision. However, the inaccuracies resulting from improper spike-to-sample ratios could be mitigated by normalization to a standard with a comparable spike-to-sample ratio. In one experiment, Mo(UBC), SDO-1, seawater and ICL-Mo were spiked with a molar proportion of double spike ranging from 0.25 to 0.35, which is below the optimal range of 0.4 to 0.8. The underspiked Mo(UBC) standard was offset by 0.11 ‰ relative to the δ98/95Mo of other Mo(UBC) standards analyzed in the optimal spike-to-sample ratio. Nonetheless, normalizing underspiked sample δ98/95Mo values to the underspiked Mo(UBC) standard and subsequent conversion to NIST-SRM-3134 using Equation (2.2) resulted in δ98/95Mo measurements which were in agreement with published values, demonstrating that a correction for sub-optimal spike-to-sample ratios is still possible. Because the error in δ98/95Mo increases exponentially as the spike-to-sample ratio approaches 0 or 1, corrections of underspiked or overspiked samples will require increasingly accurate matching of spike-to-sample ratios for the standard and the sample at either extremity. Adequate knowledge of sample concentrations prior to the addition of the double spike easily circumvents this potential source of error.  53    Figure 2.6. Values of δ98/95Mo for Mo(UBC) standards over a range in spike-to-sample ratios after normalization to the daily average value of the standard. At either extreme in spike-to-sample ratios, measurement accuracy degrades. Error bars indicate ± 2 SD for triplicate analyses. 2.3.3 Evaluation of column chemistry artefacts on isotope measurements Several studies have demonstrated the occurrence of isotopic fractionation during ion-exchange separation (Anbar et al., 2001; Siebert et al., 2001; Télouk et al., 1999; Wen et al., 2010). In this study, isotopic fractionation on the column was evaluated by adding the double spike to a Mo(UBC) standard after the standard was processed through ion exchange chromatography, thus preserving any isotope fractionation which would have occurred on the column. Its isotopic composition was compared to that of standards which were spiked before ion-exchange chromatography. The standards spiked before ion-exchange chromatography had δ98/95Mo values within error of 0 ‰ indicating that any fractionation or column matrix effects  54   were resolved with the double spike (Table 2.6.). The standard which was spiked after column chromatography was also equal within error to 0 ‰, although at the limit of our analytical precision, indicating that subtle on-the-column fractionation had occurred.  Table 2.6. Values of δ98/95Mo for Mo(UBC) standards double spiked before and after column chromatography. Errors are ± 2 SD.  δ98/95MoMo(UBC) (‰) Spiked before column 1 0.01 ± 0.03 Spiked before column 2 -0.01 ± 0.03 Spiked after column -0.04 ± 0.07 More concerning was the significant inaccuracy of up to -1.9 ‰ for δ98/95Mo measurements of unspiked Mo(UBC) standards passed through ion-exchange chromatography when the mass bias for these standards was corrected using SSB and SSB in combination with normalization to 105Pd/104Pd (SSB + Pd), rather than the double spike (Table 2.7.). A similar effect was observed during the analysis of environmental samples with various matrices, which were taken from a mining waste-rock facility and analyzed using the same protocols developed in this work (Table 2.7.). While Pd-doping somewhat attenuated this error, δ98/95Mo measurements obtained from Pd-normalization were still highly inaccurate compared to the results of double-spike analysis. The accurate δ98/95Mo of a standard spiked after column separation demonstrates that on-the-column isotopic fractionation did not cause this systematic offset towards lighter isotope compositions when using SSB or SSB + Pd. In addition, sample matrix effects are not expected given the consistent offset in both natural samples and pure Mo standards towards erroneously light isotopic compositions. Therefore, we conclude that this is a  55   result of a non-spectral column matrix effect as has been reported elsewhere (Pietruszka and Reznik, 2008; Shiel et al., 2009), which SSB and external normalization procedures fail to correct. This effect highlights a limitation of SSB and external normalization that is overcome by the adoption of a double spike. The robustness of the double spike to this effect provides good justification for the use double-spike analysis over other mass-bias correction methods whenever possible. For elements where double-spiking is not possible (e.g. Cu, Si), it may be useful to use a bracketing standard which has been passed through ion-exchange chemistry to help account for this effect. Table 2.7. Comparison of δ98/95Mo measurements using double spike, SSB, and SSB + Pd for mass bias corrections. Analyses include standards and natural samples. SSB + Pd is a combination of SSB and exponential law normalization to 105Pd/104Pd for mass bias corrections. Errors are ± 2 SD.   δ98/95MoMo(UBC) (‰)  SSB SSB + Pd Double spike     Mo(UBC) standard -1.93 ± 0.03 -1.31 ± 0.09 0.01 ± 0.03 Groundwater -0.39 ± 0.03 0.32 ± 0.41 0.89 ± 0.02 Molybdenite -1.89 ± 0.03 -1.46 ± 0.05 0.05 ± 0.05 Quartz Monzonite -1.04 ± 0.04 0.13 ± 0.16 0.41 ± 0.05 2.4 Conclusions Double-spike isotope analysis is the most rigorous method for instrumental mass bias correction, while also allowing for the correction of isotope fractionation occurring during sample preparation and purification. Precision and accuracy for δ98/95Mo at the 0.05 ‰ level is  56   achieved, and is sufficient for resolving Mo isotope fractionation in environmental samples. A detailed protocol for the design, development, and implementation of a Mo double spike is made available here which provides accurate δ98/95Mo values for a range of Mo isotope standards and natural reference materials. This protocol may be adapted for use in the implementation of double-spike analyses for other elements. When the proportion of double spike in the spike-sample mixture ranged from 0.4 to 0.8, accurate δ98/95Mo values were obtained in this study. Moreover, excursions of spike-to-sample ratios beyond this range could be corrected for, provided that a standard with a similar spike-to-sample ratio as the samples was used for normalization. This observation highlights the need for testing the range of acceptable spike-to-sample ratios to avoid inaccurate measurements. A column matrix effect that causes inaccurate biases towards lighter δ98/95Mo values at the per-mil level when using SSB and external normalization mass bias corrections is resolved by adding the double spike to samples prior to ion-exchange chemistry. With its growing use and significant advantages over other mass-bias correction methods, the double-spike approach is expected to become the method of choice for isotope systems where it can be applied.   57   CHAPTER 3. MOLYBDENUM AND ZINC STABLE-ISOTOPE VARIATIONS IN MINING WASTE-ROCK DRAINAGE AND WASTE ROCK AT THE ANTAMINA MINE, PERU 3.1 Introduction Leaching of metals during sulfide oxidation reactions in mining waste-rock dumps presents a global environmental challenge, the mitigation of which requires a detailed understanding of the geochemical behavior of metals in these systems. Molybdenum (Mo) and zinc (Zn) are two metals that can be released at elevated concentrations during weathering of sulfidic waste rock (e.g. Kaback, 1976; Sahu et al., 1994). Both of these elements are important micronutrients, but can cause toxic effects at elevated environmental concentrations: excess Mo is particularly harmful to ruminants which are susceptible to molybdenosis (Barceloux, 1999), and Zn toxicity has been reported in plants, humans, cattle, and invertebrates (Canadian Council of Ministers of the Environment, 1999b; Valko et al., 2005).  Molybdenum and zinc are released from mine waste during the oxidative weathering of the sulfide ore minerals molybdenite (MoS2) and sphalerite (ZnS), respectively: 2MoS2 + 9O2 + 6H2O → 2MoO42- + 4SO42- + 12H+    (3.1) ZnS + 2O2 → Zn2+ + SO42-       (3.2) Because aqueous Mo forms an anionic species (MoO42-) while Zn forms a weakly hydrolyzing cation (Zn2+), the response of these metals to changes in pH in mine drainage is distinct (Dockrey and Stockwell, 2012) and their mobility may be reduced by different  58   attenuation processes (e.g. adsorption, secondary mineral formation). Molybdate (MoO42-) is strongly adsorbed onto (oxyhydr)oxide minerals under moderately acidic conditions (Goldberg et al., 1996; Gustafsson, 2003b; Xu et al., 2006), but is mobile in neutral to alkaline pH conditions where its adsorption is minimal. In alkaline mine drainage, molybdate minerals powellite (CaMoO4) and wulfenite (PbMoO4) can act as alternative Mo solubility controls to adsorption (Conlan et al., 2012). In contrast, Zn exhibits an opposite response to pH relative to Mo: it is mobilized under acidic conditions while attenuation processes reduce its mobility under alkaline conditions. Attenuation mechanisms for Zn in mine drainage include adsorption onto (oxyhydr)oxide and carbonate minerals (Gupta et al., 1987; Iavazzo et al., 2012; Laurenzi et al., 2015), and the formation of a variety of secondary minerals such as Zn(OH)2, carbonates (e.g. smithsonite (ZnCO3), hydrozincite [Zn5(OH)6(CO3)2], silicates [e.g. willemite Zn2SiO4, hemimorphite Zn4Si2O7(OH)2·H2O] and sulfates (e.g. zincosite, ZnSO4) (Hirsche, 2012; Iavazzo et al., 2012; Jacquat et al., 2008; Wanty et al., 2013a; 2013b). Conventional geochemical and mineralogical investigations have yielded some insights into which processes are responsible for metal attenuation in mining waste-rock dumps. However, uncertainty remains with regard to the long-term fate of metals in mine waste, in large part due to the difficulty in accessing the interior of dumps for sampling, and also due to the inherent mineralogical and hydrological heterogeneity of waste rock (Amos et al., 2015). Recent analytical improvements in mass spectrometry now permit the accurate and precise measurement of small variations in metal stable-isotope ratios, which have been increasingly utilized as tracers  59   of metal attenuation processes in the environment (Wiederhold, 2015). Metal stable-isotope ratios may therefore become useful monitoring parameters for the characterization of mine drainage effluent (Matthies, 2015) by recording the geochemical history of attenuation processes which occur within waste rock dumps, provided these processes impart distinct isotopic signatures on the residual dissolved metal pool. A small number of studies have investigated the applicability of Zn isotopes as tracers of Zn geochemistry in mine drainage. Matthies et al. ( 2014b) monitored Zn isotopic compositions of moderately acidic (pH ~ 4 to 6.5) mine drainage in an experimental waste rock pile and found small variations in δ66Zn which averaged +0.06 ± 0.08 ‰ (2 SD, n = 43) relative to the IRMM 3702 Zn isotope standard over a 2-year monitoring period, suggesting little effect of secondary processes on Zn isotopes in waste rock drainage under those conditions. However, a subsequent study found a range in δ66Zn of 0.4 ‰ during experimental leaching of sulfidic mine tailings which was hypothesized to be caused by release of Zn from distinct mineral phases bearing different isotopic compositions (Matthies et al., 2014a). Other studies have examined Zn isotope compositions in surface waters downstream of historical acid-rock drainage (ARD) sites where limited variations in δ66Zn were found, leading to the proposition that Zn stable-isotope ratios were more likely to be tracers of Zn sources rather than attenuation processes, due to its conservative isotopic and geochemical behavior in such conditions (Aranda et al., 2012; Borrok et al., 2008; Fernandez and Borrok, 2009). All of these studies pointed to a relatively narrow range in δ66Zn in ARD. No studies so far have expanded analyses to directly include alkaline pH  60   mine drainage where Zn attenuation processes including adsorption and secondary Zn carbonate and hydroxide mineral formation are of greater importance. However, studies by Wanty et al. (2013a; 2013b) in an alkaline-pH river (pH ~ 8) with an elevated dissolved Zn load from mining activities showed that Zn was removed via the precipitation of hydrozincite and amorphous Zn-silicate minerals with δ66Zn values that were 0.35 to 0.5 ‰ heavier than the dissolved Zn load. Laboratory experiments have also confirmed that hydrozincite precipitation preferentially removes heavy Zn isotopes from solution (Veeramani et al., 2015). There have been no studies of Mo isotope compositions in waste-rock drainage to date. However, a compilation of natural and experimental samples measured thus far showed a substantial degree of fractionation in δ98Mo exceeding 6 ‰ (Goldberg et al., 2013). Molybdenum adsorption is known to drive changes in Mo isotope ratios in solution: Fractionation factors for Mo adsorption (αsolution-adsorbed) range from 1.00083 to 1.00276 depending on the adsorbent mineral (Barling and Anbar, 2004; Goldberg et al., 2009; Wasylenki et al., 2008). Molybdate (MoO42-) conversion to (oxo)thiomolybdate (MoOxS4-x2-) in sulfidic waters can also drive a preferential removal of light Mo isotopes from solution (Nägler et al., 2011), although this Mo removal pathway is not expected in waste rock drainage where sulfur speciation is dominated by SO42-.  These previous studies of Mo and Zn indicate that the isotopic variations of these metals may be useful parameters to track their geochemical fate in waste-rock drainage. However, more  61   measurements of δ66Zn are necessary to constrain our understanding of its isotopic behavior in waste rock drainage, in particular in alkaline-pH mine drainage where Zn attenuation is stronger. In addition, δ98Mo has yet to be explored as a tracer of Mo in mine waste. Therefore, the objectives of the present study were twofold: (1) to evaluate the extent of isotopic variability of Mo and Zn source minerals at a mine site with lithological heterogeneity in its waste rock, and (2) to evaluate whether isotopic fractionation in mine drainage was occurring after release of Mo and Zn to solution via sulfide mineral oxidation. The Mo and Zn isotopic compositions of source minerals were characterized by analyzing solid-phase samples that included waste rock, ore minerals collected from drill core, and ore concentrate. The role of secondary processes on Mo and Zn isotopic compositions in mine drainage was investigated by analyzing water samples from small- to intermediate-scale waste rock weathering experiments under field conditions covering a range in pH from acidic to alkaline. Samples were collected at the Antamina Cu–Zn–Mo–(Pb–Ag–Bi) mine, Peru, where multi-scale weathering experiments are under way to assess metal release and transport processes in mining waste rock dumps 3.2 Study site The Antamina deposit is a polymetallic skarn emplaced during a mid-Cenozoic quartz monzonite porphyry intrusion into Late Cretaceous limestone (Lipten and Smith, 2004). It is located at high elevation (4100 to 4700 m) in the Peruvian Andes (9° 32 S, 77° 03 W) (Figure 3.1). The average annual temperature is between 5.5 and 6.0 °C, and annual precipitation is 1200  62   to 1300 mm, of which the vast majority (80 to 90 %) falls during a distinct rainy season typically lasting from October to April (Blackmore et al., 2014).  The geology of the deposit is described in detail in Beckie et al. (2011) and Lipten and Smith (2004). A sequence of overlapping intrusions and metasomatic processes into the original carbonate terrain produced a highly heterogeneous deposit, which is also reflected in the waste rock. Major rock types can broadly be grouped into quartz monzonite intrusive, endoskarn, exoskarn, marble, hornfels, and limestone—although further subdivisions exist within these groups.   Figure 3.1. Location of the Antamina Mine, Peru.   63   Field weathering experiments were initiated in 2005 at Antamina to identify hydrological and geochemical processes in waste rock at different scales. In particular, this investigation includes the implementation of small-scale “field barrel” and intermediate-scale “constructed pile” experiments (Beckie et al., 2011). Field barrels are ~ 1 m high rain barrels filled with ~ 350 kg of a specific waste rock type (Figure C.1). Constructed pile experiments consist of 36 m long x 36 m wide x 10 m high piles of waste rock containing approximately 25,000 tons of material of specific lithologies that were constructed by end-dumping waste rock with haul trucks (Figure C.1). Both field barrels and piles are subjected to weathering under field conditions and their drainage hydrogeochemistry is routinely monitored. Further details of the construction and instrumentation of the field barrels and piles can be found in Bay (2009), Carazao Gallegos (2007), and Aranda (2010).  Quantitative mineralogy of the waste rock emplaced in field-barrel and constructed-pile experiments has been characterized previously using powder X-ray diffraction (XRD) with Rietveld Refinement (Dockrey et al., 2014; Peterson, 2014). Waste-rock lithologies in this study (intrusive, endoskarn, exoskarn, marble, and hornfels) contain variable sulfide amounts, ranging from 0.2 to 15 wt. % (Table C.1a and Table C.1b). Molybdenite abundance is greatest in endoskarn and intrusive waste rock, where it is present at concentrations close to the detection limit of XRD (~ 0.5 wt. %), and sphalerite is mainly enriched in exoskarn waste rock (0.4 to 5.1 wt. %). Both molybdenite and sphalerite are also present at low concentrations in other waste-rock types, as detectable using scanning-electron microscopy coupled with mineral-liberation  64   analysis software (SEM-MLA) (Aranda, 2010). Other sulfide minerals include chalcopyrite, pyrite, pyroaurite and pyrrhotite. Calcite is present (2 to 87 wt. %) in all lithologies except intrusive waste rock, and provides acid-buffering capacity in those materials. 3.3 Methods 3.3.1 Sample selection and preparation An array of solid-phase samples was collected to assess the source isotopic variability for Mo and Zn in Antamina molybdenites, sphalerites and waste rock. Unweathered ore-grade molybdenite and sphalerite samples were subsampled from Antamina drill core from a range of mineralized areas within the deposit. For molybdenites, individual grains weighing a few mg each were hand-picked from drill-core samples and dissolved in Savillex® PFA beakers on a hot-plate in 3 mL of a 3:1 mixture of concentrated HCl and HNO3 and then re-constituted in ~ 30 mL of 5 % vol. HNO3 – 1 % vol. HF for storage and mass spectrometric analysis. For sphalerites, drill-core hand samples containing 12 to 40 % Zn (as indicated by chemical assays done by mine geologists) were pulverized using a ring mill. Depending on Zn concentrations, 1 to 14 mg aliquots of the powdered sphalerite ores were weighed into Savillex® beakers and dissolved in 0.75 mL of 2:1 concentrated HNO3 and HCl on a hotplate for approximately 60 hours. The sphalerite samples were then evaporated and re-constituted in 2 mL of concentrated HCl. They were centrifuged to remove any residual refractory minerals which were discarded while the supernatant containing the dissolved sphalerites was retained for mass spectrometry.   65   A sample of molybdenite ore concentrate was also obtained from the mine. Molybdenum ore at Antamina is processed on site, beginning with a primary crusher followed by further grinding using a semi-autogenous grinder (SAG) mill. After milling, ore is further pulverized by a series of ball-mills and then sulfides are extracted from host rock using hydrochemical flotation (Bay, 2009). Approximately 40 mg of molybdenite ore concentrate containing ~ 50 wt. %. Mo was dissolved in 8 mL of 3:1 HCl:HNO3 on the hotplate. Three successive treatments of this mixture were required for full dissolution of the ore, after which it was evaporated and re-constituted in 5 vol. % HNO3 – 1 vol. % HF for storage prior to analysis. Waste-rock samples included exoskarn material from field barrel UBC-3-1A and intrusive material from field barrel UBC-2-3A. The exoskarn waste rock sample was collected prior to the beginning of field-barrel experiment and stored dry to avoid any weathering effects. The intrusive waste rock sample was collected directly from the field barrel and therefore had been exposed to weathering. In order to minimize secondary effects from weathering of that material, a large (cobble-sized) sample showing little surface alteration was selected. Waste rock samples were pulverized using a ring mill and aliquots of rock powders were weighed into Savillex® beakers for total hotplate acid digestion. Because of the elevated carbonate and silicate content of Antamina waste rock, the multi-stage dissolution method of Connelly et al. (2006) was used, in which silicates are dissolved in a mixture of HF-HNO3 followed by treatment with a mixture of HNO3–H3BO3 to remove any insoluble fluoride precipitates by complexation of B and F-. All acids used in the digestion of solid-phase samples were purified in-house from  66   concentrated reagent grade acids by sub-boiling distillation, and sample dissolution was conducted in metal-free Class 1000 clean laboratories at the Pacific Centre for Isotopic and Geochemical Research (PCIGR), University of British Columbia (Vancouver, Canada). Waste-rock drainage samples were collected from field barrels and experimental waste-rock piles at Antamina during three sampling trips to the mine between November 2012 and April 2014. Samples were selected to encompass drainage from a variety of waste-rock lithologies present at Antamina, in both the barrel- and the pile-scales. This included field barrels and piles containing intrusive, exoskarn, endoskarn, marble and hornfels waste rock that produce drainage ranging from pH < 3 to pH > 8. Samples were filtered in the field using 0.2-µm filters or, if unavailable, 0.45-µm filters, and preserved with reagent-grade HNO3 to pH < 2 prior to shipment to the laboratory. Field-filtered blanks for Mo and Zn were regularly monitored to ensure no contamination occurred during the sampling and preservation process. In the clean laboratory, aliquots of water samples were reconstituted in dilute (1 % vol. HNO3 0.05 % vol. HF) acid solutions for elemental analysis by ICP-MS (Agilent 7700x, Agilent Technologies, Santa Clara, CA, USA) at PCIGR. 3.3.2 Analytical methods Elemental concentrations for solid-phase and aqueous samples were determined by ICP-MS (Agilent 7700x) using an external multi-element calibration standard and indium as an interna-drift correction standard. Prior to isotopic analyses, Mo and Zn fractions were isolated  67   from the sample matrices using the ion-exchange purification scheme of Skierszkan et al. (2015): Briefly, samples were dissolved in 7 N HCl and loaded onto 2 mL of BioRad AG® MP-1M resin (100-200 mesh). Most matrix elements are removed in 7 N HCl and 4 N HCl, after which successive elutions of Fe (1 N HCl 0.5 N HF), Mo (2 N HCl 8 N HF) and Zn (0.1 N HBr 0.5 N HNO3) are possible. Replicate trials of synthetic multi-element solutions and USGS basalt BCR-2 show recoveries of 94 to 97 % for Mo and 100 to 103 % for Zn using this purification scheme. Molybdenum isotopic compositions on the pure sample fractions were measured using a Nu Plasma MC-ICP-MS (Nu 21, Nu Instruments Ltd, Wrexham UK). A 97Mo–100Mo double spike was used to correct for laboratory and instrumental mass fractionation as per Skierszkan et al. (2015). The Mo double spike was added prior to ion-exchange purification to account Mo isotopic fractionation during ion exchange. After ion-exchange purification, samples were re-constituted in 2 % vol. HNO3 0.1 % vol. HF and introduced into the plasma using a DSN-100 desolvating nebulizer (Nu Instruments Ltd, Wrexham UK). Molybdenum isotope masses 92 to 100 were monitored in static mode using 30 measurement cycles each having a 10 s integration time and a 2 s magnet settling time. On-peak zero blanks and ESA deflection blanks were measured and subtracted from raw sample isotope intensities. For each batch of samples, accuracy was monitored by analyzing a reference material (e.g., BCR-2, SDO-1 or open ocean water) along with samples through the full ion-exchange and analytical procedure. Molybdenum isotope compositions are reported using the δ98Mo notation:  68   δ98Mosample (‰) = ((Mo98Mo95)sample(Mo98Mo95)standard-1) × 1000  (3.3)  All δ98Mo data presented in this study follow the convention recommended by Nägler et al. (2014), i.e., sample Mo isotope ratios are normalized to NIST-SRM-3134 = + 0.25 ‰. Using this notation, our in-house Mo isotope standard “Mo(UBC)”, defined in Skierszkan et al. (2015), has a δ98Mo value of 0.07 ‰ relative to the NIST-SRM-3134 = +0.25 ‰. The long-term running average 2 SD reproducibility for analyses of Mo (UBC) is 0.05 ‰ (n = 165).  Zinc isotope measurements were also conducted on the Nu Plasma MC-ICP-MS, using the operational settings and methods presented in Shiel et al. (2013) and Shiel et al. (2009). For Zn isotopic analyses, samples and standards were doped with a pure Cu solution (High Purity Standards, Charleston, SC, USA lot 510232) in a 3:1 Cu:Zn ratio (based on mass concentrations) and were introduced into the instrument in wet plasma mode. Mass bias corrections were done using a combination of standard-sample bracketing and exponential-law normalization (e.g. Maréchal et al., 1999) of 66Zn/64Zn using 65Cu/63Cu. Online corrections of possible interferences of 64Ni on 64Zn were achieved by monitoring 62Ni. All sample measurements were done in triplicate, and prior to sample analyses, instrumental accuracy was monitored by measuring a secondary Zn isotope standard [PCIGR-2 Zn, which is 10.38 ± 0.12 ‰ (2 SD, n=40) lighter than our primary in-house standard (PCIGR-1 Zn) as defined in Shiel et al. (2009)]. Three-isotope  69   plots of δ68/64Zn as a function of δ66/64Zn showed that samples were free from isobaric interferences and displayed mass-dependent fractionations (Figure C.2). Zn isotope ratios were expressed using the δ66Zn ratio as follows: δ66Znsample (‰) = ((Zn66Zn64)sample(Zn66Zn64)PCIGR-1 Zn)-1) × 1000   (3.4) Repeated measurements of the PCIGR-1 Zn reference standard on wet plasma on two Nu Plasma MC-ICP-MS instruments (Nu21, NP214) at PCIGR show that it is 0.11 ± 0.05 ‰ (2 SD, n=8) heavier than the commonly used Zn isotope standard of the École Normale Supérieure de Lyon's Earth Science Laboratory, “JMC-Lyon” (JMC Zn standard solution, batch 3-0749L). For each batch of sample analyses in this study, an aliquot of PCIGR-1 Zn was also passed through ion-exchange purification and analyzed on the MC-ICP-MS for quality control. Concentrations of samples and bracketing standards were generally similar within < 10 % and always equal within < 25 %. A test of our secondary standard (PCIGR-2 Zn) showed that δ66Zn measurements remained accurate within a deviation in concentration of 30 % between the bracketing standards and the sample. In addition to the water samples that were collected from field barrels and experimental waste-rock piles for isotopic analysis, weekly or monthly water sampling was routinely conducted at the mine and samples were analyzed for full routine geochemical parameters (pH, alkalinity, major and minor cations and anions). These data were used to calculate mineral  70   saturation indices (SI) to identify possible mineral solubility controls on Mo and Zn. This analysis was extended to cover the entire rainy season during which isotope sampling events occurred to give a general understanding of saturation states of plausible Mo and Zn secondary minerals in each field barrel and waste-rock pile. Mineral SI calculations were computed using the PHREEQC code (Parkhurst and Appelo, 2013) with the WATEQ4F mineral database. Chemical reactions and mineral solubility constants involving Mo aqueous species and minerals (powellite, wulfenite) were added to the WATEQ4F database as described in Conlan et al. (2012). In addition, we included hydrozincite data in the WATEQ4F database, using the solubility constant determined by Preis and Gamsjäger (2001). 3.4 Results and Discussion 3.4.1 Molybdenum isotopic composition of waste rock, ores and mine drainage Molybdenum isotopic compositions in all solid phase samples exhibited significant heterogeneity with a range in δ98Mo spanning 1.2 ‰ (Table 3.1), and values of δ98Mo overlapped between endoskarn-hosted and intrusive-hosted molybdenite (Figure 3.2). These results follow a growing body of research demonstrating large Mo isotope fractionation in molybdenites, which are thought to be related to high-temperature Mo redox changes during ore genesis or Rayleigh-type distillation during the precipitation of molybdenite from hydrothermal vapors (Greber et al., 2014, 2011; Hannah et al., 2007; Mathur et al., 2010; Wieser et al., 2007; Yang et al., 2015). While the scale of spatial variability in molybdenite isotopic compositions  71   was not systematically evaluated in this study, others have shown that Mo isotopic fractionation can occur within hand samples and single crystals (Hannah et al., 2007; Mathur et al., 2010). Evidence for Mo isotopic heterogeneity at small scales was also seen in Antamina mine waste: a powdered whole-rock analysis of intrusive waste rock and a single molybdenite grain picked from the same material prior to crushing differed by 0.2 ‰ in δ98Mo (samples UBC-2-3A_WR and 2-3A_MoS2; Table 3.1).    72   Table 3.1. Molybdenum isotope data of mine drainage and rock and mineral samples from Antamina Sample ID Lithology Mo pH δ98Moa   2 SDb  ppm   ‰ Mine drainage samples Field barrels           Celda-07-14 endoskarn 11.5 7.2 -0.06 ± 0.05 Celda-07-13 endoskarn 13.9 8.2 -0.15 ± 0.02 UBC-2-0A intrusive 0.62 7.9 0.45 ± 0.05 UBC-2-1A intrusive 1.69 8.4 0.33 ± 0.02 UBC-2-3A intrusive 0.16 2.2 0.68 ± 0.03 UBC-5-2A intrusive 0.65 8.0 1.31 ± 0.04 UBC-3-1A-14 exoskarn 0.19 7.2 1.35 ± 0.06 UBC-3-1A-13 exoskarn 0.27 7.6 1.35 ± 0.03 UBC-5-0A hornfels 0.02 7.9 0.89 ± 0.07 UBC-3-3A exoskarn 0.015 7.8 0.73 ± 0.04 UBC-4-4A marble 0.010 8.6 2.07 ± 0.04 UBC-4-5-1A hornfels 0.010 8.1 0.48 ± 0.04 Experimental waste rock piles      UBC-2-D-12 intrusive 4.69 6.1 1.17 ± 0.02 UBC-2-D-14 intrusive 0.50 2.9 1.30 ± 0.04 UBC-3-D exoskarn 0.013 7.5 1.76 ± 0.01 UBC-5-D intrusive with hornfels 0.91 7.9 0.50 ± 0.04 Solid-phase samples Molybdenites       A1392 endoskarn --  0.18 ± 0.07 A2793d endoskarn --  -0.21 ± 0.05 A813 endoskarn --  0.61 ± 0.10 A2062 endoskarn --  0.57 ± 0.04 A813 endoskarn --  0.55 ± 0.06 C-07_MoS2 endoskarn --  0.09 ± 0.04 2-3A_MoS2 intrusive --  0.22 ± 0.05 A1363 intrusive --  -0.55 ± 0.04 A2194 intrusive --  -0.28 ± 0.04 Ore concentrate       Mo2496 N/A --  0.11 ± 0.07 Waste rock       UBC-2-3A_WR intrusive 626  0.47 ± 0.05 UBC-3-1A_WR exoskarn 82  0.47 ± 0.03 a δ98Mo expressed relative to NIST-SRM-3134 = +0.25 ‰; b 2 SD for triplicate analyses on the MC-ICP-MS.  73    Figure 3.2. Molybdenum isotopic compositions of mine drainage from experimental field barrels, and waste rock piles, and of molybdenites, waste-rock and ore concentrate. Data are spread vertically for clarity. The δ98Mo is reported relative to NIST-SRM-3134 = + 0.25 ‰. Error bars are 2 SD for triplicate analyses on the MC-ICP-MS, and are typically smaller than symbol size. Although the occurrence of such isotopic variations in molybdenite at small scales presents a challenge to define the isotopic composition of the source of Mo in waste-rock drainage, at larger scales (i.e. waste rock dumps) the input δ98Mo signature to mine drainage should reflect a deposit-scale averaged isotopic composition of molybdenites, thereby reducing the effect of these small-scale variations. Our best estimate for this averaged source value at Antamina is provided by the Mo isotopic composition of MoS2 ore concentrate, which is produced by milling and blending molybdenite ore and isolating it through hydrochemical flotation. The ore concentrate’s isotopic composition of 0.11 ± 0.07 ‰ (2 SD) should therefore reflect a δ98Mo composition averaged over larger scales within pit domains. No isotopic  74   fractionation is anticipated as a result of ore processing at the mine because MoS2 is concentrated by density separation that does not involve breaking of Mo—S bonds. The ore concentrate isotopic composition compared favorably to an average δ98Mo of 0.13 ± 0.82 ‰ (2 SD, n = 9) for all the Antamina molybdenites analyzed in this study (Figure 3.2). Upper and lower bounds defined by the range of δ98Mo found in molybdenites are -0.55 to +0.61 ‰.  Larger variations of δ98Mo spanning 2.2 ‰ were found in samples of mine drainage collected from field barrels and experimental waste-rock dumps weathering under field conditions. The range in Mo isotopic composition for mine drainage was -0.15 to +2.07 ‰ (Table 3.1) and the average δ98Mo was higher (0.89 ‰ ± 1.25, 2 SD, n = 16) relative to the solid phase, which indicated that Mo isotopic fractionation arose either during oxidative MoS2 dissolution or during subsequent aqueous reactions (Figure 3.2). There was no clear association of waste-rock lithology and δ98Mo in drainage waters as water samples originating from different waste-rock types exhibited overlap in Mo isotope ratios (Figure 3.2). The shift to more elevated δ98Mo in mine-drainage samples relative to the range observed in the solid phase (Figure 3.2) is consistent with observations of heavier Mo isotopic signatures in solutions relative to rocks and minerals at a global scale. The average δ98Mo of the continental crust is estimated at 0.3 to 0.4 ‰ (Greber et al., 2015; 2014), in comparison to an average value of 0.89 ± 0.91 ‰ (2 SD, n = 65) for rivers (Archer and Vance, 2008; Nägler et al., 2011; Voegelin et al., 2012) and 2.3 ‰ for the ocean (Nakagawa et al., 2012). Any combination of  75   three possible hypotheses can explain the heavy δ98Mo signature of mine-drainage samples: (1) δ98Mo in mine drainage reflects heterogeneity in the source δ98Mo signature (i.e. molybdenites); (2) oxidative molybdenite dissolution results in the preferential release of heavy Mo isotopes; and (3) there is retention of light Mo isotopes during Mo attenuation processes such as adsorption or formation of secondary molybdate minerals (e.g. powellite, wulfenite). Although some degree of heterogeneity in solid-phase Mo isotopic compositions was found, it is improbable that source isotopic compositions can fully explain the shift towards heavier δ98Mo values measured in mine-drainage samples (Figure 3.2). The heaviest δ98Mo measured in our array of solid-phase samples was 0.6 ‰ (molybdenite A813), which constitutes an upper bound for the solid-phase Mo isotopic compositions. In contrast, ten of the sixteen measurements of δ98Mo in drainage waters were heavier than this 0.6 ‰ upper-bound value for solid-phase δ98Mo. In addition, δ98Mo values in mine drainage coming from the two field barrels for which whole-rock δ98Mo measurements were available (UBC-2-3A and UBC-3-1A) also showed shifts towards heavier isotopic compositions in solution: Δ98Mosolution-whole rock values (defined as Δ98Mosolution-whole rock = δ98Mosolution - δ98Mowhole rock) were +0.21 and +0.88 ‰, respectively. Calculating an average isotopic separation factor as Δ98Mosolution-solid (Δ98Mosolution-solid = δ98Mosolution - δ98Moore concentrate) between mine drainage and the ore concentrates used as a proxy for the average molybdenite δ98Mo value for all water samples give a Δ98Mosolution-solid value of +0.77 ± 1.25 ‰ (2 SD, n = 16). These observations indicate that the shift towards  76   heavier isotopic compositions in drainage water cannot solely be attributed to variability in the solid-phase signature. A second hypothesis to explain the shift towards heavier δ98Mo in solution is that molybdenite dissolution preferentially releases heavy Mo isotopes to solution due to a change in bonding environment between trigonal-prismatic MoIVS2 and tetrahedral MoVIO42- (Johansson and Caminiti, 1986; Ramana et al., 2008). At present the exact mechanism of isotopic fractionation during mineral dissolution reactions is poorly understood due to the complexity and possible co-mingling of various processes during dissolution, such as the limited availability of atoms at the mineral surface for dissolution reactions, the role of interactions between ligands in solution and atoms at the mineral surface, microbial processes, mineralogical heterogeneity and overlapping kinetic and equilibrium isotope effects (Wiederhold, 2015). These complications make it difficult to predict a priori the direction and magnitude of any Mo isotopic fractionation during mineral dissolution. Empirical observations of the preferential release of heavy Mo isotopes during Mo dissolution experiments were found in studies by Liermann et al. (2011) and Voegelin et al. (2012), while Siebert et al. (2003) observed no fractionation during leaching. However, the ~ 0.3 ‰ enrichment in heavy Mo isotopes observed by Liermann et al. (2011) during dissolution of Mo-rich shale was interpreted to be driven by secondary adsorption onto Fe-Mn (oxyhydr)oxides rather than fractionation during the leaching process itself. Meanwhile, Voegelin et al. (2012) attributed 0.5 to 1.1 ‰ enrichments in δ98Mo during leaching of igneous rocks to the preferential release of a sulfide-bound Mo pool which was isotopically heavier than  77   the silicate matrix and the bulk rock. This interpretation is supported by the observation that heavy Mo isotopes preferentially accumulate in Mo-sulfide phases and in molybdenites rather than the silicate melt in high-temperature Mo ore-genesis processes (Greber et al., 2014). In the context of weathering of molybdenite-rich waste rock, molybdenite rather than silicate-bound Mo will dominate the pool of labile Mo. As discussed above, the average isotopic composition of Antamina molybdenite is not expected to exceed 0.6 ‰ and is more likely close to the ~ 0.1 ‰ value of the MoS2 ore concentrate. We therefore conclude that secondary processes occurring after oxidative MoS2 dissolution were most likely responsible for driving the shift towards heavier δ98Mo signatures observed in mine drainage relative to molybdenites, rather than the molybdenite dissolution process. Molybdenum adsorption is known to preferentially remove light Mo isotopes with fractionation factors (αsolution-solid) in the range of 1.00083 to 1.00276 (Barling and Anbar, 2004; Goldberg et al., 2009; Wasylenki et al., 2008). Adsorption can be anticipated in weathered waste rock where accelerated pyrite and silicate weathering leads to an abundance of Fe, Mn, and Al (oxyhydr)oxide minerals that can act as adsorbents. Adsorption is strongest for Mo at a moderately acidic pH range of 4 to 5, and is weaker in alkaline pH conditions (Goldberg et al., 1996; Gustafsson, 2003; Xu et al., 2006). Most mine-drainage samples had neutral to alkaline pH, where Mo adsorption onto (oxyhydr)oxide minerals is anticipated to be small, yet, surprisingly—among samples with pH > 7, the full range in δ98Mo spanning 2.2 ‰ was observed (Figure 3.3). However, acidic micro-sites have been shown to occur in weathered Antamina  78   waste rock that releases bulk neutral-pH drainage (Dockrey et al., 2014). It is therefore plausible that localized Mo adsorption in acidic micro-sites is responsible for retaining light Mo isotopes from solutions that display bulk neutral-pH, and drives the shift towards heavier δ98Mo values in mine drainage. Evidence for adsorbed Mo was found in chemical sequential extractions of weathered Antamina waste rock conducted by Laurenzi et al. (2015) where Mo was leached during the dissolution of crystalline reducible phases corresponding to iron oxides (Hall et al., 1996). It is therefore likely that Mo adsorption is occurring in Antamina waste rock and driving the shift towards heavier δ98Mo in mine drainage.  In addition to considering the role of adsorption, secondary minerals, including powellite (CaMoO4) and wulfenite (PbMoO4), are known to attenuate Mo in mine waste and have been observed at Antamina (Conlan et al., 2012). The isotopic fractionation associated with their formation is not known2. However, the chemical environment for Mo in both powellite and wulfenite is similar to aqueous molybdate anions, characterized by tetrahedrally coordinated Mo atoms bonding with oxygen atoms at interatomic distances of 1.76 to 1.77 Å (Achary et al., 2006; Cora et al., 2011; Johansson and Caminiti, 1986; Lugli et al., 1999). The lack of differences in Mo bonding and coordination in aqueous molybdate and metal molybdate minerals leads to the a priori hypothesis that any equilibrium Mo isotope fractionation should be minimal                                                  2 Isotopic fractionation factors for the precipitation of powellite and wulfenite were only later determined during laboratory experiments that constitute the focus of Chapter 5 in this thesis.  79   during mineral precipitation (Schauble, 2004), although this should be tested empirically. If Mo removal during molybdate mineral precipitation is a kinetic, rather than equilibrium process, isotope fractionation favoring removal of lighter Mo isotopes would be expected.   Figure 3.3. Molybdenum isotopic compositions against Mo concentrations and pH in Antamina mine drainage. Symbol shape indicates lithology of field barrel or pile, as in Figure 3.2. Horizontal shadow shows the isotopic composition of Antamina molybdenite ore concentrate. Error bars are 2 SD for triplicate analyses of δ98Mo and are typically smaller than symbols.  Powellite mineral saturation indices (SI) at the field-barrel scale were consistently < 0 throughout the study period in all field barrels (Figure 3.4) except Celda-07, which had significantly higher aqueous Mo concentrations (Table 3.1). With the exception of Celda-07, powellite precipitation was therefore not expected to play a major role at the field-barrel scale and should have had little impact on dissolved Mo isotope compositions. In the case of Celda-07, powellite formation was favored by high molybdenite contents in that material (0.2 wt. %; Table  80   C.1) and alkaline-pH conditions (pH 7.2 to 8.2 during this study period; Table 3.1) which leads to dissolved Mo concentrations that are 1 to 2 orders of magnitude higher than in all of the other field barrels. The occurrence of powellite in Celda-07 has also been confirmed by scanning-electron microscopy (c.f. Fig. 12.c and 12.d in Conlan et al., 2012). However, we consider that powellite formation in Celda-07 is unlikely to have caused significant Mo isotope fractionation in drainage from that field barrel. As mentioned previously, equilibrium isotope fractionation associated with powellite formation should be negligible due to similar Mo bonding environments in aqueous MoO42- and in powellite (Achary et al., 2006; Johansson and Caminiti, 1986). In addition, if significant kinetic isotope fractionation were occurring during powellite formation in Celda-07, we would expect a shift towards heavier δ98Mo in the residual aqueous Mo pool. However, Celda-07’s dissolved Mo isotope composition was the lightest among all mine-drainage samples. Furthermore, given the exceptionally high Mo concentrations released from this field barrel, mass balance would require that a very high proportion of Mo would have had to be removed by secondary processes to cause any measurable isotope effect in the remaining solution, which is unlikely given the slow reaction kinetics of powellite formation (Conlan et al., 2012). We therefore conclude that the relatively light δ98Mo value of mine drainage in Celda-07 most likely reflected direct dissolution of molybdenite with limited effect of secondary processes. Wulfenite supersaturation was frequently observed in field barrels (Figure 3.4), but its formation was usually limited by Pb availability: whenever Pb was measured above the detection  81   limit of ICP-MS, wulfenite supersaturation typically ensued (Figure 3.4). Lead solubility in aqueous solutions is exceedingly low due to its tendency to precipitate as sparingly soluble Pb-carbonate, Pb-sulfate, Pb-hydroxide or Pb-molybdate minerals (Conlan et al., 2012; Hem, 1976). The ample availability of dissolved anions such as HCO3-, SO42- and OH- in mine drainage therefore imparts competition for the association of MoO42- anions and Pb2+ cations and limits wulfenite precipitation to Mo-rich fluids coming into contact with Pb-rich materials (Hirsche, 2012). Overall, powellite and wulfenite formation were therefore unlikely to cause the observed Mo isotopic fractionation in mine drainage.  Figure 3.4. Seasonal variation in powellite and wulfenite mineral saturation indicies (SI) in Antamina field-barrel drainage. Data are shown for the entire rainy season during which a sample was drawn for Mo isotope analysis. Isotope sampling was conducted in April 2013 and/or 2014. Wulfenite data series are discontinuous because Pb concentrations were frequently below the ICP-MS detection limit. Also note that UBC-2-3A mineral SIs for powellite were typically lower than the range of the y-axis in the figure. Horizontal dashed line indicates SI = 0. Ultimately, we propose that the shift towards heavier δ98Mo in mine-drainage relative to the range found in solid-phase samples was most likely dominated by adsorption processes. The variations of δ98Mo in molybdenites remain a source of uncertainty in interpreting mine drainage  82   δ98Mo values in small-scale field weathering experiments. The role of secondary molybdate mineral precipitation in driving heavy Mo isotopic compositions in mine drainage—if any— should be more easily assessed once fractionation factors for their formation are known. In addition, diverse microbiological communities are known to thrive in Antamina waste rock (Dockrey et al., 2014). The redox-active geochemistry of Mo and its activity in biological enzymes may result in biological fractionations (Wang, 2012) which are at present poorly understood. Finally, a better understanding of Mo isotopic fractionation during mineral dissolution would also be useful constrain to what extent dissolution may have contributed to the heavier signatures observed in mine drainage. 3.4.2 Zinc stable-isotope composition of waste rock, sphalerites, and mine drainage In contrast to the Mo isotopic heterogeneity among solid-phase samples, remarkably little variations in Zn isotope ratios were observed among sphalerites with an average δ66Zn of 0.11 ± 0.01 ‰ (2 SD, n = 5). In addition, the Zn isotopic composition of an exoskarn waste rock sample (UBC-3-1A) was indistinguishable from the value defined by Antamina sphalerites, lending further support for a source δ66Zn value of 0.11 ‰. This provides a well-constrained isotopic signature for Zn leached into Antamina mine drainage, which falls within the estimated global average Zn isotopic signature for sphalerites compiled by Sonke et al. (2008) of 0.05 ± 0.20 ‰ (re-expressed relative to our in-house PCIGR-1 Zn reference standard).   83   In mine-drainage samples collected from field barrels and experimental waste-rock piles, a span of 0.65 ‰ in δ66Zn was observed (Table 3.2; Figure 3.5). While this range in fractionation of 0.33 ‰ amu-1 was smaller than the range found for δ98Mo in water samples (1.47 ‰ amu-1), the extent of Zn isotope fractionation measured in solution was larger than results from previous studies of Zn isotopic compositions in mine drainage (Aranda et al., 2012; Borrok et al., 2009; Matthies et al., 2014a, 2014b) . The homogeneous isotopic composition of Antamina sphalerites and the lack of substantial Zn isotopic fractionation during oxidative weathering of this mineral (Fernandez and Borrok, 2009) indicate that there are Zn attenuation processes in solution that induce Zn isotope fractionation in mine drainage.    84     Figure 3.5. Zinc isotopic compositions in mine drainage and sphalerites at Antamina. Data are spread vertically for clarity. Values of δ66Zn are expressed relative to the PCIGR-1 Zn standard. Error bars are 2 SD for triplicate analyses on the MC-ICP-MS    85    Figure 3.6. Zinc isotopic compositions against Zn concentrations and pH in Antamina mine drainage.  Symbol shape indicates lithology in field barrel or pile, as described in Figure 3.5. Horizontal shadow shows the isotopic composition of Antamina sphalerites. The δ66Zn is expressed relative to the PCIGR-1 Zn standard. Error bars are 2 SD for triplicate analyses on the MC-ICP-MS.     86   Table 3.2. Zinc isotopic compositions and Zn content in mine drainage, sphalerites, and waste rock from the Antamina mine Sample ID Lithological association Zn pH δ66Zna  ± 2 SDb    ppm   ‰ Water samples Field barrels UBC-2-0A intrusive 1.1 7.9 0.03 ± 0.05 UBC-2-3A intrusive 7.8 2.2 0.30 ± 0.03 UBC-3-1A-14 exoskarn 2.3 7.2 -0.06 ± 0.07 UBC-3-1A-13 exoskarn 2.8 7.6 -0.02 ± 0.05 UBC-3-2A exoskarn 15.6 7.8 -0.17 ± 0.11 UBC-3-3A exoskarn 0.3 7.8 -0.35 ± 0.07 UBC-4-4A hornfels 0.2 8.6 -0.17 ± 0.02 Experimental waste rock piles  UBC-2-D-14 intrusive 667 2.9 0.15 ± 0.05 UBC-2-D-12 intrusive 135 6.1 0.07 ± 0.04 UBC-3-D exoskarn 11.8 7.5 -0.03 ± 0.12 UBC-5-D intrusive / hornfels 0.2 7.9 -0.01 ± 0.05 Solid phase samples A1463 ore-grade sphalerite 152,316  0.10 ± 0.05 A1491 ore-grade sphalerite 230,113  0.12 ± 0.01 A1975_465-468 ore-grade sphalerite 333,067  0.10 ± 0.02 A1975_471-472 ore-grade sphalerite 69,688  0.106 ± 0.001 A2191 ore-grade sphalerite 401,653  0.10 ± 0.06 UBC-3-1A_WR exoskarn waste rock 6.34  0.15 ± 0.06 a δ66Zn expressed relative to PCIGR-1 Zn b 2 SD for triplicate analyses on the MC-ICP-MS The δ66Zn in mine drainage was associated with waste-rock lithology and pH: drainage produced by weathering of carbonate-rich waste rock types (exoskarn, hornfels) was isotopically lighter with a range in δ66Zn of -0.35 to -0.02 ‰, while carbonate-poor intrusive weathering produced mine drainage with a range of +0.03 to +0.30 ‰ (Figure 3.5). When plotted against  87   pH, lighter δ66Zn values were found in neutral to alkaline waters (Figure 3.6) that are associated with carbonate-rich waste rock weathering.  The shift towards lighter δ66Zn in alkaline-pH drainage coincides with conditions favorable to Zn attenuation processes, including adsorption onto Fe- and Mn-(oxyhydr)oxide minerals (Balistrieri et al., 2008; Fuller and Bargar, 2014; Juillot et al., 2008; Zachara et al., 1988), as well as formation of secondary Zn minerals such as Zn(OH)2, Zn carbonates (e.g. smithsonite, hydrozincite) and amorphous Zn silicates (Wanty et al., 2013a). Mineral saturation indices computed for exoskarn and hornfels field barrels with light δ66Zn compositions showed that the only secondary mineral that reached a state of supersaturation was hydrozincite (Figure 3.7). Hydrozincite supersaturation occurred in hornfels and exoskarn-bearing waste rock drainage as a result of their greater carbonate and Zn content. Saturation indices of other Zn mineral groups including sulfates and hydroxysulfates were always significantly undersaturated. Because dissolved SiO2 was not analyzed, it was impossible to calculate mineral SIs for Zn silicates.     88    Figure 3.7. Seasonal variation in ZnCO3•H2O, hydrozincite, and Zn(OH)2 mineral saturation indices (SI) in Antamina field-barrel drainage. Data are shown for the entire rainy season during which a sample was drawn for Zn isotope analysis. Isotope sampling was conducted in April 2013 and/or 2014. All mineral SIs in field barrel UBC-2-3A fall below range shown on graph. Note the differences in y-axis scales.  89   Previous observations in the literature show that adsorption and hydrozincite formation are both plausible explanations for the shift to lighter δ66Zn in alkaline-pH field barrel drainage. Experimental fractionation factors for Zn adsorption onto ferrihydrite, goethite, birnessite and carbonates are dominated by a preferential uptake of heavy Zn isotopes due to a change in coordination from octahedral aqueous Zn2+ ions to stiffer tetrahedral bonding in the sorbed phase, with fractionation factors (αsolution-solid) ranging from 0.99984 to 0.99731 (Balistrieri et al., 2008; Bryan et al., 2015; Dong and Wasylenki, 2014; Juillot et al., 2008). Abiotic and biologically mediated hydrozincite precipitation also has similar fractionation factors to Zn adsorption, ranging from 0.99965 to 0.99982 (Veeramani et al., 2015; Wanty et al., 2013b). (Oxyhydr)oxides and carbonates are ubiquitous during the weathering of sulfide and carbonate-rich waste mines (Al et al., 2000), and SEM imaging along with sequential chemical extractions of weathered waste rock collected in full-scale waste rock dumps at Antamina showed an association of Zn with carbonate and reducible (i.e. (oxyhydr)oxide mineral) phases (Laurenzi et al., 2015) supporting the hypothesis that either or both of these processes are responsible for Zn attenuation and associated isotopic fractionation. In contrast, under acidic mine-drainage conditions, δ66Zn values exhibited variations of less than 0.2 ‰ relative to Antamina sphalerites (Figure 3.5). Basal drainage from the intrusive waste-rock pile contained the highest Zn concentrations (135 to 667 mg/L), and its isotopic composition was indistinguishable from the sphalerite values (Figure 3.6). From these results, we infer that minimal Zn attenuation and isotopic fractionation is occurring under acidic-pH  90   conditions, and that the mine drainage δ66Zn value directly reflects the source (sphalerite), as has been reported in other studies of Zn isotopes in acid-mine drainage (Matthies et al., 2014a, 2014b). 3.4.3 Implications for use of metal stable-isotope ratios to monitor mine drainage geochemistry Significant and resolvable fractionation of Mo and Zn isotope compositions was observed in waste-rock drainage that was most likely driven by attenuation processes. This finding makes these metal isotope ratios useful tools to monitor metal attenuation processes in mine waste dumps, in addition to conventional geochemical analyses. Future research should continue to provide more constraints on the processes that govern Mo and Zn isotope fractionation in environmentally relevant reactions, including mineral dissolution, adsorption, secondary mineral formation, and microbial processes. These elements show opposite isotopic fractionation responses during adsorption processes as a result of differences in coordination chemistry during complexation at the mineral-water interface (Wasylenki et al., 2011). Taken together, Mo and Zn isotope ratios may serve as informative precursors to evolving pH conditions within mine waste dumps, much in a way their concentration patterns provide information about pH (Dockrey and Stockwell, 2012). A localized drop in pH as a result of persistent sulfide oxidation and consumption of mineral buffering capacity would be reflected in enhanced Mo adsorption and a shift towards heavier  91   δ98Mo signatures in mine drainage. This pH drop would simultaneously result in dissolution of the isotopically heavy pool of accumulated secondary Zn in the form of adsorbed Zn and secondary Zn carbonate, hydroxide and/or silicate minerals (Matthies et al., 2014a; Veeramani et al., 2015; Wanty et al., 2013a; 2013b). Zinc isotope ratios would therefore be predicted to shift towards a heavier δ66Zn during the flushing of these secondary Zn phases before reaching a stable signature resembling the value of sphalerites once persistent ARD conditions have led to the exhaustion of secondary Zn minerals. Monitoring Mo and Zn isotope compositions in mine drainage could thereby provide useful information for transient changes in geochemical conditions within waste-rock dumps that are otherwise difficult to assess from outflow concentration and pH measurements alone.  At this point, the application of Zn isotope ratios as a tracer for attenuation processes is simplified in contrast to the one of Mo isotopes due to the lack of Zn isotopic heterogeneity in the primary solid phases. Under acidic conditions, Zn isotope ratios are representative of source (sphalerite) isotopic composition with little effect of secondary processes. However, in neutral to alkaline mine drainage, heavy Zn isotopes are preferentially removed, most likely by adsorption and/or precipitation of secondary zinc carbonates, making Zn isotope ratios indicators of Zn attenuation.  Isotopic fractionations in mine drainage for Mo are larger than Zn. The application of Mo isotopic compositions as a tracer of attenuation processes is complicated by the isotopic  92   heterogeneity of molybdenites in waste rock. Despite this heterogeneity, δ98Mo measurements in mine drainage provide evidence for Mo attenuation, as demonstrated by the accumulation of heavy Mo isotopes in solution. The dominant Mo attenuation process that can be resolved at present in mine drainage using Mo isotopes is adsorption, although future studies of Mo isotopes should improve our ability to consider the role of other relevant Mo attenuation processes (e.g. metal molybdate precipitation), and constrain uncertainties associated with the mechanisms and magnitude of isotopic fractionation during molybdenite weathering.  The application of metal isotope ratios to track metal attenuation in mine drainage is still in its infancy, but the observations in this study and in previous studies (Aranda et al., 2012; Borrok et al., 2009, 2008; Matthies, 2015; Matthies et al., 2014a, 2014b) indicate their potential as powerful geochemical tracers in contaminated environments. Further constraints on the mechanisms of Mo and Zn isotopic fractionation in laboratory experiments and under field conditions should make Mo and Zn isotope ratios insightful tracers of geochemical processes occurring in a variety of subsurface environments including— but not limited to— mine waste dumps that are typically inaccessible for discrete sampling.     93   CHAPTER 4. MOLYBDENUM STABLE ISOTOPIC VARIATIONS AS INDICATORS OF MO ATTENUATION IN MINE DRAINAGE FROM FULL-SCALE WASTE-ROCK STORAGE FACILITIES 4.1 Introduction Metal leaching from sulfide mineral oxidation in mining waste rock constitutes a global environmental challenge whose mitigation requires a detailed understanding of the mobility of metals within mine waste-rock storage facilities (WRSFs). Molybdenum is a transition metal that can reach elevated concentrations in water as a result of the weathering of sulfidic mine waste (Goumih et al., 2013; Kaback and Runnells, 1980), in particular under alkaline conditions (Conlan et al., 2012). At elevated concentrations, excess Mo can be harmful, especially for ruminants, which are susceptible to molybdenosis (Barceloux, 1999). The World Health Organization recommends that Mo concentrations in drinking water should not exceed 70 µg/L (World Health Organization, 2011). Knowledge of the fate of this element within mine waste is therefore necessary for their proper environmental management. However, the complexity of processes, such as Mo release, adsorption, and precipitation, which may be simultaneously occurring in waste-rock environments, creates uncertainty regarding its fate. Long-term changes in pH and redox conditions in WRSFs further enhance the need to identify the processes of Mo attenuation and to anticipate changes in water quality in the decades following disposal.   94   The dominant Mo source in Mo-rich waste rock is the ore mineral molybdenite (MoS2). Oxidative dissolution of molybdenite leads to the release of aqueous Mo in the form of the molybdate (MoO42-) oxyanion (Smedley and Kinniburgh, 2017). Other sulfide minerals (e.g., pyrite) have Mo concentrations that are orders of magnitude lower (Greaney et al., 2016; Pichler and Mozaffari, 2015), and the main minerals that host Mo in silicate rocks are Ti-bearing minerals, such as titanite and ilmenite, which are highly resistant to chemical weathering in comparison with sulfides (Chappaz et al., 2014; Greaney et al., 2016; Smedley and Kinniburgh, 2017).  In oxidized circumneutral to alkaline aqueous environments, molybdate dominates the aqueous Mo pool. Molybdate is strongly adsorbed onto mineral surfaces under acidic conditions, with maximal adsorption in the range of pH 4 to 5 (Goldberg et al., 1996; Gustafsson, 2003; Xu et al., 2006). Molybdate adsorption is weaker under alkaline conditions, becoming minimal at pH > 8 (Goldberg et al., 1996). In alkaline mine drainage, an alternative molybdate sink is the precipitation of secondary minerals such as powellite (CaMoO4), wulfenite (PbMoO4), and NiMoO4 (Conlan et al., 2012; Essilfie-Dughan et al., 2011).  Recent studies have shown that Mo stable-isotope analyses can serve as a complementary tool to track the environmental fate of molybdenum (Siebert et al., 2015). Current knowledge of Mo isotopic fractionation was reviewed by Kendall et al. (Kendall et al., 2017). A survey of Mo isotope compositions in rivers around the world revealed that a creek receiving drainage from a   95   major Mo mine (Clear Creek, Colorado, USA) constituted an end-member with the highest Mo concentration (49 µg/L) and lightest Mo isotopic composition (Archer and Vance, 2008). Recent field waste-rock weathering experiments have also shown that mine drainage becomes enriched in heavier Mo isotopes relative to source minerals as a result of Mo attenuation processes (Skierszkan et al., 2016). Field-based studies examining Mo stable-isotope systematics during weathering and transport in rivers and soils demonstrate preferential removal of light Mo isotopes, leading to isotopically heavier fluids (King et al., 2016; Pearce et al., 2010b; Siebert et al., 2015; Skierszkan et al., 2016; Wang et al., 2015). The known isotopic fractionation factors for Mo attenuation reactions all indicate preferential removal of light isotopes from aqueous solutions, leading to a heavier isotopic signature of dissolved Mo. In oxidized aqueous environments, this result has been shown for adsorption onto Fe- and Mn-(oxyhydr)oxides (Barling and Anbar, 2004; Goldberg et al., 2009; Wasylenki et al., 2008). In anoxic and sulfidic environments, molybdate can be converted to thiomolybdate (MoOxS4-x2-), which preferentially accumulates light Mo isotopes and readily precipitates from solution (Nägler et al., 2011; Tossell, 2005). The evolution of the Mo isotopic composition in solution is therefore controlled by the extent of Mo attenuation via one or more of the aforementioned Mo removal pathways. As a result, Mo isotopic fractionation can be a useful indicator of Mo attenuation in WRSFs.  In this work, the primary objective was to determine the process of Mo attenuation within two large (> 130 million tons) WRSFs using Mo stable isotopes along with conventional geochemical analyses in mine drainage samples. Molybdenum stable isotopes were also analyzed   96   in waste rock, tailings, and ore samples to constrain the isotopic composition of the source of Mo at the study site. In addition, the surfaces of visibly weathered waste-rock samples were subjected to sequential chemical extractions to determine whether Mo was being retained and isotopically fractionated on mineral surfaces during reactive transport. Finally, surface water and groundwater samples from the mine site were analyzed to determine whether mine drainage Mo was isotopically distinguishable from natural water. The results obtained here provide an overview of the extent of Mo stable-isotope variations that may be encountered in rocks, mine drainage and natural waters at the catchment scale and relate these variations to processes controlling Mo during weathering and transport.  4.2 Study site 4.2.1 Local climate and geology Samples were collected at the Thompson Creek Molybdenum Mine, located in the Central Rocky Mountains in Idaho, USA (Figure 4.1). Precipitation falls mostly as winter snow, with occasional summer thunderstorms. Snowmelt represents > 70 % of runoff, and freshet (peak flow) typically occurs in April or May (Dockrey and Stockwell, 2012). Molybdenite (MoS2) was emplaced by a Late Cretaceous magmatic intrusion into shales and carbonaceous argillites (Dockrey and Stockwell, 2012). This entire sequence was covered much later by Eocene volcanic rock. Molybdenite is mainly concentrated within the intrusion and extends slightly into the metasedimentary country rock. Volcanic, metasedimentary and intrusive waste rocks have   97   been stockpiled in two WRSFs occupying different valleys adjacent to the open pit for close to three decades of mining. During operations, waste was classified as potentially acid generating (PAG) or non-acid generating (NAG) based on its carbonate and sulfur content. The Buckskin WRSF served as the primary repository for NAG material, and the Pat Hughes WRSF mainly comprised PAG material (Figure 4.1). PAG waste emplaced in the Buckskin WRSF is encapsulated within a layer of NAG waste. Consequently, drainage from the Buckskin WRSF has remained alkaline, with a pH of approximately 8, over the course of the mine life despite ongoing sulfide oxidation (Dockrey and Stockwell, 2012). Molybdenum concentrations have also risen steadily since 2008, ranging between approximately 60 and 80 µg/L. Drainage from the PAG-rich Pat Hughes WRSF was historically alkaline and had Mo concentrations ranging from tens to hundreds of µg/L. However, in 2006, drainage became acidic and since then has remained persistently at a pH of approximately 4.5 with Mo concentrations < 10 µg/L (Dockrey and Stockwell, 2012).     98    Figure 4.1. Map of the Thompson Creek Mine waste-rock storage facilities showing sampling locations and surficial geology.  Three sampling locations (TCg, BC3A and Sunbeam HS) fall outside of the extent of this map. Samples are described in Table 4.1. Surface geology is taken from the USGS (2005) 4.2.2 Site hydrogeology Both WRSFs occupy distinct catchments (Figure 4.1). Water movement generally follows regional topography; in valley bottoms, fast-flowing shallow colluvial/alluvial aquifers historically drained into Thompson Creek. Shallow bedrock, either volcanic or metasedimentary, is also recharged from the overlying colluvium/alluvium. Deeper bedrock aquifers have slower and upward groundwater flow driven by recharge in the neighboring mountains (Lorax Environmental Services Ltd., 2011a). The flow of shallow groundwater and flow within WRSFs respond strongly to seasonal variations, in particular, melting of the winter snowpack during   99   freshet. During baseflow (October to December), there is an increase in the proportion of groundwater discharge that contributes to flow at the mine drainage sampling stations (Lorax Environmental Services Ltd., 2011a). In the fall of 2014, shortly after the first sampling round in this study, cutoff walls were installed within the shallow colluvial/alluvial aquifers downgradient of both WRSFs to prevent leakage of mine drainage.  4.3 Sample collection and preparation 4.3.1 Water sampling and storage Water was sampled during the 2014 and 2015 fall baseflows and in spring 2016 during freshet. Groundwater and surface water samples included both locations that were impacted and those that were unimpacted by mine drainage. Mine drainage samples were collected from discharge points directly at the base of WRSFs. In addition, mine process water collected within the tailings pond and runoff percolating within the open pit of the mine were analyzed to determine the Mo isotopic signature of aqueous Mo directly after oxidative dissolution of MoS2. A brief description of the water sampling locations is provided in Table 4.1, and sampling locations are shown in Figure 4.1. During water-sample collection, temperature, electrical conductivity, oxidation-reduction potential, pH, and dissolved oxygen were measured with a flow-through cell connected to an MP-20 multi-probe instrument (QED Environmental Systems, Ann Arbor, Michigan, USA) that   100   was calibrated daily using pH 4 and pH 7 buffers. Electrical conductivity, oxidation-reduction potential and dissolved oxygen measurements were considered to be qualitatively accurate. Shallow groundwater and surface water were collected using a peristaltic pump and dedicated HDPE tubing and filtered in-line using 0.45-µm Geotech dispos-a-filtersTM. Deeper groundwater samples were drawn using either Hydrolift-2 Actuators connected to dedicated Waterra Inertial Lift Pumps or bladder pumps connected to dedicated tubing in each well. Prior to sampling, groundwater was purged, and field chemistry parameters and drawdown were monitored at regular intervals to ensure that representative formation water was being collected. Samples were filtered in-line and drawn after stabilization of field parameters.     101   Table 4.1. Description of water sampling locations Sample Type Description Pat Hughes Catchment   PW10 Groundwater Deep groundwater from metasedimentary bedrock aquifer PW13 Groundwater Deep groundwater from metasedimentary bedrock aquifer PW4 Groundwater Groundwater/mine drainage mixture from a well drilled into the base of the Pat Hughes WRSF PW7 Groundwater Intermediate groundwater from metasedimentary bedrock aquifer PW8 Groundwater Shallow groundwater from colluvial/alluvial aquifer downgradient of the Pat Hughes WRSF UPC Creek Upper reach of Pat Hughes Creek, upstream of mining activities PH Toe Mine drainage Mine drainage outflowing from the base of the Pat Hughes WRSF PH Spring Spring Spring draining into Pat Hughes Creek. The upper reach of the drainage contains a small amount of Pat Hughes waste rock Buckskin Catchment   BW1 Groundwater Deep artesian groundwater from metasedimentary bedrock aquifer BW3 Groundwater Deep artesian groundwater from metasedimentary bedrock aquifer BW4 Groundwater Shallow groundwater from colluvial/alluvial aquifer below bottom of Buckskin WRSF BuckC Mine drainage Mine drainage outflowing from the base of the Buckskin WRSF BuckSpring Spring Spring draining into Buckskin Creek below the base of the Buckskin WRSF Tailings Pond and Open Pit Drainage PD-14 Pit wall runoff Runoff from the mine's open pit TP-14 Process water Mine process water Other Catchments   Sunbeam HS Groundwater Hot spring (77 °C) surfacing ~15 km W of mine site TC4 Creek Thompson Creek, upstream of mining activities and upstream of the confluence with Buckskin Creek TCg Creek Thompson Creek, downstream of mining activities and of the confluence with Pat Hughes and Buckskin Creeks BC3A Groundwater Deep groundwater from metasedimentary bedrock aquifer, in Bruno Creek Drainage, 3 km E of Pat Hughes WRSF All water samples were stored in coolers with icepacks until refrigeration in the laboratory. Samples for sulfide analysis were preserved using NaOH and Zn-acetate. Alkalinity titrations were conducted using the Gran Method on the same day as sample collection. Water samples for metals and Mo isotopic analyses were collected into acid-washed HDPE bottles and   102   preserved with sub-boiled HNO3 to pH < 2. Metal blanks were monitored during each sampling round by passing ultrapure > 18.2 mΩ water though the peristaltic pump apparatus, including the tubing and filter, and preserved with HNO3 in the same manner as the samples; Mo concentrations measured in the blanks ranged from 0.002 to 0.09 µg/L. Sample Mo concentrations ranged from 0.29 to 3,260 µg/L; in the majority of samples, the blank would have contributed less than 5 % of the Mo, although it is possible in five (out of thirty) samples with Mo concentrations less than 2 µg/L that the blank contribution could have been greater. 4.3.2 Rock sampling and preparation Waste-rock and ore samples representing the major lithologies present at the site were obtained from the mine’s geology department collection or directly from WRSFs. To eliminate the effects of chemical weathering on samples collected in the field, their outer faces were removed using a rock saw. Rocks were pulverized using a ring mill, and aliquots were weighed into Savillex® PFA vials for hotplate dissolution using the method of Connelly et al. (2006). A sample of a white precipitate forming in basal drainage at the Pat Hughes WRSF was also collected and frozen in the field with dry ice. It was thawed in the laboratory, and water contained in the sample was removed by centrifugation. The precipitate was then dissolved in a MARS6 OneTouch Microwave Digestion System (CEM Corporation, Matthews, North Carolina, USA) using HCl-HF-HNO3 that was a 5x scale-up of the protocol of Axelsson et al. (2002).    103   All acids used for sample digestions were purified in-house from concentrated reagent-grade acids by sub-boiling distillation, and dissolutions were performed in metal-free Class 1000 clean laboratories at the Pacific Centre for Isotopic and Geochemical Research (PCIGR) at the University of British Columbia (Vancouver, Canada).  4.3.3 Characterization of weathered waste-rock surfaces The Mo distribution and isotopic composition on the surfaces of weathered waste rock were characterized by sequential chemical extractions and X-ray diffraction. The chemical extraction protocol was modified from Wiederhold et al. (2007) but downscaled proportionally due to the small amounts of surface coatings that were recovered from each sample. Sequential chemical extractions are operationally defined: in this case, they were designed to extract the Mo mobilized under reducing conditions in “Step 1” using 1 M NH2OH•HCl–1 M HCl and under oxidizing conditions in “Step 2” using aqua regia. Step 1 grouped amorphous and crystalline Fe-(oxyhydr)oxide minerals as well as water-soluble and exchangeable phases; we did not attempt to distinguish Mo association among these reservoirs. Step 2 targeted oxidizable sulfide minerals. The weathered waste-rock samples selected for extractions were coated in orange Fe-(oxyhydr)oxides. Surface minerals were scraped using silicon-carbide sandpaper and homogenized using a mortar and pestle. Qualitative mineralogy was obtained by X-ray diffraction (XRD) using a Siemens Bruker D5000 Bragg-Brentano diffractometer over a 3 to 80°   104   2θ range with a CoKα radiation source. Silicate and sulfide grains, which are insoluble in NH2OH–HCl/HCl, were inevitably entrained into the sample during scraping (Dragovich, 2006). As a result, the elemental abundances obtained in Step 1 constitute lower limits for surface coatings in weathered waste rock. For Step 1, 2 mL of 1 M NH2OH–HCl / 1 M HCl solution was added to 100-mg sample aliquots. These mixtures were then placed in a 90 °C water bath contained in a horizontal shaker for 4 hours and vortexed every 30 minutes. The mixtures were then centrifuged at 3,000 rpm for 15 minutes, and the resulting supernatants were collected and filtered (0.2 µm). To maximize recoveries, the remaining residues were subjected to the same procedure for 2 hours, and the supernatant was once again centrifuged, filtered and added to the initial supernatant. The extracted solutions were preserved in 2 % HNO3 for analysis of elemental concentrations by inductively coupled plasma optical-emission spectroscopy (ICP-OES). The solid residues remaining after Step 1 were dried overnight at 75 °C, after which, in Step 2, they were weighed into PFA vials and digested in 3 mL of aqua regia (a 3:1 concentrated HCl-HNO3 mixture) on a hotplate for 24 hours at 150 °C. The digestate was extracted after centrifugation for 15 minutes at 3,000 rpm, and to maximize recoveries, the remaining solids were rinsed and centrifuged three times in 1.5 mL of > 18.2 mΩ H2O, with the additional supernatant added to the initial aqua regia solution. This recovered solution was dried down, treated three times with a 200-µL drop of concentrated HNO3, and re-dissolved in a 2 % HNO3   105   solution for analysis by ICP-OES. The reagent blank for the chemical extraction represented at most 0.4 % of the Mo content in the samples. One sample was processed in duplicate, and the relative percent differences of Mo content between duplicates were 4 % for Step 1 and 13 % for Step 2. The Mo isotope analysis of Step 1 duplicates was reproduced within 0.01 ‰.  4.4 Analytical methods Anions and sulfide in water samples were analyzed within 1 week of sampling using ion chromatography and colorimetric methods, respectively, at ALS Environmental (Burnaby, Canada). Abundances of Al, Ca, Mo, S, Fe, K, Mg, Mn, Na, Si, and Zn were determined in sequential chemical extraction samples by ICP-OES (Varian 725-ES) with an external calibration standard and europium as an internal-drift correction standard. For water samples, major element concentrations were also determined by ICP-OES. The accuracy of the ICP-OES was monitored by analysis of a secondary ICP calibration standard solution: Analyses were on average accurate within 4 % relative to expected values for the secondary standard (Table D.1). ICP-OES external reproducibility was assessed by the relative standard deviation (RSD) of replicate analyses of standard solutions and was typically better than 3 % and no worse than 6 % (Table D.2). Absolute charge balance errors for water sample analyses were on average 3 % and always better than 10 % (Table 4.3). Trace element contents in rock and water samples were determined by inductively coupled plasma mass spectrometry (ICP-MS, Agilent 7700x). The instrument was calibrated   106   with a multi-element standard solution, and indium was used for internal drift correction. The external reproducibility of trace metal ICP-MS analyses was on average 4 %, as determined by calculation of the relative standard deviation of replicate analyses of standards (Table D.3). ICP-MS accuracy was verified with analyses of Mo and Pb contents in the USGS BCR-2 reference material; these were equal within error to the certified values (Table D.4). The reproducibilities for water, rock, and sequential chemical extraction samples were verified by analysis of full sample duplicates by ion chromatography, ICP-OES, and ICP-MS (Table D.5).  The geochemical code PHREEQC (Parkhurst and Appelo, 2013) was used to calculate mineral saturation indices in water samples using the wateq4f database, to which powellite and wulfenite solubility constants were added as described in Conlan et al. (2012) and Skierszkan et al. (2016). Aliquots for Mo isotope analyses were weighed into Savillex® PFA beakers and a 97Mo–100Mo double spike was added to correct for laboratory and instrumental mass fractionation. Samples were purified using ion-exchange chemistry as described in Skierszkan et al. (2015, 2016). Column blanks contained on average 2.5 ng of Mo and always less than 5 ng. Whenever possible, > 500 ng of sample Mo was loaded onto columns to minimize the contribution from the blank. Molybdenum isotope ratios were determined using a Nu Plasma MC-ICP-MS (Nu 21, Nu Instruments Ltd., Wrexham, UK) connected to a DSN-100 desolvating nebulizer for sample introduction. For each batch of samples, accuracy was monitored by analyzing a reference   107   material (e.g., BCR-2, SDO-1 or seawater) along with samples through the full ion-exchange and analytical procedure. Molybdenum isotopic compositions are reported using the δ98Mo notation: δ98Mosample (‰) =   ((Mo98Mo95 )sample(Mo98Mo95 )standard-1)× 1000  (4.1) Molybdenum isotope data presented in this study follow the convention recommended by Nägler et al. (2014), whereby δ98Mo values are normalized to NIST-SRM-3134 = + 0.25 ‰. The long-term average 2 SD reproducibility for analyses of the in-house Mo isotope standard “Mo(UBC)” in our laboratory is 0.07 ‰ (n = 244). For the Mo stable-isotope analysis of samples reported in this study, 33 % of the samples were analyzed in triplicate on the MC-ICP-MS, and the average 2 SD uncertainty for all sample replicate analyses was 0.06 ‰ (n = 31). 5 % of the samples were analyzed as full duplicates (including digestion and ion-exchange chemistry as applicable), with a 2 SD reproducibility better than 0.06 ‰. 4.5 Results 4.5.1 Source characterization: Rocks Volcanic waste rock and pyrite were negligible sources of Mo for mine drainage. Volcanic waste rock, emplaced well after the ore-building intrusion, was characterized by Mo contents < 1 µg/g and δ98Mo values of 0.3 to 0.4 ‰ (Table 4.2). These values are similar to the   108   Earth’s upper crustal Mo abundance of approximately 1 µg/g (Rudnick and Gao, 2013) and its δ98Mo of 0.3 to 0.4 ‰ (Voegelin et al., 2014). The Mo content in a sample of massive pyrite found in waste rock (PH-IW-16-1 Py) was also low at 0.165 µg/g (Table 4.2), in agreement with a recent study showing that pyrite is not an important host of Mo in igneous sulfides (Greaney et al., 2016). Mine tailings samples represent homogenized ore-grade intrusive and metasedimentary rock with minor quantities of residual sulfides, including pyrite, marcasite, pyrrhotite, and molybdenite (Lorax Environmental Services Ltd., 2011b). The average δ98Mo of tailings was homogenous at 0.1 ± 0.1 ‰ (2 SD, n = 5). In contrast, intrusive and metasedimentary rock samples had more elevated Mo contents spanning from 0.4 to > 7,000 µg/g and variable δ98Mo values ranging from -0.9 to 1.3 ‰ (Figure 4.2). Variation of Mo isotopic compositions within a single deposit is typical and is thought to be caused by redox changes and Rayleigh-type fractionation during the precipitation of molybdenite from hydrothermal vapors (Greber et al., 2014; Hannah et al., 2007; Mathur and Schlitt, 2010; Skierszkan et al., 2016).  Given the variability in solid-phase δ98Mo, an estimate for the isotopic signature of Mo entering mine drainage from molybdenite oxidative dissolution at the scale of the site’s WRSFs was calculated by averaging the δ98Mo value, weighted by Mo abundance, in all waste rock, tailings, and ore samples. Samples from Step 2 in the sequential chemical extractions were included in this weighted average calculation because they represented molybdenite mineralization, as demonstrated by their elevated average Mo content, which exceeded 1,500   109   µg/g, and XRD characterization (Table D.6 and Table D.8). The weighted average δ98Mo of waste rock, tailings and ore—hereafter referred to as “mine-waste Mo”—was 0.7 ± 1.0 ‰ (abundance-weighted 2 SD, n = 36).      110   Table 4.2. Molybdenum abundance and δ98Mo in waste rock, ore, and tailings. Continued on next page.     Mo  δ98Moa ± 2 SDb nc   Lithology µg/g ‰     Waste Rock           BS-QTZW-15-2 intrusive-hosted qtz-MoS2-FeS2  240d 0.98 0.06 2 PHIW15-1 intrusive 3,300 0.71 -- 1 PHIW15-4 intrusive 289 1.29 -- 1 PHIW16-3 intrusive 7,400d 0.95 0.03 2 PHIW16-5 intrusive 380 -0.10 -- 1 PHIW16-1 GM intrusive (feldspar matrix) 2.21 1.28 -- 1 PHIW16-1 Py massive pyrite 0.165       IW1 intrusive 3.8 -0.88 0.08 3 IW2 intrusive 0.40       IW3 intrusive 60.1 0.83 0.02 3 IW1 w/Mo intrusive 175 1.24 0.08 3 IW3 w/Mo intrusive 175 0.17 0.08 3 MSW2 metasedimentary 9.08 0.65 0.06 3 MSW3 metasedimentary 81 0.84 0.08 3 IW4 metasedimentary 52 1.22 0.07 2 MSW1 w/Mo metasedimentary 533 1.16 0.01 3 BS-MSW-15-1 metasedimentary 0.39       BS-MSW-15-2  metasedimentary 0.83       BS-MSW-15-6  metasedimentary 10.0 0.94 -- 1 BS-MSW-15-7 metasedimentary 1.40       BS-MSW-15-9 metasedimentary 28.9 0.43 -- 1 PH-MSW-16-1 metasedimentary 1050 0.11 -- 1 VW1 volcanic 0.66 0.3 0.1 3 VW2 volcanic 0.99 0.28 0.04 3 VW3 volcanic 0.28       a δ98Mo is expressed relative to NIST-SRM-3134 = +0.25 ‰. b 2 SD for replicate solution analyses by MC-ICP-MS. Dashed lines indicates the samples that were only analyzed once, in which case the 2 standard error was < 0.04 ‰ for 30 measurement cycles on the MC-ICP-MS. Italics indicate δ98Mo measurement and 2 SD for a full duplicate (including rock digestion and ion-exchange chemistry). c Number of replicate analyses on the MC-ICP-MS. d Average of duplicate analysis on ICP-MS.     111   Table 4.2. Molybdenum abundance and δ98Mo in waste rock, ore, and tailings. Continued from previous page.     Mo  δ98Moa ± 2 SDb nc   Lithology µg/g ‰     Molybdenite Ores           IHG1 intrusive 1,320 0.39 0.01 3 IHG2 intrusive 4,820 0.09 0.03 3 IHG3 intrusive 1,570 -0.02 0.09 3 MSO1 metasedimentary 1,270 -0.07 0.06 3 MSO2 metasedimentary 218 1.10 0.06 3 Mine Tailings           TAILS tailings 27 0.14 0.06 3 FT-15 tailings 55.5 0.20 -- 1 FT-16 tailings 71.5 0.09 -- 1 TC-16 tailings 59.4 0.07 -- 1 DFT-16 tailings 135d 0.25 -- 1 a δ98Mo is expressed relative to NIST-SRM-3134 = +0.25 ‰. b 2 SD for replicate solution analyses by MC-ICP-MS. Dashed lines indicates the samples that were only analyzed once, in which case the 2 standard error was < 0.04 ‰ for 30 measurement cycles on the MC-ICP-MS. Italics indicate δ98Mo measurement and 2 SD for a full duplicate (including rock digestion and ion-exchange chemistry). c Number of replicate analyses on the MC-ICP-MS. 4.5.2 Source characterization: Process water and runoff from the mine’s open pit Mine process water (sample TP-14) and runoff collected in the mine’s open pit (sample PD-14, Table 1) constituted end-member source waters representing MoS2 oxidative dissolution (i.e., prior to significant Mo attenuation reactions). As a result of extensive interaction with molybdenite, these water samples had Mo concentrations orders of magnitude higher (3,260 µg/L and 341 µg/L, respectively) in comparison with other mine drainage, surface water, and groundwater samples (Table 4.3). The δ98Mo of the process water and pit runoff ranged from 0.6 to 0.7 ‰ and was therefore isotopically indistinguishable from the mine-waste Mo. Process   112   water and pit runoff were also isotopically lighter than all other water samples, which spanned from 0.8 to 4.8 ‰ (Figure 4.3).  Figure 4.2. Molybdenum isotopic composition and Mo contents of waste- rock, ore, and mine tailings samples. The horizontal gray shading represents the average Mo isotopic composition, weighted by Mo concentration, for the entire sample set. 2 SD error bars are smaller than symbol sizes.      113   Table 4.3. Water-sample chemistry. Continued on next pages.       Field parameters   Anions Mineral SI     Date Alkalinity  Cond. C.B.E.a Cl- F- NO3- SO4-2 S2- powellite wulfenite Sample Type  mg/L as CaCO3 pH mS/cm % mg/L mg/L mg/L as N mg/L mg/L   Pat Hughes Catchment                       PW10 GW Oct-15 231 7.25 0.67 1.0 2.1 0.7 <dl 156 0.061 -3.1 N/Ab PW13 GW Apr-16 287 7.14 0.85 -0.9 <dl 0.6 <dl 180 0.099 -4.0 N/Ab PW4 GW Oct-15 27 5.62 0.53 0.8 21 0.3 1.9 212 -- -2.7 -1.7 PW4 GW Apr-16 3 4.83 1.33 -9.9 36 2.1 2.4 638 -- -2.0 -0.1 PW7 GW Oct-15 208 7.37 0.70 3.5 12 0.3 0.8 179 -- -2.9 -3.8 PW7 GW Apr-16 207 7.28 0.79 -4.1 9.2 0.3 0.9 196 -- -2.8 N/Ab PW8 GW Oct-15 118 6.60 0.66 6.2 3.7 0.13 3.1 238 -- -2.3 -2.5 PW8 GW Apr-16 70 7.10 0.50 -0.8 2.2 0.13 3.7 144 -- -2.5 -2.5 UPC Creek Apr-16 36 6.25 0.09 2.3 0.72 0.047 0.01 3.6 -- -4.6 -2.7 PH Spring GW spring Apr-16 52 7.71 0.44 3.3 1.1 0.049 4.0 128 -- -3.6 N/Ab PH Toe WRSF drainage Oct-14 11 4.54 1.77 7.9 23 2.7 12 1,170 -- -2.6 -0.4 PH Toe WRSF drainage Oct-15 11 4.86 2.02 1.5 19 3.6 12 1,290 -- -2.8 -0.5 PH Toe WRSF drainage Apr-16 18 4.42 3.56 -4.8 23 9.0 11 2,410 -- -2.5 0.1 Buckskin Catchment                       BW1 GW Apr-16 81 7.93 0.31 -2.1 1.2 0.27 <dl 59.1 0.0032 -2.6 N/Ab BW3 GW Oct-14 128 7.33 0.28 -5.8 1.8 0.21 <dl 41.7 -- -2.3 N/Ab BW3 GW Oct-15 116 7.94 0.30 4.4 1.8 0.20 <dl 41.4 -- -2.2 -3.2 BW4 GW Oct-15 130 7.16 1.75 2.8 <dl <dl 4.5 921 -- -1.9 -2.7 BW4 GW Apr-16 135 7.18 0.71 -1.4 <dl 0.21 0.9 205 -- -2.1 N/Ab BuckC WRSF drainage Oct-14 152 7.75 2.19 3.3 <dl <dl 14 1,260 -- -1.5 N/Ab BuckC WRSF drainage Oct-15 149 7.95 2.30 2.3 <dl <dl 12 1,270 -- -1.5 -2.9 BuckC WRSF drainage Apr-16 135 7.58 2.36 1.4 <dl <dl 9.1 1,120 -- -1.4 N/Ab BuckSpring GW spring Apr-16 126 7.80 0.27 1.9 0.65 0.14 0.009 7.5 -- -3.2 -5.1 WRSF = waste-rock storage facility; GW = groundwater; -- Indicates that the parameter was not analyzed; <dl indicates below detection limit; a C.B.E. = charge balance error, calculated using PHREEQC; b Saturation index calculation was not possible because one of the required elements was below detection limit     114   Table 4.3. Water-sample chemistry. Continued from previous pages and continued on next page.   Metals and metalloids   Date Ca K Mg Na Si Al Fe Mn Mo Pb δ98Moc   Sample  mg/L mg/L mg/L mg/L mg/L µg/L µg/L µg/L µg/L µg/L ‰ ± d n Pat Hughes Catchment                           PW10 Oct-15 83 3.1 44.6 4.49 1.3 1 1,120 21 0.1 0.009 2.91 0.07 1 PW13 Apr-16 70.4 4.0 64.1 7.97 4.71 1.2 783 61.1 0.29 <dl 4.8 0.1 2 PW4 Oct-15 53 1.2 13.59 42.4 1.6 180 <dl 190 7 0.05 2.13 0.04 1 PW4 Apr-16 99.9 1.4 37.2 58.3 5.8 7,100 361 10,450 32.2 0.793 1.19 0.03 1 PW7 Oct-15 92.8 2.0 38.7 22.7 4.5 <dl <dl <dl 3.3 0.006 1.55 0.04 1 PW7 Apr-16 81.8 1.81 33.9 20.0 5.6 75 15.8 0.14 4.48 <dl 1.59 0.03 1 PW8 Oct-15 100.3 2.3 19.9 40.5 5.4 2 <dl <dl 12.8 0.007 1.25 0.04 1 PW8 Apr-16 51.9 1.3 9.54 23.5 5.4 5.5 14.5 0.43 10.7 0.000118 1.18 0.03 1 UPC Apr-16 8.4 0.6 1.47 6.116 10.0 215 134 0.84 0.29 0.043 0.8 0.3 2 PH Spring Apr-16 50.9 0.5 6.45 21.8 7.0 4.8 11.5 0.083 0.787 <dl 2.91 0.04 1 PH Toe Oct-14 235 3.7 69.3 144 6.6 39,900 40 20,000 5.9 2.1 1.73 0.05 1 PH Toe Oct-15 225 3.7 77.3 149 <dl 36,800 69 17,900 3.9 2.1 1.64 0.05 3 PH Toe Apr-16 288 4.0 116 168 <dl 124,700 267 39,700 10.84 6.81 1.93 0.09 4 Buckskin Catchment                           BW1 Apr-16 20.4 0.68 2.95 34.1 5.69 4.2 127 16.8 16.0 <dl 0.94 0.04 1 BW3 Oct-14 31.5 0.31 6.65 22.04 5.2 <dl 22 20.4 22.9 <dl 0.87 0.04 3 BW3 Oct-15 36.8 0.7 7.93 23.7 3.0 <dl 25 15 22.5 0.014 0.9 0.1 3 BW4 Oct-15 249 4.3 30.4 185 3.9 120 <dl <dl 27.1 0.01 1.94 0.04 1 BW4 Apr-16 57.2 2.4 6.83 76.2 5.4 2.0 9 0.10 25.2 <dl 1.53 0.03 1 BuckC Oct-14 321 5.69 33.6 287 7.8 12 30 2.0 58 <dl 1.90 0.02 3 BuckC Oct-15 318 6.1 36.0 275.5 3.3 <dl 30 33 61 0.02 1.90 0.03 1 BuckC Apr-16 263 5.1 28.5 259 6.2 2.1 47 2.0 79.0 <dl 1.71 0.05 2 BuckSpring Apr-16 38.1 0.8 6.7 7.3 7.4 9.7 15.2 1.63 2.17 0.00052 0.95 0.03 1 WRSF = waste-rock storage facility; GW = groundwater; -- Indicates that the parameter was not analyzed; <dl indicates below detection limit. c δ98Mo expressed relative to NIST-SRM-3134 = +0.25 ; d ± 2 SD for duplicate or triplicate analyses on the MC-ICP-MS. For samples that were analyzed only once, 2 SE is shown and denoted by italic font.     115    Table 4.3. Water-sample chemistry. Continued from previous pages and continued on next page.       Field parameters   Anions         Date Alkalinity   Cond. C.B.E.a Cl- F- NO3- SO4-2 S2- Mineral SIb Sample Type  mg/L as CaCO3 pH mS/cm % mg/L mg/L mg/L as N mg/L mg/L powellite wulfenite Process Water and Pit Wall Runoff                         PD-14 Pit Wall Runoff Oct-14 56 8.81 0.70 -2.9 3.3 1.6 0.2 243 -- -1.4 -1.3 TP-14 Process Water Oct-14 47 7.53 3.86 3.0 341 0.53 1.2 1,830 -- 0.4 1.1 Other Nearby Catchments            TC4 Creek Oct-14 53 6.70 0.10 -4.9 <dl 0.08 <dl 10 -- -3.8 -2.4 TC4 Creek Oct-15 47 6.44 0.11 2.0 <dl 0.07 <dl 11 -- -3.6 -2.6 TCg Creek Oct-14 81 7.38 0.20 -2.9 1.0 0.13 <dl 31 -- -3.3 N/Ab TCg Creek Oct-15 114 6.79 0.21 -7.4 0.82 0.11 <dl 30 -- -3.2 N/Ab Sunbeam Hydrothermal GW Apr-16 121 9.40 0.43 -5.8 11 16 <dl 42 5.9 -3.5 N/Ab BC3A GW Oct-15 218 7.24 0.48 1.8 1.8 0.35 <dl 52 -- -2.7 -3.2 WRSF = waste-rock storage facility; GW = groundwater a C.B.E. = charge balance error, calculated using PHREEQC. b Saturation index calculation was not possible because one of the required elements was below detection limit -- Indicates that the parameter was not analyzed; <dl indicates below detection limit.       116   Table 4.3. Water-sample chemistry. Continued from previous pages.   Metals and metalloids   Date Ca K Mg Na Si Al Fe Mn Mo Pb δ98Moc   Sample  mg/L mg/L mg/L mg/L mg/L µg/L µg/L µg/L µg/L µg/L ‰ ± d n Process Water and Pit Wall Runoff                      PD-14 Oct-14 20.3 0.6 0.147 113.3 6.0 100 84 6 341 0.32 0.69 0.03 3 TP-14 Oct-14 648 54.9 37.4 325 <dl 60 54 7,030 3,260 0.9 0.58 0.02 3 Other Nearby Catchments             TC4 Oct-14 14.2 0.6 2.39 5.2 9.28 0.8 3 <dl 1.2 0.04 1.20 0.03 3 TC4 Oct-15 15.3 0.7 2.57 5.24 5.1 4 <dl <dl <dl 0.005 1.22 0.06 1 TCg Oct-14 25.2 0.7 5.76 9.3 7.5 24 4.9 <dl 2.2 <dl 1.48 0.03 3 TCg Oct-15 29.7 0.8 7.05 10.3 5.1 2 <dl <dl 0.6 <dl 1.47 0.05 1 Sunbeam Apr-16 1.73 2.3 <dl 87 39 235 1 2.81 23.9 <dl 1.00 0.06 1 BC3A Oct-15 74 1.9 23.2 1.3 0.7 <dl 352 37 4.9 0.016 1.53 0.04 1 WRSF = waste-rock storage facility; GW = groundwater c δ98Mo expressed relative to NIST-SRM-3134 = +0.25 ‰.  d ± 2 SD for duplicate or triplicate analyses on the MC-ICP-MS. For samples that were analyzed only once, 2 SE is shown and denoted by italic font. -- Indicates that the parameter was not analyzed; <dl indicates below detection limit.    117   4.5.3 Groundwater, surface water, and waste-rock storage facility drainage Groundwater, surface water, and WRSF drainage samples ranged from 0.6 to 4.8 ‰ (Figure 4.3). Despite marked differences in the aqueous chemistry between the acidic Pat Hughes WRSF and the alkaline Buckskin WRSF, δ98Mo in drainage from both WRSFs fell within a relatively narrow and isotopically heavy range of 1.6 to 1.9 ‰. Sulfide mineral oxidation caused sulfate concentrations in Pat Hughes and Buckskin waste-rock drainages to exceed 1,100 mg/L (Table 4.3). However, alkaline drainage from the Buckskin WRSF at BuckC (pH 7.6 to 8.0) had close to ten times higher Mo concentrations than drainage from the acidic Pat Hughes WRSF at PH Toe (pH 4.4 to 4.9), which had < 12 µg/L Mo.     118    Figure 4.3. Molybdenum isotopic compositions of water and solid-phase samples from the Thompson Creek Mine. GW = groundwater, WRSF = waste-rock storage facility. Weathered waste-rock extractions refer to sequential chemical extractions as described in section 4.3.3. Step 1 represents Mo leached under soluble and reducible aqueous conditions; Step 2 represents Mo leached under oxidizable conditions. The vertical shaded area represents the average δ98Mo weighted by Mo abundance in waste rock, ore, and tailings. Error bars are smaller than the symbol sizes.  4.5.3.1 Buckskin Creek Catchment Within the Buckskin Creek Catchment, WRSF drainage (BuckC) contained in excess of 1,100 mg/L sulfate and 58 to 79 µg/L Mo with a δ98Mo that was distinctly isotopically heavy (1.7 to 1.9 ‰). In contrast, groundwater and spring samples unimpacted by mining (BW1, BW3, and Buckspring) in the Buckskin Catchment had < 59 mg/L SO42-, < 23 µg/L Mo, and an     119   average δ98Mo of 0.9 ± 0.1 ‰ (2 SD, n = 4) (Table 4.3 and Figure 4.5). In the shallow colluvial/alluvial aquifer immediately downgradient of BuckC, a change in water chemistry at groundwater well BW4 was observed following the installation of a cutoff wall to intercept leakage of mine drainage: In the fall of 2015, 12 months after the wall installation, water at BW4 had δ98Mo (1.9 ‰) and SO42- (921 mg/L) similar to BuckC mine drainage; however, by the spring of 2016, its δ98Mo and SO42- had decreased to 1.5 ‰ and 205 mg/L, respectively. Molybdenum concentrations remained close to 25 µg/L at BW4 throughout this time.        120    Figure 4.4. Cross-sectional views along the Pat Hughes Creek Catchment showing concentrations of Mo and SO42- and δ98Mo in water samples. SO42- and Mo are plotted against left-hand y-axis in bar graphs (units mg/L and µg/L x 100, respectively); δ98Mo is plotted against right-hand y-axis (units ‰). TC4 and TCg samples are from Thompson Creek, upstream and downstream of the confluences with Buckskin and Pat Hughes Creeks. Dark gray shading in bar plots represents spring 2016 samples; light gray shading shows fall 2015 samples. WRSF = waste-rock storage facility. Hydrostratigraphy modified from Lorax Environmental Services Ltd. (2011a). See also Figure 4.5 for a cross-section of Buckskin Creek Catchment, noting that the y-axis scales are different between the cross-sections.        121    Figure 4.5. Cross-sectional views along the Buckskin Creek Catchment showing concentrations of Mo and SO42- and δ98Mo in water samples. SO42- and Mo are plotted against left-hand y-axis in bar graphs (units mg/L and µg/L x 100, respectively); δ98Mo is plotted against right-hand y-axis (units ‰). TC4 and TCg samples are from Thompson Creek, upstream and downstream of the confluences with Buckskin and Pat Hughes Creeks. Dark gray shading in bar plots represents spring 2016 samples; light gray shading shows fall 2015 samples. WRSF = waste-rock storage facility. Hydrostratigraphy modified from Lorax Environmental Services Ltd. (2011a). See also Figure 4.4 for a cross-section of Pat Hughes Creek Catchment, noting that the y-axis scales are different between the cross-sections.      122   4.5.3.2 Pat Hughes Creek Catchment In comparison with the narrow range of δ98Mo in unimpacted water within the Buckskin Creek Catchment, groundwater and surface water samples in the Pat Hughes Creek Catchment were more variable, spanning from 0.8 to 4.8 ‰ (Table 4.3 and Figure 4.4). This range fully encompassed the isotopic composition of Pat Hughes WRSF drainage monitored at PH Toe, which was 1.6 to 1.9 ‰. The δ98Mo in shallow groundwater and surface water (represented by PW7, PW8, and UPC) spanned from 0.8 to 1.6 ‰ and was therefore isotopically lighter than deep groundwater and a spring (represented by PW10 and PW13 and PH Spring); these had δ98Mo values of 2.9 to 4.8 ‰ and Mo concentrations that were < 2 µg/L. Molybdenum concentrations in unimpacted groundwater (PW10 and PW13) and surface water (UPC) were 1 to 2 orders of magnitude lower in the Pat Hughes Creek Catchment than those in the Buckskin Creek Catchment. The Mo content in the white precipitate forming at the outflow of the Pat Hughes WRSF was 39.8 µg/g, and its δ98Mo was 1.0 ‰ (Table D.7). This δ98Mo was 0.6 to 0.9 ‰ lighter than the values observed in WRSF drainage at PH Toe, demonstrating Mo isotopic fractionation between the drainage and the precipitate. Its mineralogy could not be identified using X-ray diffraction due to its amorphous and nanocrystalline structure, although analyses by ICP-MS (this study) and previous scanning-electron microscopy work (Lorax Environmental Services Ltd., 2011c) indicated that it is composed principally of Al, S, and O. A Raman spectrum for this     123   sample suggested that it was made of poorly crystalline basaluminite (Figure D.1), which is an aluminum sulfate mineral. Aluminum sulfates are known to precipitate in acidic and sulfate-rich waters (Nordstrom, 1982) and have been observed to form amorphous white precipitates where acid-sulfate and alkaline waters mix (Munk et al., 2002; Theobald et al., 1963). These processes are analogous to those at PH Toe where groundwater discharge mixes with Pat Hughes WRSF drainage.  4.5.4 Characterization of weathered waste-rock surfaces XRD showed that molybdenite was present in all of the surface scrapings of weathered waste rock that were subjected to sequential chemical extractions (Table D.8). Jarosite and gypsum were also present in three and four of the seven samples, respectively. Poorly crystalline Fe-(oxyhydr)oxides forming in mine waste-rock environments are typically not detectable using XRD (Das and Hendry, 2011) and were not identified by XRD scans. Secondary molybdate minerals (e.g., powellite and wulfenite) were also not observed. The chemical sequential extraction results indicated an accumulation of Fe-(oxyhydr)oxides and sulfate minerals on waste-rock coatings. Iron was the most abundant cation in Step 1 (NH2OH–HCl/HCl), ranging from 1 to 9 wt. %, while total sulfur contents ranged from 0.2 to 1 wt. % (Table D.6). Molybdenum contents in Step 1 ranged from 19 to 202 µg/g with an average of 96 µg/g (Figure 4.6 and Table D.6). Proportionally, a much larger amount of Mo was recovered in the aqua regia digestion of the residues (Step 2) as a result of the presence     124   molybdenite in the samples (Figure 4.6). The δ98Mo in Step 1 was depleted in heavy Mo isotopes in comparison with waste rock and mine water samples (Figure 4.3). It was also isotopically lighter compared with the Mo extracted in Step 2 in five of the seven samples (Figure 4.6).   Figure 4.6. Molybdenum and Fe contents and δ98Mo from sequential chemical extractions of weathered waste rock. Blue shading: Step 1 = 1M NH2OH–HCl/1 M HCl; red shading: Step 2 = Aqua regia. See section 4.3.3 for details of extraction procedure. 2 SD error bars on Mo isotopic compositions are smaller than symbol sizes. 4.6 Discussion The span in δ98Mo among samples in this study was 5.7 ‰, which nearly encompasses the range of 6.4 ‰ compiled from natural and environmental samples measured to date (Goldberg et al., 2013). The average δ98Mo in WRSF drainage sampled at BuckC and PH Toe of 1.8 ± 0.3 ‰ (2 SD, n = 6) was isotopically heavy in comparison with mine-waste Mo (0.7 ‰) and the Mo released in Step 1 of our chemical extraction, which averaged 0.2 ± 1.1 ‰ (2 SD, n = 7). The increase of δ98Mo in water relative to rocks corroborated studies of natural and contaminated environments that demonstrated an enrichment of heavy Mo isotopes in aqueous solutions during weathering and transport (Kendall et al., 2017; King et al., 2016; Pearce et al.,     125   2010b; Siebert et al., 2015; Skierszkan et al., 2016; Voegelin et al., 2012). This fractionation follows the chronological sequence of molybdenite oxidative dissolution and subsequent attenuation reactions that lead to molybdate removal from solution via adsorption or mineral precipitation. Molybdenum isotopic fractionation in WRSF drainage must therefore arise from either molybdenite dissolution or attenuation reactions or from a combination of these processes. 4.6.1 Absence of isotopic fractionation during molybdenite oxidative dissolution Pit wall runoff and mine process water samples PD-14 and TP-14 serve as analogs for the geochemical composition of solutions after molybdenite oxidative dissolution. Both of these waters underwent extensive interaction with molybdenite, as shown by their Mo concentrations ranging from 341 to 3,260 µg/L. These Mo concentrations were orders of magnitude higher than in all other water samples, which had 0.3 to 79 µg/L Mo (Table 4.3). The isotopic compositions of PD-14 and TP-14 were equal within 0.1 ‰ to mine-waste Mo, suggesting that at large spatial scales, isotopic fractionation from molybdenite oxidative weathering in the absence of Mo attenuation is negligible. Among all the water samples collected at the site, PD-14 and TP-14 constituted an end-member with the highest Mo concentration and isotopically lightest δ98Mo. This result is consistent with a global survey of δ98Mo in rivers, whose end-member with the highest Mo concentration (49 µg/L) and lightest δ98Mo (0.2 ‰) was a creek receiving Mo-rich mine drainage (Clear Creek, Colorado, USA; Archer and Vance, 2008). Other field and laboratory studies have also found limited Mo isotope fractionation as a result of the oxidative     126   dissolution process (Pearce et al., 2010a; Siebert et al., 2015, 2003; Voegelin et al., 2012). In contrast, all other groundwater, surface water, and mine drainage had heavier δ98Mo values spanning 0.8 to 4.8 ‰ and significantly lower Mo concentrations, which ranged from 0.3 to 79 µg/L. These observations suggest that Mo attenuation processes occurring after molybdenite dissolution are largely responsible for the shift toward heavy δ98Mo compositions during Mo removal. 4.6.2 Evidence for the role of adsorption in controlling Mo mobility in waste rock  Given the apparent lack of Mo isotope fractionation during molybdenite oxidative dissolution, the isotopically heavy δ98Mo of 1.6 to 1.9 ‰ in mine drainage at BuckC and PH Toe relative to mine-waste Mo can be explained only by an alternative source of isotopically heavy Mo or by enrichment in heavy Mo isotopes in solution as a result of Mo removal or by a combination of these processes. Examining water-sample data in plots of δ98Mo against Mo/SO42- provides support for the occurrence of Mo removal processes. The δ98Mo and Mo/SO42- are useful tracers of Mo attenuation along a reactive flowpath: Initial oxidation of molybdenite should produce a source-term water with high Mo/SO42- and a δ98Mo similar to mine-waste Mo. Subsequent aqueous molybdate adsorption would decrease Mo/SO42- and increase δ98Mo due to preferential removal of lighter Mo isotopes (Barling and Anbar, 2004; Siebert et al., 2015; Wasylenki et al., 2008). Any precipitation of common sulfate-bearing mine drainage minerals such as gypsum, jarosite,     127   or schwertmannite (Blowes et al., 2014) would increase Mo/SO42- with no change in δ98Mo; higher pyrite/molybdenite weathering ratios would decrease Mo/SO42-, also without affecting δ98Mo. Both δ98Mo and Mo/SO42- measurements are unchanged by dilution.  Assuming that pit wall runoff or process water (samples PD-14 and TP-14) are representative of a source-term end-member (high Mo/SO42- with isotopically light δ98Mo), drainage samples from both WRSFs fit an attenuation trend, wherein mine drainage is shifted to low Mo/SO42- with isotopically heavy δ98Mo as a result of Mo removal (Figure 4.7). Unimpacted groundwater in the Buckskin Catchment occupies intermediate compositions in the δ98Mo–Mo/SO42- space, more closely reflecting direct molybdenite dissolution (with limited Mo attenuation). The remarkably heavy δ98Mo of > 2.9 ‰ observed in select groundwater samples from the Pat Hughes Catchment (i.e., PW10 and PW13) is associated with deep sulfidic and ferruginous groundwater from the metasedimentary bedrock aquifer, with low Mo concentrations of < 2 µg/L. Upward hydraulic gradients and low sulfate concentrations at depth suggest that these waters are not affected by mine drainage and could represent a highly fractionated residual pool of aqueous Mo after redox-driven removal processes in the deep aquifer: These groundwaters were at FeS saturation with saturation indices ranging from 0.09 to -0.05, and various studies report that Mo is effectively precipitated out of sulfidic and ferruginous aqueous solutions (Helz et al., 2011, 1996; Vorlicek and Helz, 2002). This process also favors the enrichment of heavy Mo isotopes in residual aqueous Mo (Dahl et al., 2010; Nägler et al., 2011; Neubert et al., 2008; Tossell, 2005).      128    Figure 4.7. Molybdenum isotopic compositions against Mo/SO42- in samples from the Buckskin Creek and Pat Hughes Creek Catchments. Note the differences y- and x-axis scales between graphs. Pit wall runoff and mine process water samples represent the initial geochemical signature of Mo–rich waters resulting from primary molybdenite dissolution. WRSF = waste-rock storage facility, GW = groundwater, SW = surface water. Symbols indicate water type: diamonds = WRSF drainage, circles = groundwater and springs, triangles = surface water, crosses = pit runoff and process water. Groundwater sample symbol color indicates mining effects: blue = unimpacted by mining, purple = impacted. Characterization of surface scrapings on weathered waste rock collected by sequential chemical extractions and X-ray diffraction (XRD) provides evidence that Mo adsorption is the process responsible for the fractionation of δ98Mo in WRSF drainage. Recall that Mo contents recovered in Step 1 (NH2OH–HCl/HCl) of the sequential extractions represent a lower bound for adsorbed Mo on mineral surface coatings due to the entrainment of primary silicate and sulfide grains. The average Mo content in Step 1 of 96 µg/g therefore shows an association of Mo with minerals that dissolved in the reducing NH2OH–HCl/HCl environment, presumably Fe-(oxyhydr)oxides and soluble (hydroxy)sulfates. The average δ98Mo of this pool of molybdenum     129   is 0.2 ± 1.1 ‰ (2 SD, n = 7) and therefore isotopically light relative to WRSF drainage, which ranges from 1.6 to 1.9 ‰ (Figure 4.3). This fractionation is consistent with laboratory studies showing enrichment of isotopically light Mo isotopes in adsorbed phases (Barling and Anbar, 2004; Goldberg et al., 2009; Wasylenki et al., 2008). Furthermore, if the majority of the Mo released from molybdenite weathering were retained in adsorbed phases, isotopic mass balance dictates that their isotopic composition should be similar to that of mine-waste Mo at 0.7 ‰, which fits with the observed data. The molybdate minerals powellite and wulfenite are also plausible Mo sinks in oxidized mining waste rock (Conlan et al., 2012), but these are unlikely to be forming in the WRSFs studied here. Molybdenum concentrations are more than one order of magnitude below the threshold required for powellite saturation (Table 4.3), and wulfenite formation requires aqueous Pb2+, whose content is much lower in geologic samples (0.4 to 45 µg/g, data not shown) compared with Mo concentrations and which readily precipitates from aqueous solutions in the presence of common anions such as OH-, HCO3- and SO42- (Hirsche et al., 2017). These minerals were also not detected by XRD scans in weathered waste-rock samples. Groundwater discharge constitutes a possible alternative source of Mo to drainage collected at the base of WRSFs, but it is also unlikely to explain the heavy Mo isotopic signature at BuckC and PH Toe. In the Buckskin Creek Catchment, all groundwater and spring samples unimpacted by mining had Mo concentrations < 23 µg/L and δ98Mo of 0.9 to 1.0 ‰, while     130   BuckC mine drainage contained 58 to 79 µg/L Mo and had a δ98Mo of 1.7 to 1.9 ‰ (Figure 4.5). In the Pat Hughes Creek Catchment, unimpacted groundwater, surface water, and spring samples (PW10, PW13, UPC, and PH Spring) had a broader range in δ98Mo that spanned from 0.8 to 4.8 ‰ (Figure 4.4). However, the Mo concentrations in these samples were exceedingly low at < 2 µg/L. In addition, during fall sampling at PH Toe, when the groundwater contribution to mine drainage is maximal, δ98Mo values were 0.2 to 0.3 ‰ lighter than samples collecting during spring freshet (when the groundwater contribution is smallest), further indicating that groundwater was unlikely to be the source of heavy δ98Mo in the Pat Hughes WRSF drainage. 4.6.3 Molybdenum attenuation in WRSF drainage The narrow range of δ98Mo—from 1.6 to 1.9 ‰—in drainage at PH Toe and BuckC, despite a tenfold difference in Mo concentrations and their contrasting pH values, is an intriguing result that merits further discussion. A graphical projection of Mo removal via adsorption onto various mineral surfaces in the δ98Mo–Mo/SO42- space provides a mechanism to explain the similarity in δ98Mo of drainage from both WRSFs. Starting from the source-term composition defined by pit wall runoff or process water and applying equilibrium isotopic fractionation factors for Mo adsorption onto Fe-(oxyhydr)oxides from the literature (Goldberg et al., 2009), the resulting projections reach a δ98Mo plateau when the Mo/SO42- molar ratio decreases below approximately 10-4, corresponding to > 90 % Mo removal (Figure 4.8). All samples from BuckC and PH Toe fall within the values predicted for equilibrium adsorption onto ferrihydrite and     131   goethite, although the larger uncertainty in the fractionation factor of goethite yields a wider range in predicted δ98Mo values (Goldberg et al., 2009). The large surface area of these minerals acts as an important adsorption surface for molybdate (Brinza et al., 2008), and they are abundant oxidation products in pyritic mine waste-rock under the pH range found in both WRSFs (Blowes et al., 2014). A fractionation factor for Mo incorporation into Al-(hydroxy)sulfates can also be calculated from our δ98Mo measurements of mine drainage and of the Al-(hydroxy)sulfate precipitates at PH Toe. However, δ98Mo data at PH Toe and BuckC fall outside of the range predicted by this process in the δ98Mo–Mo/SO42- space (Figure 4.8). The fractionation factors for Mo adsorption onto other mineral surfaces that have an affinity for Mo and may be present in WRSFs, such as schwertmannite, clays, and pyrite (Antelo et al., 2012; Bostick et al., 2003; Goldberg et al., 1996), are not known at present. The hypothesis that extensive (> 90 %) Mo adsorption onto mineral surfaces explains the similar δ98Mo in drainage from both WRSFs is also consistent with predictions made from isotopic mass balance using results from Step 1 in the sequential chemical extractions, which suggests that a large proportion of liberated Mo is retained on mineral surfaces. In addition, while the Mo isotopic composition is similar in drainage from both WRSFs, the significantly lower Mo/SO42- at PH Toe in comparison with BuckC is consistent with a greater extent of Mo removal in the Pat Hughes WRSF (Figure 4.8). This result is expected: The Pat Hughes WRSF contains significantly more molybdenite and pyrite than the Buckskin WRSF, and Mo concentrations at PH Toe had reached tens to hundreds of µg/L prior to becoming     132   persistently acidic in 2006 (Dockrey and Stockwell, 2012). Since 2006, the stabilization of drainage to a pH of approximately 4.5—near the optimum for molybdate adsorption—and a decrease in Mo concentrations to < 10 µg/L provide strong circumstantial evidence that Mo transport is retarded by adsorption (Dockrey and Stockwell, 2012).   Figure 4.8. Molybdenum isotopic compositions against Mo/SO42- for all water samples. The curves are graphical projections of equilibrium Mo isotopic fractionation during adsorption onto ferrihydrite, goethite and aluminum hydroxysulfate starting from pit runoff sample PD-14, which represents the initial composition of water after dissolution of sulfidic waste rock. Waste-rock storage facility (WRSF) drainage falls within the projections for Mo adsorption onto ferrihydrite and goethite. The fractionation factors used in the projection of ferrihydrite and goethite adsorption are taken from Goldberg et al. (2009). The fractionation factor for aluminum hydroxysulfate adsorption is calculated from the difference of δ98Mo between the precipitate and the range observed in WRSF drainage at PH     133   Toe (Supplementary Table D.7 and Table 4.3). Curves show the minimal and maximal range in δ98Mo given reported uncertainties associated with the fractionation factors. The occurrence of a large amount of Mo removal via adsorption in the Buckskin WRSF is unexpected given the alkaline pH of drainage at BuckC. However, approximately 23 % of the Buckskin WRSF is composed of potentially acid-generating (PAG) metasedimentary and intrusive waste rock (Dockrey and Stockwell, 2012), which is encapsulated by NAG waste within the WRSF. Zones enriched in PAG waste may therefore generate localized acidic drainage and accumulation of secondary minerals such as Fe-(oxyhydr)oxides from sulfide weathering, which favor Mo adsorption and commensurate isotopic fractionation. Drainage from these acidic zones could be neutralized by mixing with alkaline water generated from the weathering of NAG waste and an influx of groundwater discharge, resulting in an increase of solution pH and transport of the remaining isotopically heavy aqueous Mo to BuckC.  4.6.4 Use of δ98Mo and Mo/SO42- to trace Mo sources in shallow groundwater (well BW4) δ98Mo and Mo/SO42- data provided useful tracers of Mo provenance at groundwater well BW4, within the shallow colluvial/alluvial aquifer immediately downgradient of the Buckskin WRSF (Figure 4.5). In November 2014, a cutoff wall was installed to intercept leakage of mine drainage into the colluvial/alluvial aquifer, which had the effect of changing the source of groundwater at BW4 from mine drainage to natural groundwater. In October 2015, BW4 showed a strong mine-drainage signature with sulfate concentrations of 921 mg/L compared with natural background water in the Buckskin Catchment that had < 60 mg/L SO42-. By April 2016, 17     134   months after the installation of the cutoff wall, SO42- concentrations had decreased to 205 mg/L, and δ98Mo and Mo/SO42- values had shifted towards the composition of natural artesian groundwater at BW1 and BW3 (Figure 4.7). These changes indicated that the wall was successfully limiting mine drainage infiltration into colluvial groundwater. This effect would not have been noticeable on the basis of Mo concentrations alone, which were unchanged during this time period. Mo isotopic mass balance calculations using the range of observed δ98Mo values in unimpacted groundwater in the Buckskin Catchment (0.9 to 1.0 ‰) and at BuckC (1.7 to 1.9 ‰) suggested that 22 to 44 % of the Mo at BW4 after installation of the cutoff wall came from BuckC infiltration, with this quantity predicted to decrease as the plume of mine drainage becomes gradually displaced by natural groundwater infiltration. 4.7 Conclusions δ98Mo analyses in rock, groundwater, surface water and mine drainage demonstrated that fractionations in excess of 5 ‰ can be present at a single mine site. This observation indicates that care must be taken to characterize the isotopic composition of these different Mo sources (e.g., groundwater, molybdenites), as they can be spatially variable within the catchment scale. Molybdenum isotopic fractionation appears to be limited during the oxidative dissolution of molybdenite in mine waste rock: Mo–rich (> 340 µg/L) mine process water and pit wall runoff had δ98Mo values that were within 0.1 ‰ of the average δ98Mo of mine-waste Mo (waste rock, tailings and ore), which was 0.7 ± 1.0 ‰ (average and 2 SD weighted by Mo abundance, n = 36).     135   Prior to any Mo attenuation process, the δ98Mo of aqueous molybdate therefore reflected the isotopic composition of the source of Mo, which was molybdenite-bearing waste rock. δ98Mo in alkaline and acidic mine drainage from the bases of two large (> 130 million tons) waste-rock storage facilities (WRSF) ranged from 1.6 to 1.9 ‰ and was therefore enriched in heavy Mo isotopes relative to mine-waste Mo. This fractionation indicates that the relatively low Mo concentrations of 4 to 79 µg/L in mine drainage result from Mo attenuation rather than a lack of Mo release. The dominant Mo attenuation process is adsorption onto mineral surfaces, as demonstrated by two lines of evidence: (1) the accumulation of isotopically light Mo in secondary minerals collected from the surfaces of weathered waste rock and (2) the increase of δ98Mo and decrease of Mo/SO42- in drainage from the base of WRSFs relative to Mo–rich source waters (pit runoff and process water) resulting from molybdenite oxidative dissolution. The occurrence of Mo adsorption in a WRSF generating alkaline mine drainage suggests that localized areas of depressed pH are likely present within the WRSF because acidic conditions are required to promote Mo adsorption (Goldberg et al., 1996). The δ98Mo can therefore constitute a tracer of small-scale mine-drainage acidification, a process that is not captured by monitoring only Mo concentrations and pH in outflow WRSF drainage. The minerals providing adsorption surfaces in weathered waste-rock are most likely Fe-(oxyhydr)oxides from pyrite weathering. Attenuated Mo should remain stable as long as the waste-rock environment remains oxidized (i.e., unsaturated and uncovered) and pH does not drop below 4, such that Fe-(oxyhydr)oxides do not dissolve.      136   δ98Mo data used in conjunction with Mo/SO42- also served as a useful tracer of groundwater remediation in a shallow aquifer after the installation of a cutoff wall to prevent mine drainage leakage. Altogether, Mo isotope ratio measurements constitute an effective indicator of processes controlling Mo mobility during the weathering of molybdenite-rich waste rock. This approach should become increasingly powerful as the mechanisms of Mo stable-isotope fractionations become unraveled through more research in both field and experimental settings.     137  CHAPTER 5. MOLYBDENUM STABLE-ISOTOPE FRACTIONATION DURING THE PRECIPITATION OF POWELLITE (CaMoO4) AND WULFENITE (PbMoO4) 5.1 Introduction Metal stable-isotope signatures are increasingly applied to trace metal sources and cycling in the environment as a result of improvements in the precision and accuracy of multi-collector inductively coupled-plasma mass spectrometry (MC-ICP-MS) in the last two decades (Albarède et al., 2004; Wiederhold, 2015 and references therein). Molybdenum is a redox-sensitive transition metal that displays resolvable isotopic fractionation in natural and laboratory settings (e.g. Kendall et al. 2017 and references therein). The application of Mo isotope ratios as geochemical tracers relies upon knowledge of the magnitude and mechanisms of Mo isotopic fractionation during chemical reactions that occur in the environment. Molybdenum stable isotopic compositions are reported using the δ98Mo notation: δ98Mosample (‰) = ((Mo98Mo95 )sample(Mo98Mo95 )standard-1) • 1000      (5.1) The most commonly used Mo isotope standard is the NIST-SRM-3134, whose δ98Mo is defined as +0.25 ‰ (Kendall et al., 2017; Nägler et al., 2014). Recent reviews of the isotopic and aqueous geochemistry of Mo are available (Kendall et al., 2017; Smedley and Kinniburgh, 2017). Although this element is an essential micronutrient, it can cause harmful effects, in particular towards ruminants which suffer from a harmful condition of Cu-deficiency known as molybdenosis following excessive Mo intake (Barceloux, 1999). In     138  most oxidized aqueous solutions at pH > 5, Mo is mainly present as the oxyanion molybdate (MoO42-). The protonated forms of molybdate, HMoO4- and H2MoO4, become prevalent at pH < 4. At pH < 5 and for concentrations exceeding 1 mM, molybdenum can also polymerize into larger molecules such as MoO7O256- and Mo8O264- (Torres et al., 2014). Adsorption and co-precipitation of molybdate onto organic matter, and Fe- and Mn-(oxy)hydroxides and Al-(hydroxy)sulfates leads to an enrichment of light Mo isotopes in adsorbed phases relative to aqueous Mo, with fractionation factors ranging from 0.8 to 2.8 ‰ (Barling and Anbar, 2004; Goldberg et al., 2009; King et al., 2018, Skierszkan et al., 2017; Wasylenki et al., 2008).  Another process that removes molybdate from oxidized aqueous solutions is the precipitation of the minerals powellite (CaMoO4) and wulfenite (PbMoO4), via the reactions: Ca2+(aq) + MoO42-(aq) ↔ CaMoO4(s)  Ksp = 10-8.05 (Essington, 1990)  (5.2) Pb2+(aq) + MoO42-(aq) ↔ PbMoO4(s) Ksp= 10-15.62 (Parkhurst and Appelo, 2013)   (5.3) The precipitation of powellite and wulfenite can sequester Mo from waters containing environmentally hazardous Mo concentrations, such as drainage from metalliferous mine waste (Conlan et al., 2012). The precipitation of these minerals therefore presents an important attenuation pathway for Mo; however, nothing is known about possible isotopic fractionation during their formation. Their reaction kinetics are drastically different: wulfenite precipitation occurs within seconds to minutes, while solutions supersaturated with respect to powellite take weeks to months to reach chemical equilibrium (Conlan et al., 2012). This contrast in reaction kinetics,     139  and the ease with which those minerals are precipitated under laboratory conditions, makes them well-suited to the study of Mo isotopic fractionation during mineral precipitation processes.  Many studies have demonstrated isotopic fractionation during the precipitation of metal cations. For example, faster reaction rates for the lighter isotopes can cause kinetic isotope effects that lead to isotopically light precipitates relative to the dissolved pool of metals (e.g. Druhan et al., 2015; Pearce et al., 2012; Skulan et al., 2002; Smith et al., 2015; Watkins et al., 2017). In addition, at isotopic equilibrium metal stable-isotope fractionation occurs when there is a difference in the energy of chemical bonds involving a metal between aqueous and mineral phases: heavier isotopes preferentially accumulate in the phase with lower bond energy (Schauble, 2004). Isotopic fractionation during precipitation can also occur when different dissolved species of a metal are isotopically fractionated relative to one another, if a specific aqueous species of distinct isotopic composition (relative to the bulk fluid) is preferentially incorporated into the solid phase (e.g. Chen et al., 2016; Schott et al., 2016).  However, comparably few studies exist on isotopic fractionation during precipitation of oxyanion-forming elements from aqueous solution. The largest isotopic fractionation of oxyanions [Se(VI)O42-, Cr(VI)O42-, S(VI)O42-]  has been associated to changes in redox state because these usually involve substantial variation in coordination chemistry (e.g. bond lengths and coordination number) and/or microbial reduction (Ellis et al., 2002; Gomes and Johnston 2017; Jamieson-Hanes et al., 2012; Johnson 2004). In the absence of redox changes, isotopic fractionation during oxyanion precipitation is still possible. One example includes sulfur isotopic     140  fractionation during precipitation of gypsum (Raab and Spiro 1991; Thode and Monster 1965; Van Driessche et al. 2016), a process that is geochemically comparable to powellite precipitation.  Molybdenum isotopic fractionation is therefore plausible during powellite and wulfenite precipitation, but has not yet been studied experimentally. This precipitation reaction does not involve a change in Mo coordination number or redox state, which should limit the magnitude of equilibrium isotope fractionation based on conventional theory (Schauble 2004), although kinetic isotope effects may still produce large fractionations. A lack of knowledge of Mo isotope fractionation during molybdate mineral precipitation has hindered our ability to interpret δ98Mo variations to trace Mo mobility in natural and contaminated environments where powellite and wulfenite precipitation occurs (Neubert et al., 2011; Skierszkan et al., 2016). Furthermore, variations in Mo isotope ratios are useful to indicate Mo attenuation in mine drainage (Skierszkan et al., 2016; 2017), but require a complete understanding of the pathways causing Mo isotopic fractionation.  This study’s goal was therefore to determine the direction and magnitude of Mo isotopic fractionation during powellite and wulfenite precipitation. Knowledge of fractionation factors for those reactions provides a critical step forward in the application of Mo stable isotopes to the environmental cycling of this element. Powellite and wulfenite were precipitated in laboratory batch experiments conducted in the presence/absence of sulfate and dissolved inorganic carbon (DIC) species (e.g. CO32- and HCO3-). These anions are common in Mo-rich waters such as those     141  associated with the weathering of sulfidic mine waste. In addition, they can form aqueous complexes with Ca2+(aq) and Pb2+(aq), and may therefore impact molybdate mineral precipitation due to competitive ion effects. The experiments presented here provide insight into the processes that cause isotopic fractionation during mineral precipitation. The observed isotopic fractionations also serve as a proof-of-concept that molybdate mineral precipitation can be quantified in the environment by analysis of δ98Mo in the residual aqueous Mo pool. 5.2 Methods 5.2.1 Experimental methods All powellite and wulfenite precipitation experiments were conducted in batch reactors at 20 °C. Reactors were sealed with a small headspace, except during sampling when they were opened for sample withdrawal. Water samples were collected from the reactors at different extents of Mo removal using acid-washed syringes. For elemental and isotopic analyses, samples were filtered (0.2 µm) and acidified in 2% sub-boiled HNO3. pH was measured with an Orion 250A+ pH-meter during sampling (Thermo Scientific). At the end of all experiments, the precipitates that had formed were collected on a paper filter (0.45 µm), air-dried, and set aside for mineralogical identification by X-ray diffraction (XRD). Prior to use in batch experiments, labware was washed in 10 % HNO3 for > 24 hours, and then rinsed in 18.2 mΩ H2O at least three times. Blanks, including those from the syringe sampling and filtering apparatus, were analyzed by ICP-MS (Agilent 7700x, Agilent Technologies, Santa Clara, California, USA) to ensure negligible Mo contamination.     142  Distinct designs were required for the powellite and wulfenite precipitation experiment sets because of their drastic differences in reaction kinetics (Conlan et al., 2012).  5.2.2 Powellite precipitation experiments The initial geochemical conditions in powellite experiments are presented in Table 5.1. For powellite batch precipitation experiments, solutions were prepared in 1-L Nalgene bottles that were placed on a magnetic stirrer for several weeks while the powellite precipitation reaction took place. Throughout the experimental duration, solution samples were periodically drawn for geochemical and isotopic analyses. Initial Mo stock solutions were created by dissolving Na2MoO4•2H2O (Sigma Aldrich) in ultra-pure > 18.2 mΩ H2O. Two experimental conditions were considered: First, in the “CaCl2 series”, solutions containing approximately 13 mM Ca2+ were created by the dissolution of CaCl2•2H2O salt, and were then mixed with the Mo stock solution. CaCl2 experiments represented powellite precipitation from solutions in the absence of aqueous complexation of Ca and Mo by competing anions and cations that may be found in more solute-rich waters. Second, in the “CaCO3/SO42- series”, excess CaCO3 powder (Fisher Scientific) was added to the Mo stock solutions as a Ca source, and this mixture was allowed to reach calcite saturation via the reaction:  CaCO3(s) + H2O(l) + CO2(aq) ↔ Ca2+(aq) + 2HCO3-(aq)     (5.4)      143  The CaCO3/SO42- series also included approximately 10 mM sulfate, which was amended to the solutions by addition of µL-amounts of H2SO4. Sulfate was included because it is the dominant anion in many Mo–rich mine-drainage waters affected by sulfide mineral oxidation, where powellite precipitation has been documented (Blanchard et al., 2015; Conlan et al., 2012; Hayes et al., 2014). The presence of sulfate and DIC species HCO3- and CO32- introduced anions capable of competing with MoO42- to form aqueous complexes with Ca2+ (e.g. CaHCO3+(aq), CaSO40(aq)). Sulfate concentrations were set low enough to ensure that gypsum (CaSO4•2H2O) was undersaturated throughout the experimental duration. In both the CaCl2 and CaCO3/SO42- series, “High-Mo” and “Low-Mo” conditions were considered, with initial Mo concentrations of approximately 7,000 µM and 1,000 µM, respectively3.                                                  3 Molar concentration units are used in Chapters 5 and 6 of the thesis. A table of conversions from mass-based to mole-based units is provided in Appendix G.     144  Table 5.1. Initial concentrations and conditions for powellite and wulfenite precipitation experiments.  Experiment Initial Mo concentration Ca source Pb source SO42- concentration [source] Alkalinity concentration [source]   Low/High µM         Powellite: CaCl2 series         POW1 Low 1,110 CaCl2 -- -- -- POW3 High 6,500 CaCl2 -- -- -- Powellite: CaCO3/SO42-series       POW2B2 Low 950 Calcite -- ~ 10 mM [H2SO4] Variable [calcite dissolution] POW2A High 7,300 Calcite -- ~ 10 mM [H2SO4] Variable [calcite dissolution] Wulfenite             WULF1 -- 82 -- Pb(NO3)2 -- -- WULF2 -- 114 -- Pb(NO3)2 ~ 10 mM [H2SO4 and Na2SO4] 0.8 mM; from NaHCO3 WULF3 -- 111 -- Pb(NO3)2 ~ 10 mM [H2SO4 and Na2SO4] 2.7 mM; from NaHCO3     145  5.2.3 Wulfenite precipitation experiments Due to the rapid kinetics of wulfenite precipitation (Conlan et al., 2012), wulfenite experiments were conducted by titrating a Pb-bearing stock solution into a series of reactors containing Mo stock solutions. Fresh ~ 100 µM Mo stock solutions were prepared for each wulfenite experiment by dissolving Na2MoO4•2H2O (Sigma Aldrich) in > 18.2 mΩ H2O. These stock solutions were then subdivided into 25 mL aliquots that were pipetted into individual centrifuge tubes. A Pb stock solution containing approximately 2.4 mM of Pb2+ was created by dissolution of Pb(NO3)2 (Sigma Aldrich). The extent of Mo removal via wulfenite precipitation was controlled by adding increasing amounts of the Pb stock solution into the separate aliquots of Mo stock solutions. The amounts of Pb added were planned to target a broad range of Mo removal. Three wulfenite experiments were conducted: (1) Experiment “WULF1” was conducted using only Na2MoO4•2H2O and Pb(NO3)2 as reagents; (2) experiment “WULF2” featured the same conditions as WULF1, but with the addition of approximately 10 mM sulfate (from Na2SO4 and H2SO4) and approximately 0.8 mM alkalinity (from NaHCO3); and (3) experiment “WULF3” was analogous to WULF2, but with a higher initial alkalinity of approximately 2.7 mM (Table 5.1). The presence of sulfate and DIC in experiments WULF2 and WULF3 introduced anions that could complex with Pb2+ and therefore compete with MoO42-. Although precipitation of wulfenite from supersaturated Pb-Mo solutions is nearly instantaneous (Conlan et al., 2012; Torres et al., 2014), experiments were left on a magnetic stir-plate for > 2 days prior     146  to ensure the establishment of chemical and isotopic equilibrium. At the end of the wulfenite experiments, aqueous samples were filtered and analyzed for pH; alkalinity using the Gran titration method (Edmond, 1970); elemental concentrations; and Mo isotopes ratios. Precipitates from each reactor were also collected for mineralogical identification by XRD. 5.2.4 Analytical methods 5.2.4.1 Elemental and isotopic analyses Elemental concentrations for Na, Ca, Mo, Pb, and S in aqueous samples were analyzed by inductively coupled-plasma optical-emission-spectroscopy (ICP-OES) (Varian 725-ES) with an external multi-element calibration standard, and europium as an internal drift-correction standard. The ICP-OES precision was assessed by replicate analyses of standard solutions and accuracy was monitored by analysis of a secondary ICP standard solution; precision and accuracy were better than 5 %. Charge balance errors for samples were also typically better than 5 % and no worse than 7 %. Molybdenum isotope ratios were analyzed using the protocol described in Skierszkan et al. (2015): A 97Mo–100Mo double spike was added to sample aliquots using the optimal spike-to-sample ratio to correct for instrumental and laboratory mass fractionation, after which the spiked aliquots were dried down and re-constituted in 7 N HCl. Molybdenum was extracted via ion-exchange chromatography by loading the double-spiked sample aliquots in columns containing 2 mL BioRad AG MP-1M® resin (100-200 mesh), and then successive elution steps, which included: (1) 30 mL of 7 N HCl, (2) 10 mL of 4 N HCl, and (3) 8 mL of 1N HF–0.5 N HCl. The purified Mo fraction was then eluted in 14 mL of 8 N HF-2     147  N HCl. In some samples, the elution steps targeting Fe (1 N HF–0.5 N HCl) and W (4 N HCl) were omitted because they decreased Mo recoveries; presumably because the relatively simple matrix of these experimental samples (as compared to natural samples) caused premature elution of Mo. Iron and W were not present in the laboratory reagents that were used, and replicate Mo isotopic measurements of samples using both the full and shortened chromatography schemes indicated that there was no impact on δ98Mo. Ion-exchange column blanks were < 3 ng, representing a negligible amount (< 0.3 %) of the sample Mo that was loaded onto columns, which was approximately 750 ng.  Following ion-exchange chromatography, samples were dried down and then refluxed in concentrated sub-boiled HNO3 at 100 °C for > 12 hrs, and then dried down and re-constituted in 2 % (v/v) HNO3 – 0.1 % (v/v) HF. Molybdenum isotopic analyses were conducted using a Nu Plasma multicollector ICP-MS (Nu21, Nu Instruments, Wrexham, UK) with a desolvating nebulizer (DSN-100, Nu Instruments, Wrexham UK) at the Pacific Centre for Isotopic and Geochemical Research. Molybdenum masses 92 to 100 were monitored in static mode, with each analysis consisting of 30 measurement cycles using a 10 s integration time and a 2 s magnet settling time. Electrostatic analyzer (ESA) and on-peak zero blanks were subtracted from sample isotope intensities prior to the double-spike inversion.  Thirteen out of the 56 samples from this study were analyzed for δ98Mo either as MC-ICP-MS replicates, or as full duplicates including ion-exchange chromatography. The typical 2 SD for all replicate and duplicate analyses was 0.1 ‰, and no worse than 0.2 ‰ (Table E.1). The laboratory's long-term average 2 SD for the Mo(UBC) isotope standard is 0.07 ‰ (n = 251). To     148  monitor the accuracy of δ98Mo analyses, Mo isotope standards and reference materials were processed alongside sample batches through ion-exchange chromatography and isotopic analyses. These included the Mo isotope standards Mo(UBC) and ICL-Mo, both doped with Ca and Na, as well as USGS reference materials BCR-2 and Nod-P-1. The isotopic compositions measured for standards and reference materials were equal within error to previously published values (Table E.2). 5.2.4.2 Mineralogy by X-ray diffraction The mineralogy of the precipitates was determined by powder XRD (D5000 Bragg-Brentano, Siemens AG, Munich, Germany) using a CoKα radiation source applied over a 2θ range of 3 to 80°, with step size of 0.03°.  5.2.4.3 Aqueous speciation modelling Mineral saturation indices (SI) and aqueous speciation were calculated using the geochemical code PHREEQC v.3.0.6-7757 (Parkhurst and Appelo, 2013), with the measured analytical concentrations and pH as inputs. Aqueous complexation reactions for relevant Mo species and powellite precipitation reactions were added to the wateq4f database as described in (Conlan et al., 2012). Note that because wulfenite is highly insoluble, the association constant for the aqueous complexation reaction Pb2+(aq) + MoO42-(aq) ↔ PbMoO4(aq)      (5.5)     149  cannot be determined accurately (Torres et al., 2014). This lack of an association constant precludes the species PbMoO4(aq) from PHREEQC calculations. Bicarbonate concentrations for the CaCO3/SO42- series were estimated from charge balance in PHREEQC because the determination of HCO3- from Gran alkalinity titrations was compromised by molybdate protonation and polymerization, which occurs near the titration endpoint (Torres et al. 2014). Simulations in PHREEQC showed that the powellite SIs determined for samples in those experiments were insensitive to the range of probable HCO3- concentrations in the experiment (Appendix F).  5.3 Results 5.3.1 Powellite precipitation 5.3.1.1 CaCl2 series In both CaCl2 experiments, formation of a white precipitate was observed after combining Ca2+ and MoO42- bearing solutions. This phenomenon was first noted after 6 days in the “Low-Mo” CaCl2 experiment POW1, and only 45 minutes in the “High-Mo” CaCl2 experiment POW3. Solution speciation calculations in PHREEQC conducted on the first sample drawn from experiments POW1 and POW3 (i.e. prior to the onset of powellite precipitation) showed that MoO42-(aq) and CaMoO4(aq) were the dominant Mo species, with minor NaMoO4-(aq); there was a larger proportion of NaMoO4-(aq) in POW3 (Figure 5.1). Molybdenum and Ca concentrations initially decreased rapidly in both experiments, but the Mo removal rate slowed as     150  the powellite saturation index (SI) gradually decreased towards chemical equilibrium (Figure 5.2 and Table 5.2). The Mo removal rate was faster in POW3, where concentrations decreased from 6,500 µM to 14.8 µM (corresponding to > 99 % Mo removal) within the first 14 days (Figure 5.2). Molybdenum removal in POW3 became minimal between 14 and 74 days, during which time the powellite SI remained close to zero (Figure 5.2 and Table 5.2), indicating the end of the reaction progress. In comparison, Mo removal was slower in experiment POW1, which only approached powellite saturation after 58 days, when the SI was 0.18 and Mo concentrations had decreased to 10 µM (> 99 % Mo removal) (Figure 5.2).     151   Figure 5.1. Aqueous speciation at the start of powellite experiments. Calcium at left, Mo at right. “Low Mo” and “High Mo” refer to the initial Mo concentrations as shown in Table 5.1.       152   Figure 5.2. Decreasing Mo concentrations and powellite saturation indices over time during powellite precipitation experiments. Circles show CaCl2 series, diamonds show CaCO3/SO42- series. Open symbols represent experiments at lower initial Mo concentrations, closed symbols represent experiments at higher initial Mo concentrations, as described in Table 5.1.       153  Table 5.2. Geochemical data for samples drawn from powellite precipitation experiments. Continued on next pages. Sample Time pH Mo Ca Na SO42- Powellite Gypsum fMo δ98Mo ID days   µM mM mM mM SIa SIa remaining ‰ ± 2 SDb n POW1A                         POW1A-1 0.003 -- 1,110 12.6 2.3 -- -- -- 1.00 0.05 0.08 2 POW1A-2 0.03 6.10 1,120 12.6 2.35 -- 2.23 -- 1.01 -- --   POW1A-3 0.1 5.87 1,105 12.48 2.30 -- 2.23 -- 1.00 -- --   POW1A-4 0.2 5.86 1,130 12.8 2.38 -- 2.24 -- 1.02 -- --   POW1A-5 0.9 5.81 1,120 12.7 2.37 -- 2.23 -- 1.01 -- --   POW1A-6 2.2 5.86 1,150 13.5 2.45 -- 2.25 -- 1.04 -- --   POW1A-7 5.2 5.99 324 11.8 2.3 -- 1.69 -- 0.29 1.03 0.07 1 POW1A-8 9.0 6.54 158 10.8 2.1 -- 1.37 -- 0.14 -- --   POW1A-9 13 5.79 124 12.0 2.3 -- 1.28 -- 0.11 1.78 0.07 1 POW1A-10 16 5.89 90 11.4 2.2 -- 1.13 -- 0.08 -- --   POW1A-11 19 5.79 67 11.5 2.2 -- 1.01 -- 0.06 -- --   POW1A-12 23 5.67 50 11.6 2.31 -- 0.87 -- 0.04 -- --   POW1A-13 28 5.67 36 11.7 2.29 -- 0.73 -- 0.03 2.66 0.07 1 POW1A-14 34 5.90 26 12.5 2.40 -- 0.60 -- 0.02 -- --   POW1A-15 58 5.66 10.07 11.5 2.2 -- 0.18 -- 0.01 3.13 0.07 1 a Powellite and gypsum SIs modelled in PHREEQC (Appendix F). b The 2 SD for duplicate and replicate analyses is shown when available. For samples analyzed only once, the long-term 2 SD of 0.07 ‰ ( n = 251) for the Mo(UBC) isotope standard is shown instead and noted by italicized font. The number of significant digits shown for concentration data reflects the internal 2 standard deviation precision of each measurement. -- indicates that the parameter was not measured.        154  Table 5.2. Geochemical data for samples drawn from powellite precipitation experiments. Continued from previous pages and continued on next pages. Sample Time pH Mo Ca Na SO42- Powellite Gypsum fMo δ98Mo ID days   µM mM mM mM SIa SIa remaining ‰ ± 2 SDb n POW2B2                         POW2B2-initial 0   950 -- 1.94 -- -- -- 1.00 0.11 0.07 1 POW2B2-1 0.12 6.17 930 10.8 1.92 10.0 2.07 -0.26 0.98 0.16 0.07 1 POW2B2-2 0.42 6.46 910 11.5 1.87 9.83 2.06 -0.25 0.96 -- --   POW2B2-3 1.3 6.51 956 12.2 1.97 10.3 2.10 -0.22 1.01 -- --   POW2B2-4 3.0 6.63 850 10.7 1.74 9.20 2.04 -0.30 0.89 0.24 0.07 1 POW2B2-5 4.0 6.57           -- -- -- --   POW2B2-6 5.5 6.55 870 10.9 1.81 9.5 2.05 -0.28 0.92 0.06 0.07 1 POW2B2-7 7.3 6.81 940 11.7 1.90 10.09 2.09 -0.24 0.99 0.01 0.07 1 POW2B2-8 8.9 6.86 965 11.99 1.97 10.3 2.10 -0.23 1.02 -- --   POW2B2-9 11 7.20 940 11.55 1.92 10.2 2.07 -0.25 0.99 0.06 0.07 1 POW2B2-10 13 7.09 880 10.9 1.83 9.67 2.05 -0.28 0.93 -- --   POW2B2-11 17 6.94 836 10.68 1.841 9.8 2.03 -0.27 0.88 0.10 0.07 1 POW2B2-12 21 6.91 747 10.1 1.83 9.6 1.97 -0.29 0.79 -- --   POW2B2-13 25 6.92 590 9.7 1.81 9.5 1.86 -0.30 0.62 0.30 0.07 1 POW2B2-14 31 7.00 488 9.74 1.87 9.8 1.78 -0.29 0.51 0.48 0.07 1 POW2B2-15 38 6.80 375 9.3 1.91 9.7 1.65 -0.30 0.39 0.62 0.07 1 POW2B2-16 52 7.10 238 9.2 1.93 9.79 1.46 -0.30 0.25 1.0 0.1 2 POW2B2-17 73 7.00 156 8.6 1.85 9.5 1.27 -0.31 0.16 -- --   POW2B2-18 97 7.00 146 --     -- -0.33 0.15 1.23 -- 1 a Powellite and gypsum SIs modelled in PHREEQC (Appendix F). b The 2 SD for duplicate and replicate analyses is shown when available. For samples analyzed only once, the long-term 2 SD of 0.07 ‰ ( n = 251) for the Mo(UBC) isotope standard is shown instead and noted by italicized font. The number of significant digits shown for concentration data reflects the internal 2 standard deviation precision of each measurement. -- indicates that the parameter was not measured.     155  Table 5.2. Geochemical data for samples drawn from powellite precipitation experiments. Continued from previous pages. Sample Time pH Mo Ca Na SO42- Powellite Gypsum fMo δ98Mo ID days   µM mM mM mM SIa SIa remaining ‰ ± 2 SDb n POW3                         POW3-initial 0 -- 6,500 -- 13.5 -- -- -- 1.00 0.07 0.07 1 POW3-1 0.4 6.11 6,530 10.5 14.0 -- 2.85 -- 1.00 0.12 0.07 1 POW3-2 0.6 5.93 3,400 7.59 14.1 -- 2.52 -- 0.52 0.62 0.07 1 POW3-3 0.7 -- 2,340 6.75 14.7 -- -- -- 0.36 0.86 0.07 1 POW3-4 0.9 5.78 1,680 6.00 14.3 -- 2.17 -- 0.26 -- --   POW3-5 1.7 5.52 850 5.14 13.8 -- 1.85 -- 0.13 1.51 0.06 2 POW3-6 3.4 5.41 419 4.75 15.7 -- 1.51 -- 0.064 -- --   POW3-7 4.4 6.27 259 4.16 14.1 -- 1.29 -- 0.040 1.99 0.07 1 POW3-8 5.9 5.41 115 4.38 15.2 -- 0.94 -- 0.018 2.70 0.07 1 POW3-9 7.6 5.39 45 4.08 14.5 -- 0.52 -- 0.007 -- --   POW3-10 14 6.17 14.8 3.57 12.92 -- 0.03 -- 0.0023 -- --   POW3-11 21 5.80 12.8 3.83 13.7 -- -0.02 -- 0.0020 0.69 0.07 1 POW3-12 31 5.52 11.5 4.00 14.3 -- -0.09 -- 0.0018 0.41 0.07 1 POW3-13 52 4.60 9.7 3.85 13.8 -- -0.21 -- 0.0015 0.47 0.07 1 POW3-14 73 6.56 9.0 3.75 13.5 -- -0.18 -- 0.0014 0.46 0.01 2 a Powellite and gypsum SIs modelled in PHREEQC (Appendix F). b The 2 SD for duplicate and replicate analyses is shown when available. For samples analyzed only once, the long-term 2 SD of 0.07 ‰ ( n = 251) for the Mo(UBC) isotope standard is shown instead and noted by italicized font. The number of significant digits shown for concentration data reflects the internal 2 standard deviation precision of each measurement. -- indicates that the parameter was not measured.     156  The initial and rapid decrease in Mo concentrations in both experiments was accompanied by a rise in δ98Mo in the remaining aqueous Mo pool: In POW1, δ98Mo rose from a starting value of 0.1 ‰ to 3.1 ‰ at 99 % Mo removal at the experiment’s end at 58 days, while in POW3, δ98Mo rose from 0.1 ‰ to 2.7 ‰ after 6 days and 98 % Mo removal (Figure 5.3). Interestingly, once powellite SIs approached equilibrium values after 20 days in POW3 the solution became isotopically lighter, stabilizing around 0.4 to 0.5 ‰ for the remaining several weeks of the experimental duration (Figure 5.3).  XRD scans confirmed that the white precipitates formed in both experiments were powellite. 5.3.1.2 CaCO3/SO42- series Molybdenum concentrations in the CaCO3/SO42- series also decreased through time, albeit much more slowly than in the CaCl2 series (Figure 5.2). Molybdenum concentrations stabilized at a much higher threshold of approximately 150 µM after three to four months, despite thermodynamic calculations in PHREEQC that indicated conditions well above powellite saturation, with SIs > 0.6 (Figure 5.3 and Table 5.2). In similar experiments by Conlan et al. (2012), faster reaction kinetics were also observed when CaCl2 was used as a source of Ca (rather than CaCO3) and sulfate and DIC were not present. XRD scans of precipitates collected at the end of our experiments confirmed that powellite and calcite were present, but no gypsum was detected. Speciation calculations showed that a larger proportion of the aqueous Ca pool was     157  complexed as CaSO4(aq) in the CaCO3/SO42- series, in contrast to the CaCl2 series where Ca was overwhelmingly present as the free cation Ca2+ (Figure 5.1).   Figure 5.3. Molybdenum concentrations, δ98Mo, and powellite SI vs. time during powellite precipitation experiments.  CaCl2 series at left and CaCO3/SO42- series at right. Experimental conditions are summarized in Table 5.1. Note difference in x-axis scale for experiment POW2A compared to other graphs. In the “High-Mo” CaCO3/SO42- experiment POW2A, the Mo removal rate was faster than in the “Low-Mo” experiment POW2B2, analogous to observations from the CaCl2 series. In POW2A, the decline in Mo concentrations began after only 19 hours, and continued rapidly during the first 14 days, during which time > 90 % of the Mo was removed. However, the reaction slowed thereafter, and at 139 days and 98 % Mo removal, Mo concentrations appeared     158  to stabilize at 150 µM and with a powellite SI of 0.9 (Figure 5.2). In experiment POW2B2, there was a transient decrease of approximately 10 % in Mo concentrations between 3 and 4 days, after which time Mo concentrations returned to the starting value of approximately 950 µM and remained constant between 7 and 11 days (Figure 5.3). After 11 days, a more pronounced decline in Mo concentrations occurred, albeit at a slower rate than in the other powellite experiments. After 38 days, Mo removal had reached 75 %, and between 73 and 97 days Mo removal plateaued around 83 % and Mo concentrations of ~ 150 µM. The solution remained supersaturated with respect to powellite at this stage (SI > 1) (Figure 5.2). Similar to the CaCl2 series, in both CaCO3/SO42- experiments, δ98Mo rose alongside Mo removal, reaching 1.2 ‰ after 83 % of the Mo was removed in POW2B2, and 2.2 ‰ after 98 % of Mo was removed in POW2A. The maximal δ98Mo observed in the CaCO3/SO42- series was lower than that of the CaCl2 series. 5.3.2 Wulfenite precipitation Mineral precipitates formed immediately upon addition of Pb into Mo-bearing solutions, which changed from clear and transparent to a milky white. Molybdenum removal in all three wulfenite experiments followed a linear trend with increasing Pb addition (Figure 5.4 and Table 5.3), and XRD scans of the precipitates collected at the end of experiments confirmed that wulfenite had formed—and not other Pb-bearing minerals that could conceivably have precipitated on the basis of PHREEQC SI calculations [e.g. cerrusite, anglesite, or Pb(OH)2]. In experiment WULF1, the pH was moderately acidic (pH < 6), while the addition of NaHCO3 in     159  experiments WULF2 and WULF3 created circumneutral pH conditions (pH 6.6 to 7.7). The median alkalinity in experiment WULF2 was 0.87 mM, and 2.7 mM in WULF3. Values of δ98Mo increased with greater extents of Mo removal in all experiments, becoming up to 0.8 ‰ heavier than the starting Mo stock solutions (Figure 5.5).  Figure 5.4. Molybdenum removal from solutions with the addition of Pb in wulfenite precipitation experiments. Black line shows a 1/1 slope.   0.0000.0010.0020.0030.0040.0050.000 0.001 0.002 0.003 0.004 0.005mmol Mo removedmmol Pb addedWULF1WULF2WULF3    160   Figure 5.5. Increasing δ98Mo during Mo removal via wulfenite precipitation. Error bars represent ± 2 SD for replicate MC-ICP-MS analyses when available; or, for samples that were only analyzed once, the external 2SD reproducibility of the Mo(UBC) isotope standard is shown instead (0.07 ‰, n = 251).       161  Table 5.3. Geochemical data for samples drawn from wulfenite precipitation experiments Sample pH Alkalinity Mo Na SO42- fMo δ98Mo     ID   meq/L µM mM mM remaining ‰ ± 2 SDa n WULF1                   WULF1 - Mo stock -- -- 82 0.180 -- 1.00 0.07 0.02 3 WULF1-1 5.59 -- 67.1 0.1767 -- 0.81 0.1 0.1 3 WULF1-2 5.63 -- 53.1 0.177 -- 0.64 0.18 0.07 1 WULF1-3 5.41 -- 38.1 0.168 -- 0.46 0.28 0.07 1 WULF1-4 -- -- 22.2 0.172 -- 0.27 0.51 0.07 1 WULF1-5 -- -- 8.26 0.1854 -- 0.10 0.71 0.04 3 WULF2                   WULF2 - Mo stock 6.53 0.95 114 21.5 10.5 1.00 0.08 0.07 1 WULF2-1 6.76 0.84 89 21.2 10.3 0.78 0.15 0.07 1 WULF2-2 7.04 0.80 68.2 21.5 10.4 0.60 -- -- -- WULF2-3 6.70 0.77 59.6 22.1 10.8 0.52 0.46 0.07 1 WULF2-4 7.10 0.78 45.9 21.0 10.2 0.40 0.46 0.07 1 WULF2-5 6.55 0.75 36.2 21.4 10.4 0.32 0.58 0.01 2 WULF2-6 6.71 0.81 24.8 21.3 10.3 0.22 0.66 0.07 1 WULF2-7 6.86 0.87 16.6 23.0 11.1 0.14 0.73 0.07 1 WULF2-8 6.96 0.81 10.2 21.7 10.4 0.090 0.81 0.07 1 WULF2-9 6.70 0.83 4.5 22.1 10.7 0.040 0.9 0.2 3 WULF3                   WULF3 - Mo stock 7.37 3.1 111.0 24.0 10.5 1.00 0.1 0.1 3 WULF3-1 7.49 3.0 88 24.0 10.44 0.79 0.18 -- 1 WULF3-3 6.85 2.2 53 23.0 10.0 0.48 0.41 -- 1 WULF3-5 7.39 2.8 32.23 22.9 9.99 0.29 0.58 0.06 2 WULF3-6 7.39 2.6 21.0 21.7 9.4 0.19 0.65 -- 1 WULF3-7 7.63 2.8 12.13 22.2 9.6 0.11 -- -- -- WULF3-8 7.69 2.6 4.38 20.8 9.0 0.039 0.8 0.1 3 a The 2 SD for duplicate and replicate analyses is shown when available. For samples analyzed only once, the long-term 2 SD of 0.07 ‰ ( n = 251) for the Mo(UBC) isotope standard is shown instead and noted by italicized font.  -- indicates that the parameter was not measured.       162  5.4 Discussion 5.4.1 Molybdenum isotopic fractionation during powellite precipitation 5.4.1.1 Kinetic isotope effect (KIE) Kinetic isotope effects commonly result from mineral precipitation under supersaturated conditions (Watkins et al., 2017). Under these conditions, net precipitation rates are much greater than net dissolution rates, creating a reaction pathway that is essentially unidirectional (precipitation-dominated) and enriching the mineral phase in light isotopes due to their faster reaction rates (DePaolo, 2011). The δ98Mo data from powellite experiments were fitted using the Rayleigh model (Rayleigh, 1902), which applies to incomplete and unidirectional reactions under well-mixed and closed conditions (Wiederhold, 2015). A linear form of the Rayleigh equation (Jamieson-Hanes et al., 2017; Scott et al., 2004) can be used to determine the fractionation factor (α). ln(δf + 1000) = (α – 1) lnf + ln(δo + 1000)  (5.6) In equation (5.6), δf is the δ98Mo corresponding to a given fraction f of reactant (Mo) remaining in solution, and δ0 is the initial δ98Mo of the system. α is henceforth defined as the 98Mo/95Mo ratio in the mineral phase divided by the 98Mo/95Mo ratio of the aqueous Mo. The two standard error (2 SE) of α is given by propagating the 2 SE of the best-fitting slope of the linear regression of equation (6) (i.e. α – 1) through calculations of α (Jamieson-Hanes et al.,     163  2017; Kitchen et al., 2012). The fractionation factor can also be re-expressed for convenience in per mil units as an isotope enrichment factor (ε): ε = 1000 (α – 1)     (5.7) The Rayleigh model adequately described Mo isotope data during powellite precipitation, and yielded fractionation factors ranging from 0.99934 to 0.99945 (Figure 5.6 and Table 5.4). Data from samples collected later than 14 days in experiment POW3—after powellite approached chemical equilibrium—were excluded from Rayleigh curve-fitting as they do not adhere to the unidirectional reaction regime applicable in Rayleigh fractionation (Figure 5.3). Data from the initial transient decrease in Mo concentrations between 0.1 and 11 days in experiment POW2B2 (Figure 5.3) were also excluded, because they most likely represented short-term fluctuations in Mo concentrations during the initial nucleation period, rather than the onset of powellite precipitation. Kinetic fractionation factors in all powellite experiments were equal within error, regardless of the source of Ca and of the presence/absence of sulfate and DIC (Figure 5.6 and Figure 5.7). The average fractionation factor representing the KIE during powellite precipitation under supersaturated conditions is αpowellite-aqueous of 0.99936 ± 0.00004 (2 SD, n = 4), which corresponds to an isotopic enrichment factor (ε) of -0.64 ± 0.04 ‰ (Table 5.4).     164    Figure 5.6. Isotopic enrichment factors (ε) determined for kinetic isotope fractionation during powellite precipitation under supersaturated conditions. Error bars for each experiment represent ± 2 SE; error bar for average of all experiments represents ± 2 SD and is smaller than symbol size. Table 5.4. Summary of kinetic and equilibrium Mo isotopic fractionation factors in powellite precipitation experiments. Experiment System αmineral-aqueous ± 2 SE or SD† ε or ∆a(‰) ± 2 SE or SD† Kinetic fractionation           POW1 CaCl2 Low Mo 0.99934 0.00036 -0.66 0.36 POW3 CaCl2 High Mo 0.99938 0.00008 -0.62 0.21 POW2B2 CaCO3-SO42- Low Mo 0.99936 0.00007 -0.64 0.08 POW2A CaCO3-SO42- High Mo 0.99945 0.00010 -0.55 0.10 Kinetic - Average   0.99936 0.00004† -0.64 0.04† Equilibrium fractionationb         POW3 CaCl2 High Mo 0.99962 0.00006† -0.38 0.06† a ε is shown for kinetic systems; ∆ is shown for equilibrium system. b Equilibrium isotope fractionation behavior calculated from samples collected after 31 days in experiment POW3. See section 5.4.1.2 for details. † 2 SD is shown for average of multiple analyses, 2 SE is shown for individual experiments.       165  Several hypotheses can be considered to explain the mechanism causing KIEs during powellite precipitation. Isotopic fractionation between aqueous Mo species is possible (Tossell, 2005) and could cause fractionation during mineral precipitation if certain Mo aqueous species of distinct isotopic composition are preferentially precipitated relative to others. An analogous process was invoked to explain U stable-isotope fractionation during co-precipitation with CaCO3, wherein charged U species were preferentially precipitated relative to the uncharged and isotopically lighter aqueous U species Ca2UO2(CO3)3, which remained in solution (Chen et al., 2016). However, it is unlikely that isotopic fractionation between aqueous Mo species can explain the fractionation observed in our experiments: the differences in DIC and sulfate content between the CaCl2 experiments and the CaCO3/SO4 series introduced variety in solution chemistries and aqueous complexes involving the MoO42- oxyanion (Figure 5.1), and yet the kinetic isotope fractionation factors were equal within error between experiments. Consequently, the observed KIE was more likely caused during the incorporation of Mo into the solid phase, rather than isotopic fractionation between different aqueous complexes and preferential uptake of an isotopically fractionationated species into the solid phase.     166   Figure 5.7. Best-fit curves for Mo removal via powellite precipitation using the Rayleigh model. Experimental data are overlain (blue triangles). 2 SD error bars on δ98Mo data are typically smaller than symbol size. Experimental conditions are described in Table 5.1. Incorporation of Mo into mineral phases involves a series of steps, each of which could conceivably induce isotopic fractionation: (1) transport of MoO42- from the bulk solution to the mineral-solution interface; (2) diffusion of MoO42- from the solution to the mineral surface via a diffusion boundary layer (DBL); and (3) desolvation and transport of molybdate ions along the mineral surface to edges and kinks where they are incorporated into the growing crystal lattice (Lasaga, 1998; Watkins et al., 2017). Because all experiments were stirred, we can reasonably assume that bulk fluids were isotopically homogenous, which precludes fractionation from diffusive transport in the bulk solution (note that this excludes the DBL). The mechanism for     167  enrichment in light Mo isotopes within the precipitates therefore most likely resulted during transport through the DBL and/or during desolvation and attachment on the mineral surface. Desolvation of metal cations has been linked to kinetic isotope fractionation, for example during incorporation of Mg2+ into calcite, because faster desolvation kinetics for the lighter isotopes favors their attachment to the growing crystal lattice (Mavromatis et al., 2013). In the case of an oxyanion such as molybdate, a similar step is required: aqueous molybdate has a hydration shell of ~ 12 water molecules [MoO4(H2O)122-; Johansson and Caminiti 1986] which must be shed prior to incorporation into the molybdate mineral lattice and could also favor kinetic enrichment of light Mo isotopes in the mineral phase if 95MoO42- has faster desolvation kinetics than 98MoO42-. However, because hydrating H2O molecules are second-order neighbors in the molybdate oxyanion rather than first-order neighbors as in the cation-forming aqueous species [e.g. Mg(H2O)62+; Di Tommaso and de Leeuw 2009], the relative extent of isotopic fractionation from desolvation effects is expected to be smaller for an oxyanion species such as molybdate.  5.4.1.2 Equilibrium fractionation after advanced reaction progress In experiment POW3, the change in Mo isotopic composition after 20 days most likely reflected a switch from a kinetically controlled regime to conditions reflecting chemical equilibrium (Figure 5.3). The initial KIE is consistent with mineral precipitation from supersaturated solutions favoring faster reaction rates for the lighter isotopes, but as chemical equilibrium is approached the net precipitation rate decreases and the dissolution rate becomes     168  more important. This switch in reaction pathways favors bidirectional isotopic exchange between the fluid and mineral phases and opens the door to equilibrium isotope fractionation (DePaolo, 2011; Pearce et al., 2012). The progression from kinetic to equilibrium conditions was postulated to account for a similar behavior of Mg isotopes during experimental precipitation of MgCO3 (Pearce et al., 2012).  POW3 was the only one among all powellite experiments to exhibit equilibrium isotope fractionation behavior, presumably because its reaction progress was the greatest and it therefore remained near chemical equilibrium long enough to allow isotopic exchange between precipitated powellite and aqueous Mo. The greater reaction progress in POW3 is consistent with a higher initial degree of supersaturation and a higher initial Mo concentration in that experiment, leading to more rapid initial powellite nucleation and also providing a larger powellite surface area to catalyze further powellite precipitation.  Our data cannot ascertain whether isotopic equilibrium was fully established at late time in POW3, or whether the redistribution of isotopes upon approaching chemical equilibrium was limited an outer layer of atoms at the mineral surface exchanging with the solution (Guilbaud et al., 2011). If the latter condition prevailed, chemical equilibrium could have been maintained without necessarily achieving full isotopic equilibrium. In addition, isotopic zoning has been observed during rapid precipitation of hematite from Fe(III)-bearing solutions, indicating that solid-phase isotopic heterogeneity can be preserved during mineral precipitation under highly supersaturated conditions (Skulan et al., 2002). In that experiment, stepwise re-dissolution of isotopically heterogeneous hematite caused distinct Fe isotopic signatures. However, the     169  consistency in δ98Mo values for samples collected after 30 days in experiment POW3 suggests that isotopic heterogeneity within the re-dissolving powellite under near-(chemical)-equilibrium conditions was not important, and that the fluid and mineral were most likely at isotopic equilibrium. If this interpretation is assumed to be correct, the three samples in experiment POW3 that were collected near powellite chemical equilibrium (i.e. POW3-12; 3-13; 3-14) make it possible to estimate an equilibrium fractionation factor via isotopic mass balance: The δ98Mo of the precipitated powellite (δpowellite) is given by: 𝛿𝑝𝑜𝑤𝑒𝑙𝑙𝑖𝑡𝑒 = 𝛿0 − 𝛿𝑓  ×  𝑓1 − 𝑓   (5.8) where δ0 is the initial isotopic composition of the system, and δf is the isotopic composition of the residual aqueous Mo after a fraction f of Mo has been removed. The equilibrium isotopic fractionation factor (αpowellite-aqueous) can be derived using the relationship:    𝛼𝑝𝑜𝑤𝑒𝑙𝑙𝑖𝑡𝑒−𝑎𝑞𝑢𝑒𝑜𝑢𝑠 =𝛿𝑝𝑜𝑤𝑒𝑙𝑙𝑖𝑡𝑒 + 1000𝛿𝑎𝑞𝑢𝑒𝑜𝑢𝑠 + 1000  (5.9) where δaq is the δ98Mo of the remaining solution. These calculations yield an average equilibrium isotopic fractionation factor of 0.99962 ± 0.00006 (2 SD, Table E.3). This equilibrium isotopic fractionation factor is in the same direction but of smaller magnitude than the kinetic isotopic fractionation factor for powellite precipitation, similar to the results of Skulan et al. (2002), for Fe isotope fractionation during hematite precipitation, demonstrating that larger fractionations can result from KIEs as compared to equilibrium effects. However, further refinement of the powellite equilibrium isotope fractionation factor should be made as it can only     170  be calculated from three samples of a single experiment in our dataset in which isotopic equilibrium cannot be unequivocally confirmed. 5.4.1.3 Effect of sulfate and/or DIC inhibition on powellite precipitation Overall, a smaller proportion of Mo was removed in the CaCO3/SO42- series than in the CaCl2 series, indicating that either sulfate and/or the presence of DIC inhibit powellite formation (Figure 5.2). One hypothesis explaining this inhibition is that the presence of sulfate interferes with the formation of Ca2+–MoO42- electrostatic interactions and drives a kinetic limitation on powellite nucleation. This process may occur at the growing powellite crystal surface, or in aqueous solution, and effectively “poisons” lattice growth. Evidence for lattice-growth poisoning is present in other mineral precipitation reactions: For example, the presence of PO43- can decrease calcite precipitation rates, and block calcite precipitation altogether even under conditions of calcite supersaturation (Plant and House, 2002). Sulfate—as a tetrahedral and divalent aqueous oxyanion—is geochemically similar to MoO42- and can be plausibly expected to interfere with MoO42-’s molecular interactions. Analogous inhibition of powellite formation by DIC species (HCO3- and CO32-) is also possible, although less probable due to the greater dissimilarity between those anions and MoO42- than between sulfate and MoO42-.  An alternative hypothesis to explain the inhibition of powellite formation in the presence of sulfate and DIC is that an unknown aqueous complex binds Mo and/or Ca and makes them unavailable for further powellite precipitation. This would yield apparent supersaturation with respect to powellite if the unknown complex is not included in PHREEQC SI calculations. If this     171  hypothesis were true, our CaCO3/SO42- experiments could still have reached chemical equilibrium, despite the SIs calculated in PHREEQC that were > 0.8. However, neither experiments in the CaCO3/SO42- series saw a reversal of the KIE towards apparent isotopic equilibrium (as seen in CaCl2 series experiment POW3), even after durations exceeding 3 months and an apparent end in the reaction progress as indicated by the stabilization of Mo concentrations (Figure 5.2). This result suggests that chemical equilibrium was never reached in the CaCO3/SO42- experiments, and constitutes evidence against the hypothesis that an unknown aqueous complex caused the inhibition of precipitation. 5.4.2 Molybdenum isotopic fractionation during wulfenite precipitation Molybdenum isotopic fractionation during wulfenite precipitation followed the same direction as for powellite precipitation, leading to an enrichment of heavy isotopes in solution. The data collected in wulfenite experiments were fit by linear regression using both the Rayleigh [Equation (5.6)] and Equilibrium [Equation (5.10)] models:   𝛿𝑓 = 𝛿0 − 𝑓 𝛥𝑀𝑜𝑤𝑢𝑙𝑓𝑒𝑛𝑖𝑡𝑒−𝑎𝑞𝑢𝑒𝑜𝑢𝑠   (5.10) In Equation (5.10), δf is the δ98Mo value for a given fraction f of reactant (Mo) removed from solution, δ0 is the initial δ98Mo of the system, and Δ is the isotopic separation factor, defined as the difference in isotopic composition between wulfenite and aqueous molybdate: Δ98Mowulfenite-aqueous = δ98Mowulfenite – δ98Moaqueous   (5.11)     172  The equilibrium fractionation factor (αwulfenite-aqueous) can be derived from the isotopic separation factor by the following approximation, which is valid for relatively small Δ-values (< 20 ‰) (Clark and Fritz, 1997) such as those encountered in metal stable-isotope systems: Δ98Mowulfenite-aqueous ≈ 103ln αwulfenite-aqueous    (5.12) The 2 SE uncertainty of the equilibrium fractionation factor was calculated analogously to the approach used in the Rayleigh model, wherein standard error on the linear regression’s slope was doubled and then propagated through calculations of α. In contrast to powellite formation, wulfenite chemical equilibrium is achieved almost instantaneously in Pb- and Mo-bearing aqueous solutions (Conlan et al., 2012; Torres et al., 2014), which creates favorable conditions for isotopic exchange between the precipitated wulfenite and the residual solution as a result of equal dissolution and precipitation rates at chemical equilibrium. Indeed, the Mo isotope fractionation during wulfenite precipitation was better described using equilibrium fractionation, although overprinting of KIEs was possible (Figure 5.8). Differentiating between equilibrium and Rayleigh fractionation processes is most easily done at high extents of Mo removal (> 95 %); the available data at extreme Mo removal in experiments WULF2 and WULF3 fall on the Equilibrium fractionation line (Figure 5.8b and Figure 5.8c), while similar data are not available for experiment WULF1 (Figure 5.8a). In the latter experiment, Equilibrium and Rayleigh isotope fractionation models provide reasonable fits to the data (Figure 5.8a.), with r2 values of 0.92 and 0.98, and a p-value of < 0.01 as determined from the Pearson’s Chi-squared Goodness of Fit test. However, it is geochemically unlikely that     173  experiment WULF1 had followed Rayleigh fractionation, and not experiments WULF2 and WULF3. WULF1 represented the geochemically simplest system (wulfenite precipitation in the absence of complexation effects from DIC and sulfate), and can therefore plausibly be expected to have reached chemical equilibrium earlier than the other experiments and had more time for isotopic exchange between solid and fluid phases.     Figure 5.8. Molybdenum isotope fractionation during wulfenite precipitation.  Best-fitting curves for the Rayleigh model are shown with solid lines, and equilibrium model with dashed lines. Best-fitting Rayleigh and equilibrium model fractionation factors (αwulfenite-aqueous) for each experiment are: (A) 0.99971 ± 0.00007 and 0.99929 ± 0.00032, respectively, for WULF1; (B) 0.99975 ± 0.00020 and 0.99916 ± 0.00019, respectively, for WULF2; and (C) 0.99979 ± 0.00021 and 0.99929 ± 0.00019, respectively, for WULF3. The average equilibrium αwulfenite-aqueous for all three experiments is 0.99925 ± 0.00015. The best-fitting isotope separation factors for WULF1, WULF2, and WULF3 were -0.71 ± 0.16, -0.84 ± 0.10, and -0.71 ± 0.10 ‰, respectively (Table 5.5). Since these separation factors are equal within error (Figure 5.9), it is possible to calculate their average to represent equilibrium isotope fractionation for wulfenite. This average corresponds to Δ98Mowulfenite-aqueous = -0.75 ± 0.15 ‰ (2 SD, n = 3), or αwulfenite-aqueous of 0.99925 ± 0.00015.      174  Table 5.5. Equilibrium Mo isotope fractionation factors in wulfenite precipitation experiments. Experiment αmineral-aqueous ± 2 SE or SD† ∆mineral-aqueous (‰) ± 2 SE or SD† WULF1 0.99929 0.00016 -0.71 0.16 WULF2 0.99916 0.00010 -0.84 0.10 WULF3 0.99929 0.00010 -0.71 0.10 Wulfenite - Average 0.99925 0.00015† -0.75 0.15†   Figure 5.9. Isotopic separation factors (∆98Mo) for equilibrium isotope fractionation during wulfenite precipitation. The error bars for each experiment represent ± 2 SE, and the error bar for average of all experiments represents ± 2 SD. The lack of significant change in αwulfenite-aqueous among wulfenite experiments, despite their different solution chemistries, suggests that isotopic fractionation between aqueous Mo species was not responsible for causing the overall fractionation. The mechanism of Mo isotopic fractionation during wulfenite precipitation is therefore probably caused by differences in the distortion of Mo—O bonds in aqueous molybdate and wulfenite and aqueous Mo, with heavier isotopes being enriched in the aqueous phase. The enrichment of light Mo isotopes in the mineral phase is also consistent with results from the powellite experiment POW3, after chemical     175  equilibrium was approached. Further molecular-scale study of these compounds is required to determine the exact mechanism of isotopic fractionation between solution and molybdate minerals, although similar results have been described in other metal stable-isotope systems at isotopic equilibrium, such as Fe (Dauphas et al., 2017), Mg (Pearce et al., 2012), and stable Sr (Liu et al., 2016). 5.4.3 Molybdate removal mechanisms and environmental implications The application of Mo stable isotopes to gain insight into controls on Mo transport in natural and contaminated environments depends upon knowledge of the mechanisms and magnitude of Mo isotope fractionation. This study demonstrates that molybdate mineral precipitation consistently leads to preferential removal of light Mo isotopes, and can therefore leave a traceable isotopic signature in the environment. Studies of other aqueous Mo removal processes in natural and laboratory settings also point to isotopic fractionations that are in the same direction, including during Mo adsorption and co-precipitation onto Fe- and Mn-(oxyhydr)oxide and Al-(hydroxy)sulfate minerals in oxic environments (Barling and Anbar, 2004; Goldberg et al., 2009; Skierszkan et al., 2017; Wasylenki et al., 2008), and precipitation of Fe-Mo sulfides under sulfidic conditions (Nägler et al., 2011). As a general rule of thumb, rising δ98Mo in water therefore reflects Mo removal, and the use of Mo stable isotopes in conjunction with other geochemical data (e.g. oxydo-reduction potential, pH, solution chemistry) constitutes a powerful means to identify and quantify the removal process(es) at play.     176  Compared with Mo adsorption, isotopic fractionation factors for powellite and wulfenite precipitation are smaller (Table 5.6 and Figure 5.10). This observation can be explained by a larger change in the chemical bonding environment of Mo during adsorption, which involves a switch from tetrahedral coordination in aqueous Mo to octahedral or mixed tetrahedral/octahedral coordination on the surface of (oxyhydr)oxide minerals (Kashiwabara et al., 2011; Wasylenki et al., 2011). The magnitude of Mo isotope fractionation during adsorption is primarily driven by the proportion of octahedrally coordinated Mo (Goldberg et al., 2009; Kashiwabara et al., 2017, 2011). In contrast, the Mo atom in wulfenite and powellite remains as a tetrahedrally coordinated MoO42- oxyanion, similar to aqueous Mo, and Mo—O bonds are of similar length in the aqueous and mineral phases (Achary et al., 2006; Cora et al., 2011; Johansson and Caminiti, 1986; Lugli et al., 1999). The similarity in bonding environments between those two phases theoretically reduces the opportunity for equilibrium isotopic fractionation (Schauble, 2004). However, the presence of a KIE during powellite precipitation can still cause large (> 3 ‰) Mo isotope fractionation in the residual aqueous Mo pool at high extents of Mo removal. Regardless of the fractionation behavior (kinetic or equilibrium) of the powellite and wulfenite precipitation it is readily resolvable by MC-ICP-MS.       177  Table 5.6. Summary of known Mo isotopic fractionation factors for Mo removal processes from aqueous solutions Process αsolid-solution ±a Ref. Adsorption       magnetite 0.99917 0.00060 1 ferrihydrite 0.99889 0.00015 1 goethite 0.99860 0.00048 1 hematite 0.99781 0.00054 1 Mn-(oxyhydr)oxide (δ-MnO2) 0.99731 0.00075 2 Mn-(oxyhydr)oxide (birnessite) 0.99735 0.00006 3 Humic acid, pH 4.2 0.99861 0.00016 4 Co-precipitation and/or adsorption     Fe-thiomolybdateb 0.99950 0.00030 5 Basaluminite 0.99925 0.00015 6 Molybdate mineral precipitation     powellite, kinetic 0.99936 0.00004 7 powellite, equilibrium 0.99962 0.00006 7 wulfenite, equilibrium 0.99929 0.00015 7 a Because uncertainties in different studies are presented using a variety of notations, the uncertainty reported here corresponds to the difference between the mean reported α-value and the α-value to which the reported uncertainty was added. When fractionation factors were reported using the isotopic separation notation (∆), corresponding α-values were calculated using the approximation ∆ ≈ 103lnα. The fractionation factors from studies using the δ97/95Mo notation were re-expressed in terms of δ98/95Mo. b Mo removal by Fe-thiomolybdates in sulfidic marine basins. References: 1 Goldberg et al. (2009); 2 Barling and Anbar (2004); 3 Wasylenki et al. (2008); 4 King et al. (2018);  5 Nägler et al. (2011); 6 Skierszkan et al. (2017) and Chapter 4 of this thesis.       178   Figure 5.10. Fractionation factors for Mo removal processes from aqueous solutions. All fractionation factors are < 1, which points to preferential removal of light Mo isotopes during Mo attenuation. “eq” refers to an equilibrium isotope effect, “kin” refers to a kinetic isotope effect. Adsorption (warm colors) yields larger fractionation factors than precipitation of powellite and wulfenite (blue symbols). Note that it is unclear in prior studies whether Mo scavenging by Fe-thiomolybdates and Al-(hydroxy)sulfate is via surface adsorption or co-precipitation. References: 1 King et al. (2018), 2 Goldberg et al. (2009), 3 Barling and Anbar (2004), 4 Wasylenki et al. (2008), 5 Nägler et al. (2011), 6 Skierszkan et al. (2017), 7 this study. Molybdate minerals can constitute solubility controls on Mo concentrations under alkaline-pH conditions, which are unfavorable for MoO42-(aq) adsorption (Conlan et al., 2012; Goldberg et al., 1996). Powellite forms in Mo-rich alkaline mine drainage at sites with MoS2 mineralization (e.g. Blanchard et al., 2015; Conlan et al., 2012; Hayes et al., 2014), and in natural and alkaline aquifers bearing a geogenic source of Mo (Pichler and Mozaffari, 2015). However, the rate and magnitude of powellite precipitation were significantly inhibited in our CaCO3/SO42- experiments, which did not lower Mo concentrations below 140 µM. A comparable inhibition of powellite precipitation in the presence of DIC and sulfate was also observed in similar experiments by Conlan et al. (2012). In contrast, the Mo concentrations in most natural     179  waters are orders of magnitude lower (Smedley et al., 2017) and the recommended maximum Mo concentration in drinking water is orders of magnitude lower, at 0.7 µM (World Health Organization, 2011). The inhibitory effect of DIC and/or sulfate on powellite formation imposes limitations on the viability of powellite precipitation as a Mo-attenuation mechanism and should be investigated further, as its precise cause is unresolved by the experiments herein. The faster reaction kinetics and lower solubility of wulfenite makes the precipitation of this mineral a more effective removal mechanism for Mo than powellite, even in the presence of anions competing for Pb such as sulfate and bicarbonate (Hirsche et al., 2017). Wulfenite occurs in Pb- and Mo-bearing mine waste (Conlan et al., 2012; Miler and Gosar, 2012; Petrunic et al., 2009), and should be suspected in environments where there is a source of Pb in contact with Mo-bearing solutions. Because both kinetic and equilibrium isotope fractionations were observed during powellite precipitation experiments, the use of fractionation factors under environmental conditions must take into account the dominant fractionation mechanism at play. In solutions that are highly supersaturated with respect to powellite, the KIE is likely to dominate (DePaolo, 2011). Equilibrium isotope effects were also not seen for powellite precipitation in the presence of DIC and sulfate, suggesting that the KIE will likely dominate in environments such as sulfide and carbonate-rich mine waste rock and tailings weathering. Where flow is very slow (i.e. longer residence time of supersaturated pore waters than the time required to reach chemical equilibrium with respect to powellite) equilibrium isotope fractionation may be more applicable. In contrast, the short timeframe (< 2 days) required to establish chemical and isotopic     180  equilibrium with respect to wulfenite precipitation indicates that the equilibrium fractionation factor is likely to be relevant to describe wulfenite precipitation under most environmental conditions. Finally, it should be also noted that in complex environments with dynamic flow fields, isotopic variations are different than under laboratory batch experiment conditions. For example, hydrodynamic mixing of flowpaths, some containing isotopically fractionated and others unfractionated metals, can decrease the magnitude of observed isotopic fractionation under active flow conditions (Druhan and Maher, 2017; Jamieson-Hanes et al., 2012). In contrast, extreme isotopic fractionation in a fluid reservoir might be expected if a flowpath undergoes progressive unidirectional metal attenuation at each transport step. The inclusion of metal isotope fractionation factors in reactive-transport models constitutes a field of active research that can help address the additional complexity introduced by transport processes under active-flow conditions and bridge the gap between laboratory and field-based studies (Jamieson-Hanes et al., 2012; Wanner et al., 2014).  5.5 Summary and conclusions Powellite (CaMoO4) and wulfenite (PbMoO4) are well-suited to the study of isotopic fractionation during mineral precipitation as a result of their contrasting reaction kinetics and the ease with which they can be precipitated under laboratory conditions. Laboratory batch experiments at 20°C demonstrated that the precipitation of these minerals from MoO42--bearing aqueous solutions causes a rise in δ98Mo in the remaining aqueous Mo pool. Powellite     181  precipitation from supersaturated conditions exhibited kinetic isotope fractionation that was described by an average fractionation factor of 0.99936 ± 0.00004 (2 SD, n = 4). That fractionation factor was not significantly changed by the presence/absence of sulfate and DIC. However, experiments that included sulfate and DIC saw a decrease in the magnitude and rate of powellite precipitation. In the absence of sulfate and DIC, powellite mineral saturation was approached after many weeks, which coincided with a switch from kinetic towards equilibrium isotope fractionation.  Wulfenite precipitation was better described by equilibrium isotope fractionation with αmineral-aqueous = 0.99929 ± 0.00015 (2 SD, n = 3). Wulfenite precipitation can also attenuate Mo concentrations to much lower levels, and within a much shorter timescale, as compared to powellite precipitation, which requires weeks to months to reach chemical equilibrium. Determination of the Mo isotopic fractionation factors for powellite and wulfenite precipitation contributes critical knowledge for tracing Mo removal from oxic, aqueous solutions in the environment. At present, all aqueous molybdenum removal processes (e.g. precipitation, adsorption) for which fractionation factors are known indicate that δ98Mo increases as a function of Mo removal (Kendall et al., 2017). Determining which removal processes are at play in a given environment therefore requires complementary knowledge of geochemical concentrations, pH, and redox conditions. Special care should also be given to account for hydrodynamic mixing and various source and sinks which introduce additional complexity in the interpretation of isotopic data under field conditions. Nonetheless, the growing documentation of isotopic fractionation caused by various Mo removal reactions make isotopic analyses a promising     182  avenue for the provision of quantitative constraints on metal attenuation in natural and contaminated environments.     183  CHAPTER 6. MOLYBDENUM ATTENUATION MECHANISMS IN VARIABLY OXIC MINE TAILINGS DRAINAGE: INSIGHTS FROM Mo STABLE ISOTOPES AND X-RAY SPECTROSCOPY 6.1 Introduction Metal release during weathering of sulfide-rich mine waste is a key environmental concern at mine sites. Molybdenum is a redox-sensitive element that can reach toxic concentrations in water near mine sites hosting Mo mineralization (Goumih et al., 2013; Kaback and Runnells, 1980; Smedley and Kinniburgh, 2017). Ruminants are particularly sensitive to Mo toxicity: their exposure to excessive Mo can cause molybdenosis, a harmful condition of Cu-deficiency. The maximum recommended concentration of Mo in drinking water for human consumption is 0.73 µM (70 µg/L, World Health Organization, 2011). To prevent adverse environmental effects of Mo leaching in mine waste, a detailed understanding of the mechanisms of Mo attenuation is required. Oxidation of residual molybdenite (MoS2) in mine waste releases aqueous molybdate (MoO42-) following the reaction:  2MoS2 + 9O2 + 6H2O → 2MoO42- + 4SO42- + 12H+   (6.1) A first important process of molybdate attenuation is its adsorption onto a variety of minerals, including (oxyhydr)oxides of Al, Mn, and Fe; clays; and pyrite (Bostick et al., 2003; Goldberg et al., 1996). These sorbent minerals are common in mine waste (Blowes et al., 2014). Adsorption of Mo is highly pH-dependent: it is strongest under moderately acidic conditions (pH     184  3 to 5) where Mo forms strong inner-sphere complexes; as pH rises above ~ 8, Mo forms only weaker outer-sphere complexes and adsorption becomes minimal (Goldberg et al., 1996; Goldberg and Forster, 1998; Gustafsson and Tiberg, 2015). Molybdenum stable-isotope fractionation can be used to indicate Mo adsorption because light Mo isotopes are preferentially removed from solution during adsorption onto mineral surfaces and organic matter, which leads to enrichment of heavy isotopes in aqueous Mo (Barling and Anbar, 2004; Goldberg et al., 2009; King et al., 2018; Wasylenki et al., 2008). This process drives the Mo in mine waste-rock drainage towards isotopically heavy values (Skierszkan et al., 2017, 2016). However, the molecular-scale mechanisms behind this isotopic fractionation on adsorption mineral surfaces have not been explained by analysis of environmental samples: spectroscopic and isotopic studies of Mo adsorption available at present are generally limited laboratory experiments on synthetic sorbent minerals [e.g. Arai (2010); Gustafsson and Tiberg (2015); Kashiwabara et al., 2009; 2011; 2017; Wasylenki et al., 2008; 2011)]. Further evaluation of the mechanism of Mo adsorption under environmental conditions, where mixed assemblages of mineral phases are likely to be present and have different surface chemistry than pure phases synthesized in the laboratory, is necessary to constrain the sorption behavior of Mo. A second Mo attenuation process in mine waste is the precipitation of molybdate minerals such as powellite [CaMoO4], ferrimolybdite [Fe2(MoO4)3], wulfenite [PbMoO4], and nickel(II) molybdate [NiMoO4] (Bissonnette et al., 2016; Blanchard et al., 2015; Essilfie-Dughan et al., 2011; Hayes et al., 2014; Hirsche et al., 2017). The precipitation of powellite and wulfenite     185  leads to progressive enrichment of heavy Mo isotopes in solution as a function of Mo removal (Chapter 5), indicating that molybdate precipitation can also be traced by Mo isotopic analysis. In sulfidic aquatic environments, precipitation of (oxo)thiomolybdate (MoOxS4-x2-) complexes with Fe(II) and organic matter (Helz et al., 1996) provides a third Mo attenuation pathway. (Oxo)thiomolybdate precipitation also preferentially removes light Mo isotopes from solution (Nägler et al., 2011), although isotopic fractionation factors for this process have not yet been determined experimentally. This process provides a redox-driven pathway for aqueous Mo removal, although it has only been studied in marine and lacustrine settings thus far (Dahl et al., 2010; Nägler et al., 2011). Although the aforementioned studies suggest possible controls on the mobility of Mo in mine waste, there have been relatively few studies of the behavior of this element in mine tailings from sulfide ore deposits (Goumih et al., 2013; Langedal, 1997). When compared to mine waste rock, mine tailings typically have finer grain size and often include dissolved organic solvents used in mineral processing, which can promote reducing and anoxic conditions (Lindsay et al., 2015). Reducing conditions can drive changes in the mobility of Mo through the reductive dissolution or phase transformation of (oxyhydr)oxides of Fe and Mn, which are important sorption surfaces for metals in oxic mine waste (Lindsay et al., 2015; Qin, 2016; Skierszkan et al., 2017). Because Mo sorption onto (oxyhydr)oxides of Fe and Mn causes isotopic fractionation (Barling and Anbar, 2004; Goldberg et al., 2009; King et al., 2018; Wasylenki et al., 2008), a change in the mobility and isotopic composition of Mo in mine drainage might be expressed in reducing mine waste facilities where Fe and Mn (oxyhydr)oxide sorption surfaces are     186  unavailable. In comparison to oxic environments, a decrease in sorption site availability could enhance Mo leaching and yield smaller isotopic fractionation. Other pathways for isotopic fractionation under reducing conditions could include the precipitation of powellite or thiomolybdate phases or Mo adsorption onto phases such as clays and pyrite that are stable in the absence of oxygen. The primary objectives of this work were (1) to qualitatively indicate the extent of Mo attenuation in variably oxic mine tailings drainage by application of Mo stable isotope analyses; and (2) to examine the molecular mechanisms of Mo sorption in mine drainage using XAS. The study site features a molybdenite mine tailings management facility that discharges anoxic water, which has elevated Fe(II) and is saturated with respect to powellite, into a seepage collection pond (Figure 6.1). In this pond, Fe(II) oxidizes and widespread Fe-(oxyhydr)oxide precipitation occurs, providing a well-constrained setting to examine Mo sorption under field conditions. The Mo isotope systematics in source minerals and waste-rock drainage from the same site have recently been characterized (Skierszkan et al., 2017), providing a basis to compare Mo attenuation in mine tailings drainage to that in waste-rock drainage. The results presented herein provide constraints on the processes that control Mo mobility in mine tailings using Mo isotopes as a tracer of attenuation. Special consideration is also given to the mechanism of Mo sorption onto Fe-(oxyhydr)oxides.       187   Figure 6.1. Conceptualized cross-section of the Bruno Creek Tailings Management Facility showing the major flowpaths and design features of the facility (modified from Lorax, 2011b). Cyclone separators at the crest of the tailings dam discharge tailings sands and process water as a slurry on the dam, and tailings slimes and process water into the tailings pond. A drain is constructed at the base of the facility to maintain unsaturated conditions within the embankment sands. Anoxic process water seeps through the water-saturated tailings slimes below the tailings pond. Process water drains through the water-unsaturated embankment sands. A small component (~ 5 % of flow) of groundwater also enters the BCTMF. All water is channeled towards the Rock Toe, and is then discharged into the Seepage Return Dam pond. A pump-back system below the SRD captures any groundwater leakage and returns it either to the SRD pond, or sends it to the mill or tailings pond for use in the ore extraction process. Localized zones of acidic pH exist within the sands near the base of the tailings embankment. Note that the drawing is not to scale.       188   Figure 6.2. Photographic overview of the study site.  Photo A shows the full extent of the Bruno Creek Tailings Management Facility (BCTMF) as seen in Google Earth. Inset shows the site location. The crest of the BCTMF dam is several hundred meters long and more than 200 m tall, and the tailings pond is visible behind the dam. Photo B shows the Rock Toe, where basal seepage exits the BCTMF. Photo C shows the Seepage Return Dam (SRD) pond, which collects basal seepage from the BCTMF.      189  6.2 Study site 6.2.1 Bruno Creek Tailings Management Facility The Bruno Creek Tailings Management Facility (BCTMF), located at the Thompson Creek Molybdenum (TCM) Mine in the Central Rocky Mountains, Idaho, USA, in contained by the world’s second-tallest centre-line tailings dam at more than 200 m in height (Figure 6.2; Golder Associates, 2018). The geological setting, hydrology, and climate at the mine are described in Skierszkan et al. (2017). Between 1982 and 2014, ore was crushed and milled on site and molybdenite was subsequently extracted by a flotation process. Flotation tailings were partitioned using a cyclone separator: the coarser-grained tailings sands were deposited along with process water as slurry to build a centre-line dam while the finer-grained tailings slimes were discharged into a tailings pond upgradient of the dam. The cyclone separator was exclusively operated during ice-free months; in winter the undifferentiated tailings sands and slimes were discharged together in the pond. Beginning in 1997, a pyrite concentration step was added to decrease the amount of pyrite entering the embankment and mitigate the potential for acid-rock drainage. Mine tailings at the TCM comprise quartz, feldspar, and plagioclase, with lesser amounts of pyrite, molybdenite, calcite, and other sulfide and aluminosilicate minerals (Lorax Environmental Services Ltd., 2011b). 6.2.2 Hydrology and geochemistry of the BCTMF A conceptual cross-section of the BCTMF shows its major design and hydrological features (Figure 6.1). Hydrology of the facility is governed by grain size distribution, with     190  coarser sands found in the embankment while finer slimes and whole tailings are deposited upgradient in the tailings pond. The sands that form the embankment are generally water-unsaturated; any process water that was discharged on the embankment during dam-building rapidly flows to an underdrain at the base of the dam, the "Rock Toe", which was constructed to prevent water saturation of the embankment. The finer-grained slimes in the pond are generally water-saturated. Consolidation of tailings slimes and the local hydraulic gradient cause seepage of water from these slimes to the base of the dam. The Rock Toe collects all flow within the BCTMF, which enters a pond held behind a Seepage Return Dam (SRD). The water budget at the Rock Toe is dominated by tailings slime drainage and process water, with a smaller amount of drainage through the sands and a minor component (5 %) of natural groundwater discharge (Lorax Environmental Services Ltd., 2011b). Below the SRD, groundwater seepage is returned to the SRD using a pump-back system to prevent downstream water contamination. Water from the tailings pond and the SRD can also be returned to the mill for use in the ore extraction process. Aqueous geochemistry of process water undergoes considerable redox evolution along the flowpath from the tailings pond through the tailings embankment into the SRD pond. Anoxic conditions prevail in the shallow sediments in the tailings pond slimes as a result of the oxidation of organic solvents (used in molybdenite extraction) and residual sulfide minerals. This oxidation drives a depletion of NO3- and release of Fe(II) and Mn(II) via reductive dissolution of (oxyhydr)oxide minerals. H2S production from sulfate reduction has also been observed locally in sediments at the bottom of the tailings pond (Lorax Environmental Services Ltd., 2011b).     191  However, sulfide was not detected in piezometers screened within the tailings sands at the base of the embankment, and sulfate concentrations in excess of 9 mM at the Rock Toe suggest that sulfate reduction only removes a minor amount of the total available sulfate. The Rock Toe seepage and piezometers screened at the base of the embankment historically contained tens to hundreds of µM of Fe(II) and Mn(II) (Lorax Environmental Services Ltd., 2011b). Upon discharge of Rock Toe seepage into the SRD pond, widespread precipitation of ochreous sludge—presumably Fe-(oxyhydr)oxide—occurs (Figure I.1). Historical data indicate that the median Mo concentration decreased by approximately 50 % between the tailings pond and the Rock Toe (Lorax Environmental Services Ltd., 2011b). Rock Toe seepage has remained close to powellite saturation, suggesting that precipitation-dissolution of this mineral may attenuate Mo (Lorax Environmental Services Ltd., 2011b). Boreholes drilled into the BCTMF in 2009 generally showed circumneutral-pH conditions, with the exception of localized acidic zones near the bottom of the embankment in an area where sulfide-rich tailings were deposited prior to the implementation of the tailings depyritization process (Lorax Environmental Services Ltd., 2011b). Shake-flask extractions of borehole samples indicated that Mo was generally mobile within the tailings embankment, with the exception of the acidic-pH samples: those samples did not release Mo, probably as a result of stronger Mo sorption under acidic conditions (Lorax Environmental Services Ltd., 2011b).  Baseline characterization of Mo isotopic compositions of solid mine waste and waste-rock drainage at the TCM are available in Skierszkan et al. (2017). The average δ98Mo of solid mine waste (including ore minerals, tailings, and waste rock) is 0.7 ‰. Mine tailings have Mo     192  contents ranging from 27 to 135 ppm and δ98Mo of 0.2 ± 0.2 ‰. Mine process water has a δ98Mo of 0.6 ‰. The two waste-rock storage facilities (WRSFs) at the TCM are characterized by isotopically heavy drainage ranging 1.6 to 1.9 ‰ as a result of Mo sorption onto Fe-(oxyhydr)oxide minerals.  6.3 Sample collection and analyses 6.3.1 Field sampling Water samples were collected from the outflow of the Rock Toe and from the SRD pond during the 2014 and 2015 baseflows, and during the 2016 freshet. Detailed sampling methods are described in Skierszkan et al. (2017) and Appendix H. Samples of the ochreous precipitate forming below the Rock Toe and in the SRD pond were collected into containers and immediately frozen in the field using dry ice (Figure I.2). In addition to grab samples, the sample “SRD-W1 PPTS” consists of Fe-(oxyhydr)oxides suspended in the turbulent water discharging into the SRD from a culvert below the Rock Toe. For this sample, water and suspended solids were collected together into a polypropylene bottle. The suspended particles in this mixture were extracted by vacuum filtration in the field using a 0.45-µm filter. The recovered solids were frozen in the field, while the filtered solution was preserved with sub-boiled HNO3 for metals and Mo isotopic analyses. In the laboratory, ochreous precipitate samples were thawed and centrifuged to isolate them from supernatant water, which was discarded. The ochreous precipitate samples were rinsed once in > 18.2 mΩ H2O, and centrifuged again with the supernatant discarded. The     193  samples were then re-frozen, and freeze-dried. Subsamples of this dry material were removed for digestions to determine their elemental abundance and Mo isotopic composition. Digestions were conducted using a MARS6 OneTouch Microwave Digestion System (CEM Corporation, Matthews, North Carolina, USA) and a combination of HCl-HF-HNO3 that was a 5x scale-up of the protocol of Axelsson et al. (2002). The acids used for sample digestion were purified in-house from concentrated reagent-grade acids by sub-boiling distillation. 6.3.2 Molybdenum sorption experiment A Mo sorption experiment was conducted on Rock Toe discharge water to capture the process of Mo attenuation during precipitation of Fe-(oxyhydr)oxides from this Fe(II)-rich tailings drainage. For this experiment, approximately 20 L of Rock Toe water was collected into a HDPE pail in the field and then allowed to oxidize over a period of 6 weeks during which time water samples were periodically withdrawn for chemical analyses (Figure I.3). During the field campaign, the pail was kept outdoors and in the shade at ambient temperatures, which varied between approximately -5 °C and +5 °C. Upon returning to the laboratory, the pail was refrigerated at 4 °C to simulate temperatures at the field site. The water in the pail was stirred during sampling events to ensure that water samples represented the bulk fluid. At the end of the experiment, the water was siphoned off and the settled precipitates were poured into a Nalgene bottle. These precipitates (samples B-1A and Mn-Buck) were then frozen, and extracted by centrifugation and freeze-drying as described previously for other ochreous precipitate samples.     194  6.3.3 Analytical methods 6.3.3.1 Elemental abundance and Mo isotopic composition Details explaining the analytical methods that were used for the determination of elemental abundances are described in Appendix H. Dissolved anions and sulfide were determined by ion chromatography and colorimetry, respectively. Metals contents of water and ochreous precipitate samples were analyzed by inductively coupled plasma mass spectrometry (ICP-MS, Agilent 7700x) and inductively coupled plasma optical-emission spectroscopy (ICP-OES, Varian 725-ES). Instrumental replicate and full duplicate analyses of samples and reference materials indicated that ICP-OES and ICP-MS analyses were generally precise and accurate within 6 % (Skierszkan et al., 2017). In ochreous precipitate samples, total carbon and nitrogen contents were determined using an elemental analyzer (Elementar Americas, Mt. Laurel, New Jersey). Total inorganic carbon (TIC) was determined by CO2 coulometry (UIC Coulometrics Inc., Joliet, Illinois). Total organic carbon was calculated as the difference between total carbon and TIC.  Molybdenum isotopic analyses of water and ochreous precipitate samples were completed with multi-collector ICP-MS (MC-ICP-MS, Nu 21, Nu Instruments Ltd., Wrexham, UK) as per Skierszkan et al. (2015, 2017). These analyses were carried out at the Pacific Centre for Isotopic and Geochemical Research (PCIGR), University of British Columbia, Vancouver, Canada. Briefly, a97Mo–100Mo double spike was added to sample aliquots to correct for laboratory and instrumental mass fractionation and Mo was extracted from the sample matrices     195  by anion-exchange chromatography using BioRad AG®-MP-1M resin (100–200 mesh). All Mo isotope data in this study are reported as δ98Mo relative to NIST-SRM-3134 = + 0.25 ‰ (Nägler et al., 2014): δ98Mosample (‰) = ( (Mo98Mo95 )sample(Mo98Mo95 )NIST-SRM-3134-1) × 1000  (6.2)  MC-ICP-MS accuracy was verified by analyzing at least one reference material (BCR-2, SDO-1, Nod-P-1, or seawater) along with each sample batch through the full ion-exchange and analytical procedure. Reproducibility was verified with replicate analyses of samples, which was always better than 0.1 ‰ (Table I.2). The geochemical code PHREEQC (v. 3.0.6-7757; Parkhurst and Appelo, 2013) was used with the wateq4f database for aqueous speciation and mineral saturation index calculations. Solubility constants for schertmannite (Bigham et al., 1996), powellite, and wulfenite (c.f. Conlan et al., 2012) were added to the wateq4f database. Absolute charge balance errors calculated in PHREEQC were on average 3 %, and always better than 10 %.  6.3.3.2 Ochreous precipitate mineralogical characterization The mineralogy and morphology of ochreous precipitate samples was determined using X-ray diffraction (XRD), Raman spectroscopy, and transmission-electron microscopy (TEM) (Appendix H). X-ray diffractograms were collected using either a Bruker D8 Focus or D8 Advance Bragg-Brentano diffractometer. Raman spectroscopy provides an alternative to XRD     196  for phase identification in Fe-(oxyhydr)oxide samples of poor crystallinity (Das et al., 2016; Das and Hendry, 2011; Hanesch, 2009). Raman spectra were collected using a Renishaw InVia Raman microscope equipped with a 785-nm solid-state laser. The instrument was calibrated prior to sample analyses using an internal Si standard measured at a Raman shift of 520 cm-1. Low laser powers of 0.1 to 0.5 % were used and spectra were monitored during acquisitions to avoid laser-induced phase transformations (e.g. transformation of ferrihydrite to hematite, (Hanesch, 2009). Several scans were conducted across the sample stage to assess microscopic heterogeneity. The spectra that were obtained were compared to previously characterized spectra for common Fe-(oxyhydr)oxide minerals available in Das et al. (Das et al., 2016; Das and Hendry, 2011) and in the RUFFTM database (Lafuente et al., 2015). The TEM images were collected on a Hitachi HT7700 instrument and selected area electron diffraction images were collected on TEM spots to determine the crystallinity of the particles of interest. Atomic-scale chemical coordination of Mo in ochreous precipitate samples was determined by Mo K-edge X-ray absorption spectroscopy measurements performed using the Hard X-Ray Microanalysis beamline (HXMA, 06ID-1) at the Canadian Light Source (University of Saskatchewan, Saskatoon, Canada) (Appendix H). At least three scans per sample were collected to improve the signal-to-noise ratio. Scans were collected from -200 eV up to k = 14  Å-1 relative to the theoretical Mo K-edge (20,000 eV). Steps of 0.25 eV were made in the X-ray near-edge spectra (XANES) and steps of 0.05 Å-1 were made in the extended X-ray absorption fine structure (EXAFS) region. ATHENA and ARTEMIS (Demeter v.0.9.25, Ravel and Newville, 2005) were used for data reduction and analyses. Linear combination fitting of the     197  XANES was performed for each sample and included the following reference materials: MoVIO3, MoIVO2, MoVI -adsorbed onto ferrihydrite, MoVI -adsorbed onto goethite, and MoVI -adsorbed onto hematite (Das et al., 2016). The program FEFF 6.0 was used to extract the phase and amplitude functions from the structure of Fe2(MoO4)3 for the Mo—O and Mo—Fe scattering paths. Non-linear least-squares fitting was performed on the Fourier transform of background-subtracted k3-weighted χ(k) spectra from k = 2.0 to 11.0 Å-1.  6.4 Results 6.4.1 Aqueous geochemistry The chemistry of basal seepage of the BCTMF at the Rock Toe sampled during this study was generally anoxic and had circumneutral pH (Table 6.1)—consistent with historical data for the facility. Anoxic conditions were indicated by elevated Fe(II) and Mn(II) concentrations, which ranged from 149 to 273 µM and 98 to 132 µM, respectively (Table 6.1 and Table 6.2). Nitrate—another redox indicator—was present in mine process water but was not detected in the Rock Toe. Trace amounts of sulfide (0.14 µM) were measured in the October 2015 sampling round, but not in other sampling events (Table 6.1). Molybdenum concentrations were relatively constant, varying between 7.7 and 8.4 µM (Table 6.3). These concentrations are approximately 2 orders of magnitude greater than those of typical natural waters (Smedley and Kinniburgh, 2017). Mineral saturation indices (SI) calculated in PHREEQC showed that the Rock Toe remained consistently close to powellite saturation (Table 6.3). Molybdenum isotopic compositions at the Rock Toe between the 2014 and 2016 sampling events remained at 1.0 ± 0.1     198  ‰ (2 SD). This δ98Mo is approximately 0.3 to 0.4 ‰ heavier than tailings pond water in the BCTMF, indicating some isotopic fractionation along the flowpath leading to the Rock Toe. The oxidation of Rock Toe seepage during flow to the SRD pond was reflected by several changes in aqueous geochemistry. Iron, Mo, and sulfate concentrations decreased and δ98Mo increased relative to the Rock Toe (Figure 6.3). The pH range of 6.9 to 7.1 in the SRD was slightly more elevated than in the Rock Toe (pH 6.6 to 6.7). Water in the SRD was generally undersaturated with respect to molybdate minerals (Table 6.3). Aqueous geochemistry was more variable between sampling events in the SRD in comparison to the Rock Toe: during peak discharge conditions (April 2016) SRD pond chemistry more closely reflected Rock Toe drainage in comparison to baseflow sampling events (October 2014 and October 2015), where aqueous concentrations of Fe, Mo, and sulfate were lower and larger Mo isotopic fractionation was observed (Figure 6.3).      199  Table 6.1. Water-sample chemistry: Field parameters and anions concentrations            <dl indicates that the parameter was below detection limit.  -- indicates that the parameter was not analyzed.       Field parameters Anions Sample     Alkalinity Specific conductivity Dissolved oxygen ORP Cl- F- NO3- NO2- SO4-2 S2- ID Date pH mM mS/cm mg/L mV mM mM mM mM mM µM TP Oct-14 7.53 0.92 3.86 7.60 77 9.9 0.028 0.086 0.0057 19 <dl Rock Toe Oct-14 6.71 3.1 2.64 0.45 -144 5.5 0.023 <dl <dl 15 <dl Rock Toe Oct-15 6.63 3.1 2.77 0.94 -- 5.6 0.023 <dl <dl 14 0.14 Rock Toe Apr-16 6.57 2.9 -- -- -- 5.4 0.022 <dl <dl 14 <dl SRD Oct-14 7.09 3.1 1.63 1.49 40 3.0 <dl 0.028 <dl 8.5 <dl SRD Oct-15 7.10 2.9 2.30 6.75 -- 3.0 0.011 0.021 <dl 8.3 <dl SRD Apr-16 6.88 2.6 -- -- -- 4.8 <dl <dl <dl 12 <dl SRD-W1 Oct-15 6.18 -- 2.75 6.54 43 -- -- -- -- -- -- B1-12d Dec-16 8.06 2.5 -- -- -- -- -- -- -- -- --     200  Table 6.2. Water-sample chemistry: Metal concentrations and Mo isotope ratios     Metals       Sample   Ca K Mg Na Si Al Fe Mn Mo Pb Sr δ98Mo ± 2 SD* n† ID Date mM mM mM mM µM µM µM µM µM µM µM ‰     TP Oct-14 16 1.4 1.5 14 <dl <dl 1.0 128 34 0.0044 132 0.58 0.02 3 Rock Toe Oct-14 13 0.85 1.4 9.0 246 <dl 273 135 8.4 0.00095 95 1.03 0.02 3 Rock Toe Oct-15 12 0.90 1.4 9.1 66 9.4 192 98 8.4 0.00051 100 0.91 0.01 1 Rock Toe Apr-16 10 0.77 1.2 7.9 <dl 0.52 149 100 7.7 0.000040 100 0.94 0.12 4 SRD Oct-14 8.4 0.20 1.7 3.4 316 <dl 0.5 2.4 0.47 0.00051 31 2.63 0.07 3 SRD Oct-15 7.6 0.31 1.3 3.9 <dl 7.1 7.2 19 1.9 0.00017 41 1.26 0.01 2 SRD Apr-16 9.0 0.64 1.1 6.7 <dl 0.26 118 90 6.3 0.000064 83 1.00 0.03 2 SRD-W1 Oct-15 13 0.89 1.5 9.1 61 10 182 105 8.2 0.0011 99 0.96 0.01 1 B1-12d Dec-16 -- -- -- -- -- -- -- -- -- -- -- 0.93 0.02 1 <dl indicates that the parameter was below detection limit.  -- indicates that the parameter was not analyzed. * Represents the two-standard-deviation of replicate δ98Mo analyses on the MC-ICP-MS. For the samples that were analyzed only once, 2 SE is shown and denoted by italic font. † Number of MC-ICP-MS replicates.       201  Table 6.3. Water-sample chemistry: Mineral saturation indices (SI) Sample   Molybdates Fe-(oxyhydr)oxides* ID Date Powellite Wulfenite FeMoO4 Fe(OH)3(a)† Goethite Hematite Jarosite‡  Schwertmannite§                 min max min max TP Oct-14 0.58 1.3 -1.2 2.6 7.9 18 -4.0 0.04 5.9 11 Rock Toe Oct-14 -0.03 0.34 0.74 5.2 10 23 2.0 6.1 25 30 Rock Toe Oct-15 -0.04 0.12 0.60 5.1 10 22 1.6 5.8 24 29 Rock Toe Apr-16 -0.10 -0.94 0.49 4.9 10 22 1.1 5.3 23 28 SRD Oct-14 -1.3 -1.2 -3.1 2.4 7.7 17 -7.4 -3.3 2.4 7.4 SRD Oct-15 -0.71 -1.0 -1.3 3.7 8.9 20 -3.5 0.92 13 18 SRD Apr-16 -0.21 -1.0 0.35 4.6 10.0 22 0.43 4.4 21 26  Table 6.3. (continued) Water-sample chemistry: Mineral saturation indices (SI) Sample   Other minerals ID Date Pyrolusite* Siderite Melanterite Calcite Gypsum               TP Oct-14 11 -1.9 -6.3 -0.05 0.05 Rock Toe Oct-14 12 0.29 -3.9 -0.39 -0.07 Rock Toe Oct-15 12 0.05 -4.0 -0.50 -0.09 Rock Toe Apr-16 12 -0.12 -4.2 -0.63 -0.15 SRD Oct-14 10 -2.0 -6.8 -0.10 -0.35 SRD Oct-15 11 -0.87 -5.6 -0.18 -0.38 SRD Apr-16 12 0.11 -4.3 -0.37 -0.23 * Mineral SI shown for Fe-(oxyhydr)oxides and pyrolusite were simulated in PHREEQC by allowing the water to come to full equilibrium with atmospheric O2 and CO2. This SI therefore represents minerals that can thermodynamically precipitate from full oxidation of TP, Rock Toe, and SRD water. † Fe(OH)3 = amorphous iron-(oxyhydroxide). ‡ Mineral SI for jarosite are the minimum and maximum values among the jarosite minerals present in PHREEQC's wateq4f database: Jarosite-Na, Jarosite-K, and Jarosite-(ss). § Mineral SI for schwertmannite are the minimum and maximum values for the window of schwertmannite solubility constants recommended by Bigham et al. (1996). See section 6.3.3.1 for full details.     202   Figure 6.3. SO42-, Fe, Mo, and δ98Mo in basal tailings drainage at the Rock Toe in comparison to the SRD pond. Blue colors indicate baseflow hydrological conditions; red colors indicate freshet. During baseflow, oxidation of Rock Toe water leads to precipitation of Fe-(oxyhydr)oxide minerals and loss of Fe(II) as the water flows into the SRD. Molybdenum is concurrently removed, leading to increasing δ98Mo in SRD waters as compared to the Rock Toe. During freshet, higher flushing rates along the flowpath cause SRD waters to be more geochemically similar to Rock Toe discharge. Light shading = Rock Toe; Darker shading = SRD. Error bars for δ98Mo are ± 0.07 ‰ (2 SD external reproducibility of Mo(UBC) isotope standard).     203  The Fe minerals that might precipitate in the SRD as a result of the oxidation of Rock Toe drainage were investigated with a PHREEQC simulation wherein Rock Toe waters were allowed to equilibrate with atmospheric PO2 and PCO2. This simulation showed that oxidation drives high degrees of supersaturation for several Fe-(oxyhydr)oxide minerals including ferrihydrite, hematite, goethite, jarosite, and schwertmannite (Table 6.3). Oxidized Rock Toe drainage plots in the stability field of ferrihydrite in Eh-pH space (Bigham et al. 1996) and ferrihydrite precipitation is favored over more crystalline phases such as goethite and hematite during the rapid oxidation of Fe(II)-bearing waters at circumneutral pH (Blowes et al. 2014; Cornell and Schwertmann 2003). In addition, jarosite and schwertmannite precipitation is typically restricted to lower pH conditions (pH < 4) than those found in the SRD pond (Bigham and Nordstrom, 2001; Blowes et al., 2014; Cornell and Schwertmann, 2003). 6.4.2 Molybdenum sorption experiment The oxidation of Rock Toe water was also studied through the Mo sorption experiment. The 20-L pail of Rock Toe water, initially clear, turned translucent and brown after approximately 32 hours; after 5 days it was orange-brown and visibly more turbid (Figure I.3). ICP-MS analyses of this water revealed a 99 % decrease in aqueous Fe concentrations within the first 7 days (Figure 6.4). However, at the experiment's end after 44 days, Mo concentrations had only decreased by 4 %. The lack of significant Mo removal during this experiment was accompanied by a lack of significant change in δ98Mo in the remaining aqueous pool relative to initial conditions: δ98Mo remained at 0.9 ‰ (Table 6.2). Substantially greater Mo attenuation was present in the SRD pond than what was observed in this experiment—Mo concentrations     204  were up to 17 times lower and δ98Mo was higher (1.3 to 2.6 ‰) in the SRD pond in comparison to Rock Toe drainage. The main differences between the sorption experiment and conditions in the pond that could have contributed to less Mo attenuation in the experiment were: (1) a rise in pH to pH 8.1 in the sorption experiment; (2) batch vs. active-flow conditions; and (3) a visually greater Fe-(oxyhydr)oxide-to-water ratio in the pond in comparison to the sorption experiment (Figures I.1 to I.3)  Figure 6.4. Concentrations of dissolved Fe, Mo, and Mn during the oxidation of a 20-L pail of Rock Toe water. Fe(II) oxidation lead to precipitation of Fe-(oxyhydr)oxide minerals. Negligible amounts of Mo and Mn were removed during this time. See section 6.3.2 for experimental details. 6.4.3 Ochreous precipitate sample characterization 6.4.3.1 Geochemistry of ochreous precipitates Most ochreous precipitate samples collected in the SRD and at the end of the Mo sorption experiment contained between 36 and 42 wt. % Fe and < 1 wt. % of organic carbon and other metals. Exceptionally, sample SRD-S1 had a comparatively lower Fe content of 11 wt. % and a     205  greater abundance of other constituents including 4.2 wt. % Al; 2.5 wt. % organic carbon; 1.5 wt. % K; and 1.2 wt. % Mn (Table 6.4). All ochreous precipitate samples had elevated Mo contents, which ranged between approximately 760 and 4,500 ppm (Table 6.4). This content translates to a range in Fe-to-Mo molar ratios of 156 to 745 in the precipitates, in comparison to ratios spanning from 19 to 32 in Rock Toe seepage. Precipitate samples also had significantly lighter δ98Mo values of -0.3 to 0.2 ‰ in comparison to SRD and Rock Toe waters, whose δ98Mo values ranged from 0.9 to 2.6 ‰ (Figure 6.5).       206   Figure 6.5. Comparison of δ98Mo values in the SRD, Rock Toe, and Fe-(oxyhydr)oxides, shown along with TCM mine tailings, process water, waste rock, and ore δ98Mo from Skierszkan et al. (2017). Process water is unfractionated (0.7 ‰) relative to the average δ98Mo of solid mine waste (tailings, waste rock and ore). The ~ 0.4 ‰ increase in δ98Mo between the tailings pond and the Rock Toe is attributed to Mo attenuation. Molybdenum sorption to Fe-(oxyhydr)oxides in the SRD drives the SRD’s δ98Mo towards even heavier values.      207  Table 6.4. Elemental abundances and Mo isotope ratios in ochreous precipitate samples Sample N-total C-total TIC TOC Na Mn Fe Mg Al K Ca Mo δ98Mo ± 2 SD* n† Fe/Mo ID % % % % wt. %. wt. %. wt. %. wt. %. wt. %. wt. %. wt. %. ppm ‰     molar ratio SRD-S1 0.33 2.9 0.40 2.5 0.98 1.2 11 0.41 4.2 1.5 0.36 756 -0.28 0.02 1 146 SRD-S2 0.13 1.8 0.77 1.0 0.25 0.77 38 0.030 0.31 <dl 0.43 4464 -0.02 0.02 1 247 SRD-S4 0.058 1.3 0.71 0.56 0.10 0.54 42 0.020 0.10 <dl 0.38 4386 0.24 0.02 1 166 SRD-W1 PPTS‡  -- -- -- -- -- -- -- -- -- -- -- -- 0.05 0.05 2 247 Mn-Buck‡  -- -- -- -- -- -- -- -- -- -- -- -- -0.27 0.02 1 727 B-1A -- -- -- -- 0.13 1.0 36 0.040 0.070 0.060 0.60 839 -0.28 0.01 1 745 <dl indicates that the parameter was below detection limit, -- indicates that the parameter was not analyzed. * Represents the two-standard-deviation of replicate δ98Mo analyses on the MC-ICP-MS. For the samples that were analyzed only once, 2 SE is shown and denoted by italic font. † Number of MC-ICP-MS replicates. ‡ For the samples SRD-W1 PPTS and Mn-Buck, elemental abundances could not be determined because the sample mass prior to digestion was not known. However, elemental ratios could be calculated from ICP-MS analyses.     208  6.4.3.2 Mineralogical characterization of ochreous precipitates Mineralogical analyses showed that poorly crystalline Fe-rich material dominated the ochreous precipitate samples. Despite this high Fe content, XRD scans did not yield distinctive Fe-(oxyhydr)oxide mineral peaks (Figure I.5). However, Raman spectra corresponding to poorly crystalline ferrihydrite were present in samples SRD-S2, SRD-S4, and B-1A (Figure 6.6a) as shown by the broad Raman bands around ~361, ~508, and ~707 cm-1 (Das and Hendry, 2011). The broadening of Raman bands in these samples relative to pure synthetic ferrihydrite is typical for environmental samples of low crystallinity (Hanesch, 2009). Transmission-electron microscopy further corroborated that these samples were predominantly made of nm-sized Fe-rich aggregates lacking crystalline structure as indicated by their morphology and broad and diffuse electron diffractograms (Figure 6.7) (Petrunic et al., 2006). These results generally pointed to a high abundance of poorly crystalline ferrihydrite in the ochreous precipitates. In addition to ferrihydrite, minor amounts of other phases were present in these samples. Sample SRD-S2 contained quartz (Figure I.5 and Error! Reference source not found.) and other phases, as shown by  additional minor Raman peaks that could not be definitively ascribed because they were either very small or masked by the stronger quartz peaks (Figure I.6). While TEM indicated that SRD-S2 predominantly had armorphous material, structures resembling re-crystallization of ferrihydrite to more stable phases were also observed (c.f. Figure 14.17 in Cornell and Schwertmann, 2003); their bladed morphology and water-rich formation environment suggested that these could be nano-scale goethite (Figure I.4.c).      209  Sample SRD-S1 stands out as having a more diverse mineral assemblage at the expense of Fe-(oxyhydr)oxide. X-ray diffractograms showed the presence of several aluminosilicates, including quartz, feldspars and clays (Figure I.5). A Raman spectrum indicative of hematite and/or maghemite was also obtained for SRD-S1 (Figure 6.6.b), although hematite and maghemite are not easily differentiated by Raman spectroscopy (Hanesch, 2009). A single peak matching hematite was also observed in the XRD pattern of that sample, although the presence of this mineral could not definitively be identified using the diffractogram. Transmission-electron microscopy imaging of SRD-S1 also revealed minor amounts of algal detritus and ~ 500-nm sized particles with better-defined edges and electron diffraction (i.e. greater crystallinity) (Figure I.5.a and Figure I.5.b). However, high fluorescence for most Raman spectra collected for SRD-S1 and TEM imaging indicated that a major component of that sample was amorphous material (Figure 6.6a and Figure 6.7).        210   Figure 6.6. Select Raman spectra for ochreous precipitate samples. Panel A shows sample spectra compared to: aferrihydrite (Fh) and Mo adsorbed onto ferrihydrite from Das and Hendry (2011); and bDas et al. (2016). All samples have higher fluorescence as a result of poor crystallinity. Samples SRD-S2, B-1A, and SRD-S4 have broad bands around 361, 508, and 707 cm-1 which correspond to poorly crystalline ferrihydrite. Panel B shows a spectrum from sample SRD-S1 that is consistent with hematite and/or maghemite (Das et al., 2011).       211   Figure 6.7. Selected transmission-electron microscopy images of Fe-(oxyhydr)oxide ochreous precipitates collected in the SRD.  The approximate areas analyzed for electron diffraction and EDX are shown in yellow circles. Electron diffractograms are shown at bottom right; broad and diffuse rings indicate amorphous or nanocrystalline structure. Additional TEM images are shown in Error! Reference source not found.. 6.4.3.2.1 X-ray absorption spectroscopy X-ray absorption near-edge spectra were used to determine the oxidation state and molecular coordination environment of Mo in ochreous precipitate samples. There was no evidence of reduced Mo (e.g. MoV or MoIV) in the samples, which would have caused a shift to lower absorption edge energies in the XANES (Figure 6.8). MolybdenumVI in tetrahedral (Td) coordination was the dominant form of Mo in samples SRD-S4, SRD-S2, and B-1A as indicated     212  by the pre-edge peak at approximately 20,000 eV (Figure 6.8). This pre-edge peak was suppressed in samples relative to the Na2MoO4 reference compound, which could be a result of distortion of the MoO4 tetrahedra due to interactions with coordinating sorption surfaces (Gustafsson and Tiberg, 2015). In sample SRD-S1, the maximum absorption energy was shifted to a slightly lower value of 20,037 eV compared to other samples, where it was around 20,040 eV. This lower maximum absorption energy corresponded closely to that of the octahedral (Oh) MoO3 reference compound—which has a six-coordinate crystal structure—suggesting the possibility that some Mo in SRD-S1 had Oh coordination. The chemical coordination of sorbed Mo was further constrained by linear combination fitting (LCF) of the XANES (Table 6.5). In most cases, Td MoVI adsorbed to ferrihydrite was the dominant Mo coordination environment (≥ 99 %). Alternative fits that were conducted for SRD-S2 and B-1A suggested that either MoVI adsorbed to hematite or goethite (< 24 %) were possible; however, the presence of these minerals was not definitively confirmed by Raman spectroscopy, XRD, or TEM. Sample SRD-S1 had a characteristically different Mo coordination environment: a significant fraction (57 %) of MoVI was fitted to octahedral (Oh) coordination by using the MoO3 reference compound as an analogue. The balance of the Mo in SRD-S1 was fitted as tetrahedrally coordinated Mo adsorbed to ferrihydrite, hematite, and amorphous Al(OH)3. Al(OH)3 was not identified in this sample, but SRD-S1 had a relatively higher wt. % Al in the form of clays and other aluminosilicate minerals. The tendency of clay minerals to adsorb Mo (Goldberg et al., 1996; Goldberg and Forster, 1998; Sun and Selim, 2018) provided a rationale for the inclusion of a Mo–Al phase in LCF. Molybdenum also forms Oh complexes with organic     213  matter and during adsorption onto Mn-(oxyhydr)oxides (Gustafsson and Tiberg, 2015; Wasylenki et al., 2011); however, the Mn content in SRD-S1 was comparable to other ochreous precipitate samples where Oh coordination of Mo was not observed, and organic ligands were tested in the EXAFS fitting but yielded poor results. Fitting of the EXAFS spectra indicated that Mo was attenuated via inner-sphere complexation to sorption surfaces in the SRD pond (Table 6.6). Molybdenum coordination models with good fits to the EXAFS data were obtained for all samples except for SRD-S1 (Figure 6.9). In samples SRD-S4, SRD-S2, and B-1A a first-shell O atom at an interatomic distance of 1.76 to 1.77 Å and a Mo coordination number ≈ 4 were fitted to the first large peak in the Fourier-transformed k3·χ(k) spectra (Table 6.6), consistent with Td MoVI (Arai, 2010) and LCF results. The second-shell neighbor could be fitted as Fe in two types of coordination environments: (1) edge-sharing bidentate mononuclear bonding of MoVI tetrahedra (Mo—Fe1) and (ii) corner-sharing bidentate-binuclear bonding of MoVI tetrahedra (Mo—Fe2). Molybdenum in SRD-S4 was only coordinated in an edge-sharing configuration, while Mo in SRD-S2 and B-1A was coordinated in both edge- and corner-sharing configurations. Sample SRD-S1 had a more complex Mo coordination environment with multiple bonding configurations and neighbors. The magnitude of the first shell peak in the Fourier-transformed k3·χ(k) spectrum was suppressed and the longer-range backscatterers were located at shorter distances (Figure 6.9). The general Mo coordination for that sample was explained by a five-shell but heavily constrained model involving Mo—O, Mo—Al, and Mo—Fe interactions (Table 6.6). Octahedrally coordinated MoVI was represented by two O shells fitted at 1.76 Å and     214  2.17 Å. However, coordination numbers of these O neighbors are lower than expected for pure Oh Mo, which could be explained by the additional presence of Td Mo. An Al shell was fit at 2.50 Å—a similar interatomic distance to that of the edge-sharing Mo—Fe shells that were fitted in other samples, which suggests a structural similarity between those Mo—Fe and the Mo—Al shells. Molybdenum-Al interaction is also consistent with a greater presence of aluminosilicate phases in SRD-S1 as identified by XRD. A nearby edge-sharing Fe shell and a further corner-sharing Fe shell were also fit to SRD-S1. This coordination model is complex, and the accuracy of its parameters is physically limited by the amount of information available in the EXAFS signal, but it suggests that the Mo in SRD-S1 has both Oh and Td coordination with O and is adsorbed to Fe and Al phases in multiple configurations. In addition, it is worth noting that in all ochreous precipitate samples, outer-sphere surface complexes are possible but cannot be confirmed because their EXAFS signal would be masked by the stronger amplitude of inner-sphere complexes (Gustafsson and Tiberg, 2015).        215  Table 6.5. Linear combination fitting results for Mo K-edge XANES spectra of ochreous precipitate samples. Sample Mo(VI)-Ferrihydrite Mo(VI)-Hematite MoO3 Mo(VI)- Al(OH)3 Sum R-factor SRD-S4 100 ± 0 %    100% 0.0019 SRD-S2 99 ± 6 %    99% 0.0009 B-1A PPTS 100 ± 0 %    100% 0.0007 SRD-S1 19 ± 9 % 19 ± 5 % 51 ± 5 % 12 ± 5 % 101% 0.0009 Table 6.6. Summary of Mo K-edge EXAFS fitting parameters for ochreous precipitate samples.  Eo (eV) CN R (Å) σ2 (Å2) R-Factor χν2 SRD-S4        Mo–O -4.1(9) 3.7(2) 1.764(4) 0.0036(5) 0.006 15.2  Mo–Fe1  0.8(2) 2.79(2) 0.01a   SRD-S2        Mo–O -5.0(8) 3.8(2) 1.763(3) 0.0041(4) 0.003 27.6  Mo–Fe1  1.0(1) 2.79(1) 0.01a    Mo–Fe2  0.9(3) 3.56(2) 0.01a   B-1A PPTS        Mo–O -3.6(9) 3.8(2) 1.770(4) 0.0047(5) 0.005 7.0  Mo–Fe1  1.2(2) 2.80(1) 0.01a    Mo–Fe2  0.7(3) 3.54(3) 0.01a   SRD-S1        Mo–Oeq -6(3) 3.3(2) 1.76(1) 0.006a 0.016 56  Mo–Oax  1.3(3) 2.17(2) 0.006a    Mo–Al  1a 2.50(4) 0.006a    Mo–Fe1  0.5a 2.65(3) 0.006a    Mo–Fe2  1a 3.29(2) 0.008a   Fitting was performed over a range from kmin = 2.0 Å-1 to kmax = 11.0 Å-1. The amplitude reduction factor, S02, was fixed at 0.9. Eo = energy shift. R = interatomic distance. σ2 = Debye-Waller factor. χν2 = reduced chi square. The numbers in parentheses are the Artemis-provided errors calculated from the diagonal of the covariance matrix and scaled by the square-root of χν2.  a Parameter held constant in the fitting procedure     216   Figure 6.8. Normalized Mo K-edge X-ray absorption near-edge spectra of reference compounds and ochreous precipitate samples. Solids lines denote measured spectra and blue circles denote linear combination fitting results.      217   Figure 6.9. Molybdenum K-edge k3-weighted χ(k) background-subtracted spectra (left) and the corresponding Fourier transforms (right).  Solids lines denote experimental data and blue circles denote fitting results.      218  6.5 Discussion In the ensuing discussion, results are combined to provide a conceptual model for Mo attenuation in the BCTMF. Attenuation within the tailings impoundment and embankment is interpreted based on aqueous geochemistry results and is compared to the attenuation behavior of Mo in waste-rock environments at the same site. Later, X-ray absorption spectroscopy and geochemistry in the SRD pond samples are used to constrain the mechanisms of Mo attenuation during sorption onto Fe-(oxyhydr)oxide following oxidation of Fe(II) bearing mine tailings drainage. 6.5.1 Processes of Mo attenuation in the BCTMF The shift towards lower Mo concentrations and higher δ98Mo between mine process water and the Rock Toe is indicative of Mo chemical attenuation in the BCTMF (Figure 6.10). At a δ98Mo of 1.0 ± 0.1 ‰, Rock Toe seepage is also isotopically heavier than mine tailings and waste rock, which have δ98Mo of 0.2 ‰ and 0.7 ‰, respectively (Skierszkan et al., 2017). This isotopic offset rules out tailings and waste rock weathering as a source of isotopically heavy Mo to Rock Toe seepage. Groundwater discharge into the BCMTF is also unlikely to explain this isotopic fractionation: groundwater—which contributes only approximately 5 % of flow at the Rock Toe (Lorax Environmental Services Ltd., 2011b)—would need a δ98Mo of 7 ‰ to explain the Mo isotopic composition of the Rock Toe based on isotopic mass balance. A δ98Mo value that high falls well outside of the range reported for natural waters (Smedley and Kinniburgh, 2017) as well as for background surface water and groundwater at the TCM— the latter two     219  having a median δ98Mo of 1.2 ‰ (c.f. Table 3 in Skierszkan et al., 2017). Therefore, we infer that the Mo isotopic fractionation between process water and the Rock Toe results from attenuation processes within the BCTMF; however, the extent of attenuation is presumed to be modest given that isotopic fractionation between process water and the Rock Toe is relatively minor at 0.3 to 0.4 ‰. Knowledge of Mo isotopic fractionation factors for various attenuation processes makes it possible to constrain their role by isotopic modelling of the geochemical evolution along the flowpath from the tailings pond to the Rock Toe. Consistent mineral saturation with respect to powellite in drainage at the Rock Toe suggests that precipitation-dissolution of this mineral may be an important control on Mo transport within the BCTMF. Under condition close to isotopic equilibrium, an isotopic fractionation factor between powellite and aqueous Mo of αpowellite-aqueous = 0.99962 ± 0.0006 was recently estimated (Chapter 5). Using this fractionation factor, the Mo isotopic composition of tailings drainage can be plotted against Mo/Cl, which progressively decreases as a function of Mo removal. This ratio is used because (1) Cl is a conservatively behaved tracer; and (2) Mo and Cl concentrations are approximately three orders of magnitude greater in TCM mine process water when compared to unimpacted groundwater in the Bruno Creek drainage [i.e. groundwater well BC3A, Skierszkan et al. (2017)], which makes Mo/Cl insensitive to dilution from groundwater discharge. However, this model showed that precipitation of powellite alone (at isotopic equilibrium) cannot account for the geochemical composition of the Rock Toe (Figure 6.11): the 0.4 ‰ increase of δ98Mo between process water and the Rock Toe would would imply that 99 % of the Mo was precipitated as powellite, which     220  is inconsistent with the ~ 50 % decrease in Mo concentrations that has historically been observed between process water and Rock Toe drainage (Lorax Environmental Services Ltd., 2011b), and the ~ 75 % decrease that was observed in samples collected during this study. An additional process is therefore required to explain the attenuation of Mo within the BCTMF.      221   Figure 6.10. Molybdenum isotopic compositions against Mo concentrations in the flowpath from the tailings pond to the Rock Toe and SRD pond. Removal of Mo is associated to a rise in δ98Mo and decrease in Mo concentrations, as shown by the arrow. Also shown are data from drainage of oxidized mine waste-rock storage facilities (WRSFs) at the TCM, which have lower Mo concentrations and higher δ98Mo as a result of greater Mo removal in comparison to the Rock Toe. Tailings Pond and WRSF data are from Skierszkan et al. (2017). Error bars for δ98Mo are ± 0.07 ‰ (2 SD external reproducibility of Mo(UBC) isotope standard, Appendix A).  Figure 6.11. Simulations of the effect of Mo removal from mine process water on δ98Mo as a function of Mo/Cl ratios. Mo/Cl ratios decrease because Cl- is a conservative aqueous solute while Mo is attenuated via precipitation of powellite (CaMoO4) and sorption onto ferrihydrite (Fh). Mine process water (TP) is used as an input. The blue shading shows the effect of Mo removal via powellite precipitation at isotopic equilibrium using the equilibrium fractionation factor of Skierszkan et al. (2017). Red shading shows the effect of a combination of powellite precipitation (at isotopic equilibrium) accounting for 60 % of Mo removal, and sorption onto ferrihydrite accounting for 40 % of Mo removal. The fractionation factor for sorption onto ferrihydrite used in this model was Δ98Mo = 1.00 ± 0.40 ‰ (see section 6.5.3.1).The thick line indicates the values predicted by the model and the shading shows the uncertainty envelope resulting from uncertainties in the isotopic fractionation factors.     222  Kinetic isotope effects (KIE) could account for some of the additional isotopic fractionation that is required to drive δ98Mo to the values observed at the Rock Toe. Both equilibrium and kinetic isotope fractionation can occur as a result of powellite precipitation, but the latter is larger in magnitude (Chapter 5). However, long water residence times in the BCTMF suggest that kinetic isotope effects are probably of minor importance in comparison to equilibrium isotope fractionation. Isotopic equilibrium takes weeks to months to be established during powellite precipitation (Chapter 5). In contrast, water residence times in the BCTMF are most likely several years to decades: groundwater velocities in tailings impoundments are typically less than 15 m a-1 (Aubertin et al., 1996; Blowes et al., 2014) and flowpaths from the tailings impoundment to the Rock Toe are hundreds of meters in length. As a result, the probability that KIEs drive the δ98Mo towards higher values than expected from pure isotopic equilibrium conditions is small—over the decades since the BCTMF's construction, ongoing dissolution-precipitation of powellite makes conditions of isotopic equilibrium more likely.  A hypothesis accounting for Mo adsorption onto Fe-(oxyhydr)oxides in the BCTMF embankment sands provides an alternative explanation for the observed isotopic fractionation between mine process water and the Rock Toe. Approximately 35 % of the water reporting to the Rock Toe comes from flow through the embankment sands, where water-unsaturated conditions allow the presence of Fe-(oxyhydr)oxide minerals formed via oxidative dissolution of pyrite (Lorax Environmental Services Ltd., 2011b). There are localized acidic zones near the base of the BCTMF embankment, where shake-flask extractions of tailings showed that Mo was insoluble and presumably adsorbed onto Fe-(oxyhydr)oxide minerals (Lorax Environmental     223  Services Ltd., 2011b). Taking into account the possibility that Mo removal through adsorption onto Fe-(oxyhydr)oxides is occurring along this water-unsaturated flowpath, a good fit to the data at the Rock Toe is obtained through a model that assigns 60 % of Mo removal to powellite precipitation and 40 % of Mo removal to sorption onto ferrihydrite (Figure 6.11). This model also more realistically incorporates geochemical processes occurring within the BCTMF, wherein a combination of powellite precipitation in the anoxic and water-saturated flowpath (through the slimes) and Mo sorption in the water-unsaturated flowpath (through the sands) could control Mo transport.  Another possible attenuation mechanism within the BCTMF is the precipitation of (oxo)thiomolybdate phases in slimes below the tailings pond. These slimes are locally susceptible to sulfide generation (Lorax Environmental Services Ltd., 2011b) and the abundance of organic solvents and sulfate in process water provides a driving force for localized microbial sulfate reduction below the sediment-water interface. Sulfide was detected in porewater below the tailings pond sediment-water interface at one of two sites studied in a previous investigation (Lorax Environmental Services Ltd., 2011b) and it was detected in the Rock Toe at low levels in one of three sampling events in this study. The lack of appreciable sulfide at the Rock Toe suggests that sulfide is removed along the flowpath; plausible removal processes include H2S degassing in the drain at the base of the BCTMF and precipitation of sulfide minerals involving Fe(II) and possibly (oxo)thiomolybdate. Given (1) the molar excess of Fe(II) relative to sulfide in BCTMF drainage, (2) the low solubility and rapid formation kinetics of FeS (Rickard, 1995), and (3) the readiness with which MoO42- converts to MoS42- and binds to FeS in sulfidic     224  conditions, it is possible that some Mo is removed via co-precipitation with FeS in the anoxic flowpath (Helz et al., 1996). However, no evidence for (oxo)thiomolybdate was present in XAS analyses of ferrihydrite samples collected in the SRD pond;  (oxo)thiomolybdate oxidation to MoO42- takes many weeks (Erickson and Helz, 2000) and its sorption behavior—although unknown for ferrihydrite—is similar to MoO42- during sorption onto pyrite (Bostick et al., 2003) suggesting that it is unlikely to be present in Rock Toe seepage. The stoichiometric excess of Mo over sulfide in the BCTMF suggests that formation of (oxo)thiomolybdate—if present—is limited by sulfide availability. In addition, quantifying the effect of Mo removal via this pathway on the δ98Mo at the Rock Toe is difficult given the lack of a well-calibrated fractionation factor but ab initio calculations suggest that isotopic fractionations in excess of 6 ‰ exist between MoO42- and MoS42- (Nägler et al., 2011; Tossell, 2005), and data from sulfidic marine water columns also confirm that light Mo isotopes are preferentially removed from solution in Fe(II) and S2- rich settings (Nägler et al., 2011). Therefore, Fe(II)-(oxo)thiomolybdate interactions cannot be ruled out as an explanation for some degree of Mo attenuation within the BCTMF, because the isotopic fractionation would presumably be in the same direction as observed for powellite precipitation and Mo sorption onto Fe-(oxyhydr)oxides. Further mineralogical analyses of tailings samples within anoxic zones of the BCTMF embankment would be required to fully rule out this attenuation process of Mo.  From the analysis above, we can conclude that powellite precipitation at isotopic equilibrium is unlikely to constitute the sole source of Mo attenuation in the BCTMF. It is probable that Mo sorption onto ferrihydrite is occurring in unsaturated flowpaths in the     225  embankment. Precipitation of powellite under conditions of kinetic isotope fractionation is also possible, and (oxo)thiomolybdate precipitation is plausible but (if present) probably of minor importance and would require further mineralogical analyses to be confirmed. 6.5.2 Comparing molybdenum mobility in anoxic mine tailings to waste-rock environments Regardless of the combination of processes that are attenuating Mo in the BCTMF, a comparison of the δ98Mo between the BCTMF and the two WRSFs at the mine shows that Mo is more mobile in an anoxic tailings environment than in oxic waste rock. Differences in Mo concentrations alone between the BCTMF and the WRSFs at the TCM are not sufficient to assess the extent of Mo mobility: Mo concentrations are 1 to 2 orders of magnitude higher in the BCTMF (Figure 6.10), which could simply be because its predominant water source is Mo-rich process water, while the water source to the two WRSFs is a combination of groundwater discharge and meteoric precipitation affected by MoS2 oxidation. However, the BCTMF has a δ98Mo of approximately 1.0 ‰, in comparison to the δ98Mo of 1.6 to 1.9 ‰ in the WRSFs, which are driven to higher values as a result of sorption processes (Skierszkan et al., 2017) (Figure 6.10). Simulations of Mo removal via combined powellite precipitation and ferrihydrite precipitation support that Mo is more mobile in the BCTMF: only ~ 60 % of total Mo removal was required to match the Rock Toe’s δ98Mo (section 6.5.1 and Figure 6.11) which, as a first-order estimate, is considerably smaller than the > 90 % Mo removal required to drive the isotopic fractionation observed in drainage in the two WRSFs (Skierszkan et al., 2017).     226  This contrast in Mo mobility between tailings and waste rock is explained by hydrological and redox differences between these materials. Mine tailings facilities typically have negligible air circulation and greater water saturation compared to WRSFs, where coarser grain sizes promote convective gas transport and drain-down—leading to greater oxygenation in waste rock (Amos et al., 2015; Blowes et al., 2014; Lorca et al., 2016). As a result, Fe-(oxyhydr)oxides—ubiquitous in WRSFs—are not available to sorb Mo in water-saturated portions of mine tailings facilities, rendering it more mobile. Because Mo adsorption onto Fe-(oxyhydr)oxides in the BCTMF is therefore limited to the oxic and water-unsaturated embankment sands, but the facility’s water budget is dominated by water-saturated and anoxic flow through the slimes, there is overall less adsorptive attenuation of Mo and commensurate removal of light Mo isotopes in comparison to the oxic WRSFs. Pyrite and clay minerals remain stable [unlike to Fe-(oxyhydroxides)] under anoxic conditions and can also adsorb Mo (Bostick et al., 2003; Sun and Selim, 2018), although there is limited information on Mo adsorption behavior to these phases in anoxic environments and Mo isotopic fractionation factors associated to them are unknown. Assuming that the direction and magnitude of Mo isotopic fractionation factors for adsorption onto clays and pyrite is comparable to known values for the more well-studied (oxyhydr)oxides of Fe and Mn, the lack of isotopic fractionation in Rock Toe seepage in comparison to WRSF drainage suggests that they are not as important in attenuating Mo as (oxyhydr)oxides in mine waste. Experimental constraints on the behavior of Mo isotopes during adsorption onto clays and pyrite under anoxic conditions are necessary to explicitly confirm this hypothesis. In the absence of Mo adsorption, powellite precipitation and/or Fe-    227  (oxo)thiomolybdate provide alternative attenuation pathways under anoxic conditions; however, Mo concentrations in powellite-saturated water remain 1 order of magnitude higher than the recommended drinking water guideline of 0.73 µM (World Health Organization, 2011), and Fe-(oxo)thiomolybdate precipitation, if any, is limited by a lack of available sulfide.  6.5.3 Molybdenum attenuation during oxidation of Fe(II)-rich tailings drainage in the SRD The discharge of anoxic drainage from the Rock Toe into the SRD provided a well-constrained setting to examine the behavior of Mo during oxidation of Fe(II)-bearing waters and concurrent Mo sorption onto (oxyhydr)oxide minerals. The accumulation of Mo and consistent enrichment of light Mo isotopes in the ochreous precipitate samples indicated the importance of sorption to control Mo in this setting. The XAS data for these samples indicated that Mo in the SRD was predominantly attenuated as inner-sphere Td MoVI complexes on ferrihydrite, in both edge-sharing and corner-sharing bonding configurations. Inner-sphere complexation constitutes a strong attenuation mechanism that is less easily reversed than weaker outer-sphere complexation (Xu et al., 2013). It is also worthwhile noting that despite the presence of Fe(II) and Mn(II) at similar concentrations in water entering the SRD, Mn-(oxyhydr)oxides were not a major sink for Mo. Numerous other studies have documented the role of Mn-(oxyhydr)oxides in Mo adsorption and isotopic fractionation (Barling and Anbar, 2004; Wasylenki et al., 2008; Xu et al., 2013). However, Mn-bearing minerals were not observed in XRD and Ramana spectra. Furthermore,     228  the precipitate B-1A from the Mo sorption experiment had 36 wt. % Fe and only 1.0 wt. % Mn; Mn(II) removal from solution was negligible over the six-week duration of the experiment (Figure 6.4)—probably as a result of the slow oxidation rate of Mn(II) at low temperatures (Knocke et al., 1991). The SRD ochreous precipitate samples SRD-S2 and SRD-S4, collected immediately below the Rock Toe, had similar Fe and Mn contents as B-1A, overall indicating that Mn-(oxyhydr)oxides are not responsible for Mo sorption in the SRD pond: ferrihydrite was the main sorbent phase. 6.5.3.1 Determination of isotopic separation factor for sorption to Fe-(oxyhydr)oxides The isotopic separation factor (Δ98Mo) for Mo sorption to poorly crystalline ferrihydrite can be calculated from δ98Mo data for ferrihydrite and co-genetic water samples in our dataset. Sample SRD-S1 was excluded from this calculation because its Mo was sorbed to a more complex mixture of mineral phases. The average Δ98Mo value between aqueous Mo and molybdate-sorbed on ferrihydrite was 1.00 ± 0.40 ‰ (2 SD, = 5) (Table I.1), which matches the previously published value of 1.11 ± 0.15 ‰ for Mo adsorption to synthetic ferrihydrite in laboratory experiments (Goldberg et al. 2009). The wider range in Δ98Mo in our samples is expected because they had poorer crystallinity and possibly contained minor amounts of other Fe-(oxyhydr)oxide phases than lab-synthesized ferrihydrite, leading to a more complex sorption environment. Given the considerable structural and geochemical differences that can be expected between poorly crystalline ferrihydrite precipitating from mine drainage and pure, lab-synthesized ferrihydrite, it is encouraging that our results corroborate a circa 1 ‰ fractionation factor for Mo adsorption onto ferrihydrite as applicable under field conditions. Kinetic isotope     229  effects are unlikely in the SRD pond because isotopic equilibrium for Mo adsorption occurs within 48 hours or less (Goldberg et al., 2009), and longer water contact times were present in the SRD. 6.5.3.2 Molecular mechanism of Mo isotopic fractionation and sorption to Fe-(oxyhydr)oxides Our combination of Mo stable-isotope and XAS data enable exploration of the molecular mechanism causing Mo isotope fractionation during adsorption onto Fe-(oxyhydr)oxides in a mine drainage setting, building upon results from recent investigations of Mo sorption under laboratory conditions (Arai, 2010; Gustafsson and Tiberg, 2015; Kashiwabara et al., 2017, 2009, Wasylenki et al., 2011, 2008). At isotopic equilibrium, heavy isotopes are concentrated in phases that have stiffer bonds (Schauble, 2004). According to Schauble (2004), metal bond stiffness should correlate with higher oxidation states, lower coordination numbers, and/or a change in the neighboring coordinating element to create more covalent-type bonding. A change in Mo coordination among aqueous species has been proposed to explain Mo isotopic fractionation during sorption (Wasylenki et al., 2008). In the hypothesis of Wasylenki et al. (2008), species of Mo with different coordination numbers co-exist in solution and are isotopically fractionated relative to one another; preferential incorporation of one of these species at the mineral surface leads to preservation of isotopic fractionation. For example, acidic hydrothermal submarine waters can contain Oh polymolybdate species (e.g. Mo8O244-, Mo7O246-) which become adsorbed onto Mn-(oxyhydr)oxides (Kuhn et al., 2003). However, polymerization     230  of MoO42- only occurs at pH < 6 and at ΣMo concentrations exceeding 1,000 µM (Torres et al., 2014), conditions which are not found in the SRD. Speciation calculations in PHREEQC show that the Mo in the SRD is dominated by MoO42-(aq) and CaMoO4(aq) with minor NaMoO4-(aq)—all in Td coordination. The circa 1 ‰ isotopic fractionation between our samples of water and sorbed Mo therefore does not appear to be caused by aqueous speciation effects. Furthermore, XAS data rule out a change in oxidation state, bond length and/or coordination number as explanations for the isotopic fractionation that we observed during Mo sorption to ferrihydrite. Only MoVI was observed in XANES (Figure 6.8). Average Mo—O bond lengths determined by EXAFS were 1.76 to 1.77 Å (Table 6.6), which is equal to that of aqueous molybdate (Johansson and Caminiti, 1986). The XANES and EXAFS results for samples SRD-S2, SRD-S4, and B-1A also indicated that the isotopic fractionation was not caused by coordination number changes because Mo remained in Td coordination (Table 6.6). Therefore, a change in coordination—often cited as an important driver of metal stable isotope fractionation during adsorption because it affects bond length (Borg et al., 2012; Bryan et al., 2015; Kashiwabara et al., 2017, 2011)—was not a factor for these samples. Recent experimental work suggests that changes in bond length and/or coordination number are not the only requirement for metal stable-isotope fractionation (Nelson et al., 2017), which echoes our results. For example, an isotopic fractionation of 0.6 ‰ has been observed for 66Zn/64Zn during Zn adsorption to quartz and silica gel, wherein Zn remained in Oh coordination and Zn—O bond lengths did not change (Nelson et al., 2017). We infer that the enrichment of light Mo isotopes in the ferrihydrite structure is driven by greater vibrational movement of Mo—    231  O bonds in sorbed molybdate that is not reflected in bond stretching. Distortion of Mo—O bonds is visible through attenuation of the pre-edge peak in the XANES (Figure 6.8). A change in vibrational movement in aqueous and sorbed molybdate could be driven by a weaker and less covalent Mo—O bond in the latter: the electrostatic force generated by Fe3+ upon the bridging O2- atoms in the ferrihydrite structure (i.e. Mo—O—Fe) exerts a pull on the oxygen’s electron cloud which could come at the expense of the covalent character of the Mo—O. This interaction would lead to greater Mo—O bond vibration and therefore enrich light Mo isotopes in the sorbed phase.   Figure 6.12. Molybdenum isotope ratios against Fe/Mo molar ratios in Fe-(oxyhydr)oxide-rich ochreous precipitate samples collected in the SRD pond.  The Mo isotopic composition of water at the Rock Toe, which is the source of Mo to the SRD, is also shown with the red line and red shading, which indicate its δ98Mo and 2 SD (n = 3). The negative correlation between δ98Mo and Fe/Mo is hypothesized to be explained by the proportion of inner-sphere vs. outer-sphere complexation of Mo: at low Fe/Mo ratios a greater amount of surface adsorption (outer-sphere complexation) with decreased MoO4 molecular distortion decreases the fractionation factor. At high Fe/Mo ratios, there is proportionally more co-precipitation of Mo (inner-sphere complexation), which has a more distorted coordination environment and therefore greater isotopic fractionation. This relationship is decoupled in sample SRD-S1, shown in grey, because its Mo includes octahedral coordination to hematite and clay that drives much larger Mo-O bond distortion and isotopic fractionation.     232  This hypothesis also relates the magnitude in isotopic fractionation between samples to the ratio of inner-sphere and outer-sphere complexes of sorbed Mo. Molecular distortion of inner-sphere complexes of molybdate on ferrihydrite described above appears to drive isotopic fractionation. However, outer-sphere complexes should have smaller isotopic fractionation because the Mo—O bonds of molybdate are only subjected to comparably weaker electrostatic interactions with the hydration shell and with aqueous cations (Kashiwabara et al., 2009). Evidence for the relationship between sorption type (inner vs. outer-sphere) and isotopic fractionation is found in the experiments of Goldberg et al. (2009), where the Mo isotopic fractionation during Mo adsorption onto goethite was inversely correlated to pH: at higher pH proportionally greater outer-sphere complexation decreased the magnitude of Δ98Mo. In our samples, a negative correlation between Fe/Mo molar ratios and Δ98Mo (all samples except SRD-S1—which has a uniquely distinct Mo coordination environment as described previously) is consistent with this hypothesis (Figure 6.12). Samples B-1A, SRD-S2, and SRD-W1 PPTS reflect conditions of rapid precipitation of ferrihydrite upon oxygenation of Rock Toe drainage. The initial rapid precipitation of ferrihydrite colloids from Mo-rich fluids could entrain Mo via a co-precipitation process leading to a greater proportion of inner-sphere Mo which has more molecular distortion that adsorbed outer-sphere Mo, and therefore is more isotopically fractionated. Upon sedimentation of these colloids, Fe/Mo decreases as a result of continued outer-sphere Mo complexation as the ferrihydrite surface is exposed to Mo-rich solution, but co-precipitation and inner-sphere complexation become proportionally less important, which decreases the isotopic fractionation of bulk sorbed Mo. Sample SRD-S4 could therefore have a     233  smaller Δ98Mo as a result of an increase ratio of adsorbed outer-sphere Mo to co-precipitated, inner-sphere Mo, because it was collected further along the flowpath and therefore had a greater likelihood of sorbing Mo as outer-sphere complexes in addition to the initial inner-sphere co-precipitation mechanism.  Nonetheless, this hypothesis highlights the potential impact of pH on Mo sorption mechanisms and isotopic fractionation. At pH = 8, Mo adsorption on ferrihydrite is limited to outer-sphere complexation because both molybdate and ferrihydrite have negative charge which decreases attraction between them and requires bridging H2O atoms for adsorption to take place (Kashiwabara et al., 2009). In the SRD environment, the pH was below the point of zero charge for ferrihydrite, enhancing electrostatic attraction of Mo to the mineral surface and creating conditions that favor the ligand-exchange process required for inner-sphere complexation (Cornell and Schwertmann, 2003). Inner-sphere complexation formed by co-precipitation is further favored by the fact that ferrihydrite nanoparticles are forming in Mo-rich aqueous solution, as opposed to the previous laboratory studies where ferrihydrite was synthesized in the absence of Mo, and adsorption experiments were carried out subsequently (Goldberg et al., 1996; 2009; Kashiwabara et al. 2009; 2011). Indeed, coordination numbers determined for SRD-S2 and B-1A in EXAFS are greater than what is expected for a pure adsorption mechanism which suggests that Mo co-precipitates with ferrihydrite during rapid oxidation of Fe(II) and MoO42--rich fluid. The distribution of outer-sphere and inner-sphere complexes of Mo on ferrihydrite have implications for Mo mobility because outer-sphere complexes are more easily desorbed. In contrast, inner-sphere co-precipitated Mo is more strongly retained within the ferrihydrite     234  structure: this result was confirmed by the fact that in the Mo sorption experiment pH rose from an initial value of 6.6 up to a value 8.1 at the experiment’s end, but Mo was nonetheless retained in ferrihydrite precipitates (B-1A) as inner-sphere complexes. Furthermore, the fractionation of Mo isotopes during inner-sphere complexation makes δ98Mo variation indicative of this sorption mechanism; in contrast, the magnitude of Mo isotopic fractionation for outer-sphere complexation is probably smaller and therefore not as likely to change aqueous δ98Mo (Kashiwabara et al., 2009). Explicit evaluation of this hypothesis would be possible with specific-pH sorption experiments designed to isolate inner- vs. outer-sphere complexation. 6.5.3.3 Multiple bonding configurations in sample SRD-S1 The coordination environment in SRD-S1 with respect to Mo was markedly distinct from the other samples, which is reflected in a decoupling in the relationship between Δ98Mo and Fe/Mo in that sample: a larger Δ98Mo value was found than what would be predicted from the Fe/Mo alone based on the relationship between these parameters found in other ochreous precipitate samples (Figure 6.12). The XANES and EXAFS data for SRD-S1 suggest that a major proportion its Mo is under Oh coordination—a unique attribute of that sample. Octahedral coordination increases average interatomic Mo—O distances and therefore increases molecular vibration that can enrich light isotopes (Kashiwabara et al., 2017, 2011; Schauble, 2004). While it is not possible to confirm the exact linkages of the Oh MoVI in SRD-S1 due to the lack of reference compounds, XANES and EXAFS fitting suggested that it could be explained by coordination to hematite and an Al-bearing phase. Tetrahedral MoVI can transform and incorporate into hematite (identified in Raman spectra) as Oh MoVI during ferrihydrite re-    235  crystallization (Das et al., 2016) and hematite was observed in Raman spectroscopy of that sample. Evidence for an Al—Mo interaction is indicated by the XANES linear-combination fitting of MoVI-adsorbed to Al(OH)3; Al(OH)3 was not found in sample mineralogical analyses but Al is a major component of SRD-S1 (Table 6.4) and XRD showed the presence aluminosilicate clays, which are known to adsorb MoVI (Goldberg et al., 1996; Goldberg and Forster, 1998; Sun and Selim, 2018). Al(OH)3 can plausibly serve as an analogue for other Al-(oxyhydr)oxide surfaces available to adsorb Mo, such as those found in clays. While it is unknown whether Mo adsorption to clays takes Td or Oh coordination, layered double hydroxides (LDH) are a structurally similar phase to clays and do include both coordination environments for Mo (Ma et al., 2017): the proportion of Td and Oh Mo in LDF is governed by surface loading, which forces Mo into either Td coordination on edge and interlayer sites, or Oh coordination within LDH layers. Therefore, Mo sorption to aluminosilicate clay minerals provides a plausible explanation for some of the Oh Mo in SRD-S1. This result suggests that sorption onto clays could also attenuate Mo around the perimeter of the SRD pond where ferrihydrite precipitation is relatively less important.  6.6 Summary and conclusions: Mo attenuation behavior in mine waste Based on the discussion above, we propose a conceptual model for Mo attenuation in mine waste that is strongly dependent upon redox conditions. In general, sulfide mine tailings seepage is more likely to become anoxic than waste-rock drainage (Blowes et al., 2014; Lindsay et al., 2015). This redox difference between tailings and waste rock causes a greater degree of Mo attenuation and isotopic fractionation in the latter, via adsorption onto Fe-(oxyhydr)oxides     236  which are reductively dissolved in anoxic tailings seepage at the Thompson Creek Mine (TCM). More widespread Mo adsorption in waste-rock storage facilities at the TCM is reflected by isotopically heavier δ98Mo values in excess of 1.6 ‰ in waste-rock drainage. In contrast, seepage from the partially anoxic BCTMF had an isotopically lighter δ98Mo of 1.0 ± 0.1 ‰. The smaller isotopic fractionation in tailings seepage is indicative of a modest amount of Mo removal (~ 60 %) in the tailings environment, which could be explained by a combination of powellite precipitation in the anoxic portion of the facility and Mo sorption to ferrihydrite in the oxic flowpath. In addition, precipitation of Fe-(oxo)thiomolybdate phases and/or Mo adsorption onto clays and pyrite in the anoxic flowpath might also be occurring but could not be confirmed without more direct evidence from mineralogical analyses, and could not be evaluated in Mo removal models without better constraints on the isotopic fractionation factor for those mechanisms. The dependence of Mo mobility upon redox conditions in mine waste suggests that long-term understanding of redox evolution is critical to assess Mo leaching during mining operations and in the years and decades following mine closure—especially because transient redox changes in mine waste are not uncommon (e.g. Johnson et al., 2000). Upon discharge of anoxic tailings seepage containing Fe(II), oxygenation of the water causes rapid precipitation of poorly crystalline ferrihydrite, which is responsible for scavenging Mo via a combination of co-precipitation and surface adsorption. During this process, MoO4 tetrahedra form corner-sharing bidentate binuclear and edge-sharing bidentate mononuclear bonds with ferrihydrite. The isotopic fractionation (Δ98Mo) observed during this process is 1.0 ± 0.4 ‰ with preferential removal of light Mo isotopes. The magnitude of Mo isotopic     237  fractionation during this sorption mechanism is hypothesized to be controlled by the degree of inner vs. outer-sphere complexes of Mo, with larger fractionations (and more stable Mo attenuation) associated to the latter as a result of greater distortion of the MoO4 tetrahedra. Further downstream, Mo also appears to be associated to clay minerals and hematite in octahedral coordination, although the significantly greater abundance of ferrihydrite at the present site suggests that these attenuation pathways are of secondary importance. This study provides new insight on Mo geochemistry in mine waste using Mo isotopic analyses and X-ray absorption spectroscopy alongside more conventional techniques. Sorption of Mo onto ferrihydrite and possibly powellite precipitation were the mechanisms of attenuation present in this study; in general rising Mo isotopic compositions in mine waste are indicative of Mo attenuation. Further knowledge of Mo isotopic fractionation factors and molecular-scale sorption mechanisms should continue to improve our ability to provide more quantitative constraints on Mo transport dynamics in mine waste and other environments.     238  CHAPTER 7. CONCLUSION 7.1 Overview of research objective Billions of tons of mine waste are produced annually as a result of extraction of mineral resources that are essential in human societies. The potential for contamination of water resources via the leaching of potentially toxic metals from this waste drives a need to understand the chemical reactions that control the mobility of metals in the environment. The overarching objective of this thesis was to apply stable-isotope analyses of Mo (and Zn)—elements of potential environmental concern in mine waste—to assess the processes controlling geochemical attenuation of these metals. This work was undertaken by (1) developing a protocol suitable for Mo and Zn stable-isotope analyses of mine waste samples; (2) collecting and interpreting geochemical, isotopic, and mineralogical data from mine sites; and (3) conducting laboratory experiments to quantify isotopic fractionation factors for specific mineral precipitation reactions that can attenuate Mo in mine waste. Emphasis was placed on Mo because this element’s chemistry is poorly understood in mine waste and its complex aqueous and isotopic properties make it well suited for isotopic study. 7.2 Summary of main findings 7.2.1 Analytical protocol development Application of Mo and Zn stable isotope data to trace attenuation in mine waste required the development of a suitable analytical protocol including ion-exchange chromatography and     239  isotopic analyses by MC-ICP-MS (Chapter 2). Previous chemical separation methods that were tested in this thesis were unsuitable for samples typical of mine waste that contain high levels of several trace metals (Appendix B). The ion-exchange chromatography developed in Chapter 2 presents a suitable means to extract multiple elements of potential environmental concern—including Mo, Zn, and Cd—through a single chromatographic column. In addition, the double-spike MC-ICP-MS protocol developed for Mo isotopic analyses in this thesis has yielded precise and accurate isotopic compositions for dozens of samples, and several geological reference materials (e.g. BCR-2, SDO-1, seawater, Nod-P-1) and Mo isotope standards. This dataset provides a sizeable contribution to the relatively new field of Mo stable-isotope geochemistry, and the ongoing characterization of the δ98Mo values of reference materials and standards gives a fundamental basis for the validation of Mo isotopic analyses to the international community of isotope geochemists. In general, the analytical protocol developed in Chapter 2 enabled the application of Mo and Zn stable-isotope analyses of mine waste samples and opened the door to their application to trace attenuation. 7.2.2 Application of the stable-isotope fractionation of Mo (and Zn) in mine waste Chapters 3 to 6 showed the occurrence of widespread Mo (and Zn) isotopic fractionation in mine drainage and confirmed that metal stable-isotope data provide evidence for geochemical attenuation at mine sites. This finding is particularly relevant when considering that metal concentrations may decrease due to dilution by clean water during hydrological mixing, whereas a change in δ66Zn or δ98Mo provides a robust indication of attenuation processes. However, because many attenuation processes can occur concurrently and influence the overall isotopic     240  fractionation of these metals, further aqueous and mineralogical data are required to constrain the mechanisms of attenuation. 7.2.2.1 Zinc stable-isotope fractionation and attenuation in mine waste Zinc stable-isotope data from Chapter 3 exemplify some of the strengths and limitations of the application of stable isotope analyses to trace metal attenuation. Variations of 0.7 ‰ in δ66Zn in drainage from mine waste-rock weathering experiments at the Antamina mine could not be explained solely by the isotopic composition of primary Zn-bearing minerals, and therefore confirmed that geochemical attenuation was causing isotopic fractionation. The δ66Zn of mine drainage was lower under alkaline-pH conditions—which favor Zn attenuation via adsorption and precipitation of secondary Zn-bearing (oxyhydr)oxide, hydroxycarbonate, and hydroxysulfate phases. Hydroxycarbonate and hydroxysulfate minerals are confirmed hosts of attenuated Zn in Antamina waste rock (Beckie et al., 2016). However, all of those attenuation processes preferentially remove heavy Zn isotopes (Bryan et al., 2015; Dong and Wasylenki, 2016, 2014; Veeramani et al., 2015), which makes changes in δ66Zn data alone insufficient to confirm which among them is dominant without further constraints from mineralogical analyses. Nonetheless, the finding that Zn stable-isotope ratios evolve following Zn attenuation confirms their use to indicate attenuation. This result expands upon previous studies of Zn isotopes in acid-rock drainage (low-pH) settings, which had concluded that Zn isotopes were more likely to be applicable as tracers of Zn sources—rather than attenuation—as a result of the limited isotopic fractionation that they had found (Aranda et al., 2012; Balistrieri et al., 2008; Fernandez and Borrok, 2009; Matthies et al., 2014). Since the publication of those early studies of δ66Zn in     241  mine drainage, there has been growing research—much of it concurrent with this thesis—confirming the occurrence of isotopic fractionation during a variety of Zn attenuation processes (Balistrieri et al., 2008; Bryan et al., 2015; Dong and Wasylenki, 2016; Ducher et al., 2018; Jamieson-Hanes et al., 2017; Veeramani et al., 2015; Wanty et al., 2013a; 2013b). 7.2.2.2 Molybdenum stable-isotope fractionation and attenuation in mine waste This thesis provided new insight on the fate of Mo in mine waste by applying Mo stable-isotope analyses. Only a handful of works had previously studied Mo in mine waste—none had applied Mo stable-isotope analyses (Frascoli and Hudson-Edwards, 2018). That knowledge gap existed despite the abundant evidence of processes that fractionate Mo isotopes under Earth-surface conditions (Kendall et al., 2017) and the potential environmental concerns associated with elevated Mo concentrations in mine-impacted waters (Smedley and Kinniburgh, 2017). As with Zn, variation of Mo stable-isotope ratios in mine drainage could not be explained purely by source-mineral isotopic compositions: a general conclusion is that geochemical attenuation causes increasing δ98Mo in solution (Figure 7.1 and Chapters 3 to 6). This observation was first described at the Antamina mine, where δ98Mo in mine drainage from waste-rock weathering experiments varied by more than 2 ‰ and was shifted towards isotopically heavy values in comparison to waste rock and molybdenite (Chapter 3). Similar offsets in δ98Mo between mine drainage and solid mine waste were also observed in full-scale waste-rock storage facilities at the Thompson Creek Mine (TCM; Chapter 4). Adsorption was identified as an important mechanism driving Mo attenuation and isotopic fractionation in waste     242  rock based on the discovery of isotopically light Mo found at high abundance in chemical sequential extractions targeting Fe-(oxyhydr)oxide-rich weathered waste rock at the TCM. The δ98Mo data from drainage suggested that a large degree (> 90 %) of Mo removal was occurring via adsorption from mine drainage over a pH range of ~ 4 to ~ 8, indicating that this process is not restricted to acid-mine drainage conditions as is generally assumed (Frascoli and Hudson-Edwards, 2018). However, the efficacy of adsorption onto Fe-(oxyhydr)oxides to attenuate Mo is contingent upon redox conditions: in anoxic mine tailings drainage Mo is more mobile as a result of a lack of Fe-(oxyhydr)oxide adsorption surfaces (Chapter 6). This change was reflected by higher Mo concentrations and less Mo isotopic fractionation in a full-scale anoxic tailings management facility at the TCM (Chapter 6).      243    Figure 7.1. Summary of Mo attenuation processes and the direction of associated isotopic fractionation (α). Values of α < 1 indicate that light Mo isotopes are preferentially removed, causing increasing δ98Mo in residual aqueous solution. The general aqueous environments under which attenuation mechanisms are grouped as : strong (solid black lines), weak (dotted black lines), or negligible (thin dotted lines). Gray shading is used for conditions that have not been experimentally tested. This figure encompasses results of multiple studies on Mo, including Chapters 4, 5, and 6 of this thesis, and: Barling and Anbar 2001; Bostick et al. 2003; Goldberg et al. 2009; Goldberg et al. 2008; Helz et al. 1996; 2011; Kashiwabara et al. 2009; 2011; 2017; King et al. 2018; Nägler et al. 2011;  Sun and Selim 2018; Wasylenki et al. 2008; 2011; Tossel et al. 2005; and Xu et al. 2013. Question marks indicate uncertain isotopic and/or unknown fractionation factors. The abbreviations ''I.S.'' and ''O.S.'' represent inner-sphere and outer-sphere Mo complexation, respectively.     244  In addition to adsorption, the precipitation of molybdate minerals can also attenuate Mo and fractionate Mo isotopes in mine drainage (Chapter 5). This relationship was confirmed for powellite and wulfenite precipitation, and is most likely also present for other molybdate minerals that were not studied in this thesis [e.g. ferrimolybdite, nickel(II) molybdate] although this hypothesis is at present untested. The provision of Mo isotopic fractionation factors for powellite and wulfenite precipitation is critical to the application of Mo stable-isotope data in mine waste (Chapters 3 and 6), where these minerals commonly form (Conlan et al., 2012). Results from Chapter 5 also draw attention to a process of Mo isotopic fractionation that was previously unknown: the pre-existing literature was limited to interpreting δ98Mo data in terms of other mechanisms—adsorption, (oxo)thiomolybdate-molybdate interactions, and hydrothermal processes—as causes for variations of δ98Mo on Earth (Kendall et al., 2017; Smedley and Kinniburgh, 2017). Consequently, the role of molybdate mineral precipitation should hitherto be included when interpreting δ98Mo signatures in natural and contaminated environments where this process is plausible—a consideration that was not possible prior to this investigation. Further research will undoubtedly unveil additional processes that fractionate Mo stable isotopes and that have not yet been considered when interpreting δ98Mo data. A final contribution of this thesis was insight into the molecular mechanisms for Mo attenuation and isotopic fractionation during sorption onto Fe-(oxyhydr)oxides characterized in Chapter 6. Only a handful of studies had previously investigated molecular mechanisms of Mo sorption using field samples (Gustafsson and Tiberg, 2015). The ~ 1 ‰ isotopic fractionation observed for Mo sorption onto ferrihydrite forming under field conditions corroborated a     245  previous laboratory study that had used synthetic ferrihydrite as a sorbent (Goldberg et al., 2009). Molybdenum was mainly found in ferrihydrite as inner-sphere complexes, which are strongly bonded to the ferrihydrite structure (Xu et al., 2013) and therefore represent an effective attenuation mechanism. X-ray absorption spectroscopy (XAS) analyses confirmed that there was no change in Mo oxidation state, coordination number, or Mo—O bond lengths during this sorption process. These parameters were probably also unchanged during the precipitation of wulfenite and powellite in Chapter 5, suggesting that minor molecular distortions of molybdate tetrahedra during sorption and mineral precipitation can cause resolvable metal stable-isotope fractionation. This conclusion indicates the presence of molecular mechanisms of isotope fractionation beyond the more classical explanations that have been widely cited (i.e. oxidation state, coordination number, bond lengths; Schauble 2004). Future studies combining XAS, isotopic analyses and thermodynamic modelling of bond lengths should continue to constrain mechanisms of metal isotope fractionation and attenuation to new levels of detail. 7.3 Future research directions This thesis provided several examples of information gained through Mo and Zn isotopic data that augmented knowledge gained through other analytical methods. An overarching conclusion was that metal stable-isotope fractionation presently serves to indicate metal attenuation, but that the precise mechanism(s) of attenuation must be constrained with supporting evidence from aqueous geochemistry and mineralogy. Once these complementary techniques are used in conjunction, quantitative estimates of metal removal become possible through a modeling approach on the basis of metal-isotope data (Figure 7.2).     246   Figure 7.2. Conceptual flow-chart demonstrating the implementation of metal stable-isotope analyses to trace attenuation along with other techniques. The δi/jM refers to isotopic composition of a metal M with isotopes i and j.      247  Going forward, metal stable-isotope data will be useful to assess the mobility of metals in the environment, but their interpretation is contingent on an understanding of the fundamental processes causing isotopic fractionation—which are just beginning to be unraveled through experimental, theoretical, and field-based studies. Many—if not most—geochemical reactions involving metals can cause resolvable metal stable-isotope fractionation; much work is needed to quantify fractionation factors associated to these reactions and how they might vary under different geochemical conditions. For example, both kinetic and equilibrium isotope fractionation can be present in a single attenuation reaction and have considerably different effects on the isotopic composition of the residual fluid, as shown for powellite precipitation (Chapter 5). Molybdenum isotopic fractionation during (ad)sorption is probably linked to (pH-dependent) mechanisms of surface attachment, including coordination number changes and inner- vs. outer-sphere complexation (Chapter 6; Goldberg et al., 2009; Kashiwabara et al., 2017, 2009; Wasylenki et al., 2008). There is also little information available on pathways of Mo attenuation and isotopic fractionation under anoxic conditions, including adsorption of molybdate on clays and sulfides which may constitute alternative sorption surfaces in the absence of Fe-(oxyhydr)oxides. The magnitude of Mo and Zn isotopic fractionation during oxidative dissolution of sulfides—the starting point for release of these elements in mine waste—remains relatively unknown, in part due to the difficulty in determining it experimentally (Fernandez and Borrok, 2009; Liermann et al., 2011). Another important consideration in interpreting metal isotope data under field conditions is the role of hydrodynamic mixing, which causes localized isotopic fractionations to be averaged over hydrological transport domains     248  (Druhan and Maher, 2017). Given the diversity and complexity of controls on metal stable-isotope fractionation, reactive transport models that include metal stable-isotope fractionation present a promising avenue of research to quantify metal removal (e.g. Jamieson-Hanes et al., 2012). To conclude, metal stable-isotope data constitute an invaluable and emerging tracer of metal mobility—a welcome new tool whose application should continue to expand given the environmental and socio-economical ramifications tied to metal release, transport, and attenuation.       249  References Achary, S.N., Patwe, S.J., Mathews, M.D., Tyagi, A.K., 2006. High temperature crystal chemistry and thermal expansion of synthetic powellite (CaMoO4): A high temperature X-ray diffraction (HT-XRD) study. J. Phys. Chem. Solids 67, 774–781. doi:10.1016/j.jpcs.2005.11.009 Aranda, S., Borrok, D.M., Wanty, R.B., Balistrieri, L.S., 2012. Zinc isotope investigation of surface and pore waters in a mountain watershed impacted by acid rock drainage. Sci. Total Environ. 420, 202–13. doi:10.1016/j.scitotenv.2012.01.015 Al, T.A., Martin, C.J., Blowes, D.W., 2000. Carbonate-mineral / water interactions in sulfide-rich mine tailings. Geochim. Cosmochim. Acta 64, 3933–3948. Albarède, F., Beard, B., 2004. Analytical Methods for Non-Traditional Isotopes. Rev. Mineral. Geochemistry 55, 113–152. Albarède, F., Télouk, P., Blichert-Toft, J., Boyet, M., Agranier, A., Nelson, B., 2004. Precise and accurate isotopic measurements using multiple-collector ICPMS. Geochim. Cosmochim. Acta 68, 2725–2744. doi:10.1016/j.gca.2003.11.024 Amos, R.T., Blowes, D.W., Bailey, B.L., Sego, D.C., Smith, L., Ritchie, A.I.M., 2015. Waste-rock hydrogeology and geochemistry. Appl. Geochemistry 57, 140–156. doi:10.1016/j.apgeochem.2014.06.020 Anbar, A.D., Knab, K., Barling, J., 2001. Precise determination of mass-dependent variations in the isotopic composition of molybdenum using MC-ICPMS. Anal. Chem. 73, 1425–31. Arai, Y., 2010. X-ray Absorption Spectroscopic Investigation of Molybdenum Multinuclear     250  Sorption Mechanism at the Goethite - Water Interface. Environ. Sci. Technol. Lett. 44, 8491–8491. Aranda, C., 2010. Assessment of waste rock weathering characteristics at the Antamina mine based on field cell experiments. MASc thesis, University of British Columbia, 269 p. Aranda, S., Borrok, D.M., Wanty, R.B., Balistrieri, L.S., 2012. Zinc isotope investigation of surface and pore waters in a mountain watershed impacted by acid rock drainage. Sci. Total Environ. 420, 202–13. doi:10.1016/j.scitotenv.2012.01.015 Archer, C., 2007. The application of transition metal isotope systems to biogeochemical studies of the early Earth. PhD Thesis, Royal Holloway University of London, 342 p. Archer, C., Vance, D., 2008. The isotopic signature of the global riverine molybdenum flux and anoxia in the ancient oceans. Nat. Geosci. 1, 597–600. doi:10.1038/ngeo282 Arnold, G.L., Anbar, A.D., Barling, J., Lyons, T.W., 2004. Molybdenum isotope evidence for widespread anoxia in mid-Proterozoic oceans. Science 304, 87–90. doi:10.1126/science.1091785 Arnold, T., Schönbächler, M., Rehkämper, M., Dong, S., Zhao, F.-J., Kirk, G.J.D., Coles, B.J., Weiss, D.J., 2010. Measurement of zinc stable isotope ratios in biogeochemical matrices by double-spike MC-ICPMS and determination of the isotope ratio pool available for plants from soil. Anal. Bioanal. Chem. 398, 3115–25. doi:10.1007/s00216-010-4231-5 Asael, D., Tissot, F.L.H., Reinhard, C.T., Rouxel, O., Dauphas, N., Lyons, T.W., Ponzevera, E., Liorzou, C., Chéron, S., 2013. Coupled molybdenum, iron and uranium stable isotopes as oceanic paleoredox proxies during the Paleoproterozoic Shunga Event. Chem. Geol. 362,     251  193–210. doi:10.1016/j.chemgeo.2013.08.003 Aubertin, M, Bussière, B, Chapuis, R.P., 1996. Hydraulic conductivity of homogenized tailings from hard rock mines. Can. Geotech. J. 33, 470–482. Axelsson, M.D., Rodushkin, I., Ingri, J., Ohlander, B., 2002. Multielemental analysis of Mn-Fe nodules by ICP-MS: optimisation of analytical method. Analyst 127, 76–82. doi:10.1039/b105706p Balistrieri, L.S., Borrok, D.M., Wanty, R.B., Ridley, W.I., 2008. Fractionation of Cu and Zn isotopes during adsorption onto amorphous Fe(III) oxyhydroxide: Experimental mixing of acid rock drainage and ambient river water. Geochim. Cosmochim. Acta 72, 311–328. doi:10.1016/j.gca.2007.11.013 Barceloux, D.G., 1999. Molybdenum. Clin. Toxicol. 37, 231–237. Barling, J., Arnold, G., Anbar, A.D., 2001. Natural mass-dependent variations in the isotopic composition of molybdenum. Earth Planet. Sci. Lett. 193, 447–457. doi:10.1016/S0012-821X(01)00514-3 Barling, J., Anbar, A.D., 2004. Molybdenum isotope fractionation during adsorption by manganese oxides. Earth Planet. Sci. Lett. 217, 315–329. doi:10.1016/S0012-821X(03)00608-3 Barling, J., Weis, D., 2012. An isotopic perspective on mass bias and matrix effects in multi-collector inductively-coupled-plasma mass spectrometry. J. Anal. At. Spectrom. 27, 653. doi:10.1039/c2ja10382f Barling, J., Weis, D., 2008. Influence of non-spectral matrix effects on the accuracy of Pb     252  isotope ratio measurement by MC-ICP-MS: implications for the external normalization method of instrumental mass bias correction. J. Anal. At. Spectrom. 23, 1017. doi:10.1039/b717418g Bay, D.S., 2009. Hydrological and hydrogeochemical characteristics of neutral drainage from a waste rock test pile. MSc Thesis, University of British Columbia, 331 p. Beckie, R.D., Aranda, C., Blackmore, S.R., Peterson, H.E., Hirsche, T., Javadi, M., Blaskovich, R., Haupt, C., Conlan, M., Bay, D., Harrison, B., Brienne, S., Klein, B., Mayer, K.U., 2011. A study of the mineralogical , hydrological and biogeochemical controls on drainage from waste rock at the Antamina Mine, Peru : An overview, in: Proceedings of the Tailings and Mine Waste Conference, Vancouver, Canada, November 6-9, 2011. Beckie, R.D., Javadi, M., Laurenzi, L., Lorca, M.E., Skierszkan, E.K., Mayer, K.U., St-Arnault, M., 2016. Antamina waste rock study – 2016 lessons learned report. Beard, B.L., Johnson, C.M., Cox, L., Sun, H., Nealson, K.H., Beard, B.L., Johnson, C.M., Cox, L., Sun, H., Nealson, K.H., Aguilar, C., 1999. Iron Isotope Biosignatures. Science 80, 285, 1889–1892. Bermin, J., Vance, D., Archer, C., Statham, P.J., 2006. The determination of the isotopic composition of Cu and Zn in seawater. Chem. Geol. 226, 280–297. doi:10.1016/j.chemgeo.2005.09.025 Bigeleisen, J., Mayer, M.G., 1947. Calculation of Equilibrium Constants for Isotopic Exchange Reactions. J. Chem. Phys. 15, 261–267. doi:10.1063/1.1746492 Bigham, J.M., Nordstrom, D.K., 2001. Iron and aluminum hydroxysulfates in acid sulfate waters.     253  Rev. Mineral. Geochemistry 40, 351–403. doi:10.2138/rmg.2000.40.7 Bigham, J.M., Schwertmann, U., Traina, S.J., Winland, R.L., Wolf, M., 1996. Schwertmannite and the chemical modelling of iron in acid sulfate waters. Geochim. Cosmochim. Acta 60, 2111–2121. Bissonnette, J., Essilfie-Dughan, J., Moldovan, B.J., Hendry, M.J., 2016. Sequestration of As and Mo in uranium mill precipitates (pH 1.5-9.2): An XAS study. Appl. Geochemistry 72, 30–33. doi:10.1016/j.apgeochem.2016.06.007 Blackmore, S., Smith, L., Mayer, K.U., Beckie, R.D., 2014. Comparison of unsaturated flow and solute transport through waste rock at two experimental scales using temporal moments and numerical modeling. J. Contam. Hydrol. 171, 49–65. doi:10.1016/j.jconhyd.2014.10.009 Blanchard, P.E.R., Hayes, J.R., Grosvenor, A.P., Rowson, J., Hughes, K., Brown, C., 2015. Investigating the geochemical model for molybdenum mineralization in the JEB tailings management facility at McClean Lake, Saskatchewan: An X-ray absorption spectroscopy study. Environ. Sci. Technol. 49, 6504–6509. doi:10.1021/acs.est.5b00528 Blowes, D.W., Ptacek, C.J., Jambor, J.L., Weisener, C.G., Paktunc, D., Gould, W.D., Johnson, D.B., 2014. The Geochemistry of Acid Mine Drainage, 11th ed, Treatise on Geochemistry: Second Edition. Elsevier Ltd. doi:10.1016/B978-0-08-095975-7.00905-0 Bonnand, P., Parkinson, I.J., James, R.H., Karjalainen, A.-M., Fehr, M.A., 2011. Accurate and precise determination of stable Cr isotope compositions in carbonates by double spike MC-ICP-MS. J. Anal. At. Spectrom. 26, 528–535. doi:10.1039/c0ja00167h Borg, S., Liu, W., Etschmann, B., Tian, Y., Brugger, J., 2012. An XAS study of molybdenum     254  speciation in hydrothermal chloride solutions from 25-385°C and 600 bar. Geochim. Cosmochim. Acta 92, 292–307. doi:10.1016/j.gca.2012.06.001 Borrok, D.M., Nimick, D.A., Wanty, R.B., Ridley, W.I., 2008. Isotopic variations of dissolved copper and zinc in stream waters affected by historical mining. Geochim. Cosmochim. Acta 72, 329–344. doi:10.1016/j.gca.2007.11.014 Borrok, D.M., Wanty, R.B., Ridley, W.I., Lamothe, P.J., Kimball, B.A., Verplanck, P.L., Runkel, R.L., 2009. Application of iron and zinc isotopes to track the sources and mechanisms of metal loading in a mountain watershed. Appl. Geochemistry 24, 1270–1277. doi:10.1016/j.apgeochem.2009.03.010 Bostick, B.C., Fendorf, S., Helz, G.R., 2003. Differential adsorption of molybdate and tetrathiomolybdate on pyrite (FeS2). Environ. Sci. Technol. 37, 285–91. Breillat, N., Guerrot, C., Marcoux, E., Négrel, P., 2016. A new global database of δ98Mo in molybdenites: A literature review and new data. J. Geochemical Explor. 161, 1–15. doi:10.1016/j.gexplo.2015.07.019 Brinza, L., Benning, L.G., Statham, P.J., 2008. Adsorption studies of Mo and V onto ferrihydrite. Mineral. Mag. 72, 385–388. doi:10.1180/minmag.2008.072.1.385 Bryan, A.L., Dong, S., Wilkes, E.B., Wasylenki, L.E., 2015. Zinc isotope fractionation during adsorption onto Mn oxyhydroxide at low and high ionic strength. Geochim. Cosmochim. Acta 157, 182–197. doi:10.1016/j.gca.2015.01.026 Buchachenko, A.L., 2013. Mass-Independent Isotope Effects. J. Phys. Chem. 117, 2231–2238. Bura-Nakić, E., Andersen, M.B., Archer, C., de Souza, G.F., Marguš, M., Vance, D., 2018.     255  Coupled Mo-U abundances and isotopes in a small marine euxinic basin: Constraints on processes in euxinic basins. Geochim. Cosmochim. Acta 222, 212–229. doi:10.1016/j.gca.2017.10.023 Carazao Gallegos, J.C., 2007. The design, construction, instrumentation and initial response of a field-scale waste rock pile exeriment. MASc Thesis, University of British Columbia, 247 p. CCME (Canadian Council of Ministers of the Environment). (1999a) Canadian Water Quality Guidelines for the Protection of Aquatic Life - Molybdenum. CCME (Canadian Council of Ministers of the Environment). (1999b) Canadian Soil Quality Guidelines for the Protection of Environmental and Human Health - Zinc. Chappaz, A., Lyons, T.W., Gregory, D.D., Reinhard, C.T., Gill, B.C., Li, C., Large, R.R., 2014. Does pyrite act as an important host for molybdenum in modern and ancient euxinic sediments? Geochim. Cosmochim. Acta 126, 112–122. doi:10.1016/j.gca.2013.10.028 Chen, X., Romaniello, S.J., Herrmann, A.D., Wasylenki, L.E., Anbar, A.D., 2016. Uranium isotope fractionation during coprecipitation with aragonite and calcite. Geochim. Cosmochim. Acta 188, 189–207. doi:10.1016/j.gca.2016.05.022 Clark, I.D., Fritz, P., 1997. Environmental Isotopes in Hydrogeology. CRC Press. Cloquet, C., Carignan, J., Lehmann, M.F., Vanhaecke, F., 2008. Variation in the isotopic composition of zinc in the natural environment and the use of zinc isotopes in biogeosciences: a review. Anal. Bioanal. Chem. 390, 451–63. doi:10.1007/s00216-007-1635-y Compston, W., Oversby, V.M., 1969. Lead isotopic analysis using a double spike. J.     256  Geophyiscal Res. 74, 4338–4348. Conlan, M.J.W., Mayer, K.U., Blaskovich, R., Beckie, R.D., 2012. Solubility controls for molybdenum in neutral rock drainage. Geochemistry Explor. Environ. Anal. 12, 21–32. doi:10.1144/1467-7873/10-RA-043 Connelly, J.N., Ulfbeck, D.G., Thrane, K., Bizzarro, M., Housh, T., 2006. A method for purifying Lu and Hf for analyses by MC-ICP-MS using TODGA resin. Chem. Geol. 233, 126–136. doi:10.1016/j.chemgeo.2006.02.020 Cora, I., Czugler, M., Dódony, I., Rečnik, A., 2011. On the symmetry of wulfenite (Pb[MoO4]) from Mežica (Slovenia). Acta Crystallogr. Sect. C Cryst. Struct. Commun. 67, 33–35. doi:10.1107/S0108270111015769 Cornell, R.M., Schwertmann, U., 2003. The Iron Oxides: Structure, Properties, Reactions, Occurences and Uses, Second Edition. Wiley-VCH Verslag GmbH & Co. KGaA. doi:10.1180/minmag.1997.061.408.20 Creech, J., Baker, J., Handler, M., Schiller, M., Bizzarro, M., 2013. Platinum stable isotope ratio measurements by double-spike multiple collector ICPMS. J. Anal. At. Spectrom. 28, 853–865. doi:10.1039/c3ja50022e Dahl, T.W., Anbar, A.D., Gordon, G.W., Rosing, M.T., Frei, R., Canfield, D.E., 2010. The behavior of molybdenum and its isotopes across the chemocline and in the sediments of sulfidic Lake Cadagno, Switzerland. Geochim. Cosmochim. Acta 74, 144–163. doi:10.1016/j.gca.2009.09.018     257  Das, S., Hendry, M.J., 2011. Application of Raman spectroscopy to identify iron minerals commonly found in mine wastes. Chem. Geol. 290, 101–108. doi:10.1016/j.chemgeo.2011.09.001 Das, S., Essilfie-Dughan, J., Jim Hendry, M., 2016. Sequestration of molybdate during transformation of 2-line ferrihydrite under alkaline conditions. Appl. Geochemistry 73, 70–80. doi:10.1016/j.apgeochem.2016.08.003 Dauphas, N., John, S.G., Rouxel, O., 2017. Iron Isotope Systematics. Rev. Mineral. Geochemistry 82, 415–510. doi:10.2138/rmg.2017.82.11 de Laeter, J.R., Hidaka, H., Peiser, H.S., Rosman, K.J.R., Taylor, P.D.P., 2003. Atomic Weights of the Elements: Review 2000 (IUPAC Technical Report). Pure Appl. Chem. 75, 683–800. DePaolo, D.J., 2011. Surface kinetic model for isotopic and trace element fractionation during precipitation of calcite from aqueous solutions. Geochim. Cosmochim. Acta 75, 1039–1056. doi:10.1016/j.gca.2010.11.020 Di Tommaso, D., De Leeuw, N.H., 2010. Structure and dynamics of the hydrated magnesium ion and of the solvated magnesium carbonates: Insights from first principles simulations. Phys. Chem. Chem. Phys. 12, 894–901. doi:10.1039/b915329b Dockrey J. W. 2010. Microbiology and geochemistry of neutral pH waste rock from the Antamina Mine, Peru. MSc Thesis, University of British Columbia. 225 p. Dockrey, J., Lindsay, M., Mayer, K., Beckie, R., Norlund, K., Warren, L., Southam, G., 2014. Acidic Microenvironments in Waste Rock Characterized by Neutral Drainage: Bacteria–Mineral Interactions at Sulfide Surfaces. Minerals 4, 170–190. doi:10.3390/min4010170     258  Dockrey, J.W., Stockwell, J.S., 2012. Early Indicators of Acid Seepage Generation – A comparative study of long-term seepage quality from two waste rock dumps in the Central Rocky Mountains, in: Proceedings of the 9th International Conference on Acid Rock Drainage, May 20-25, Ottawa, Canada. Dong, S., Wasylenki, L.E., 2016. Zinc isotope fractionation during adsorption to calcite at high and low ionic strength. Chem. Geol. 447, 70–78. doi:10.1016/j.gca.2015.01.026 Dong, S., Wasylenki, L.E., 2014. Zinc isotope fractionation during adsorption and incorporation with calcite, Goldschmidt. Geochim. Cosmochim. Acta 84, 3943. Dótor-Almazán, A., Armienta-Hernández, M.A., Talavera-Mendoza, O., Ruiz, J., 2017. Geochemical behavior of Cu and sulfur isotopes in the tropical mining region of Taxco, Guerrero (southern Mexico). Chem. Geol. 471, 1–12. doi:10.1016/j.chemgeo.2017.09.005 Druhan, J.L., Brown, S.T., Huber, C., 2015. Isotopic Gradients Across Fluid–Mineral Boundaries. Rev. Mineral. Geochemistry 80, 355–391. doi:10.2138/rmg.2015.80.11 Druhan, J.L., Maher, K., 2017. The influence of mixing on stable isotope ratios in porous media: A revised Rayleigh model. Water Resour. Res. 53, 1101–1124. doi:10.1002/ 2016WR019666.  Dragovich, D., 2006. Microchemistry of Small Desert Varnish Samples, Western New South Wales , Australia 23, 445–453. Ducher, M., Blanchard, M., Balan, E., 2018. Equilibrium isotopic fractionation between aqueous Zn and minerals from first-principles calculations. Chem. Geol. 483, 342–350. doi:10.1016/j.chemgeo.2018.02.040     259  Edmond, J.M., 1970. High precision determination of titration alkalinity and total carbon dioxide content of seawater by potentiometric titration. Deep. Res. Part I 17, 737–750. Egal, M., Elbaz-Poulichet, F., Casiot, C., Motelica-Heino, M., Négrel, P., Bruneel, O., Sarmiento, A.M., Nieto, J.M., 2008. Iron isotopes in acid mine waters and iron-rich solids from the Tinto-Odiel Basin (Iberian Pyrite Belt, Southwest Spain). Chem. Geol. 253, 162–171. doi:10.1016/j.chemgeo.2008.05.006 Ehrlich, S., Butler, I., Halicz, L., Rickard, D., Oldroyd, A., Matthews, A., 2004. Experimental study of the copper isotope fractionation between aqueous Cu(II) and covellite, CuS. Chem Geol. 209, 259–269. doi:10.1016/j.chemgeo.2004.06.010 Ellis, A.S., Johnson, T.M., Bullen, T.D., 2002. Chromium isotopes and the fate of hexavalent chromium in the environment. Science 295, 2060–2. doi:10.1126/science.1068368 Erickson, B.E., Helz, G.R., 2000. Molybdenum(VI) speciation in sulfidic waters: Stability and lability of thiomolybdates. Geochim. Cosmochim. Acta 64, 1149–1158. doi:10.1016/S0016-7037(99)00423-8 Essilfie-Dughan, J., Pickering, I.J., Hendry, M.J., George, G.N., Kotzer, T., 2011. Molybdenum speciation in uranium mine tailings using X-ray absorption spectroscopy. Environ. Sci. Technol. 45, 455–460. doi:10.1021/es102954b Essington, M.E., 1990. Calcium Molybdate Solubility in Spent Oil Shale and a Preliminary Evaluation of the Association Constants for the Formation of CaMoO4(aq), KMoO4-(aq), and NaMoO4-(aq). Environ. Sci. Technol. 24, 214–220. Eugster, O., Tera, F., Wasserberg, G.J., 1969. Isotopic analyses of barium in meteorites and in     260  terrestrial samples. J. Geophys. Res. 74, 3897–3908. Fernandez, A., Borrok, D.M., 2009. Fractionation of Cu, Fe, and Zn isotopes during the oxidative weathering of sulfide-rich rocks. Chem. Geol. 264, 1–12. doi:10.1016/j.chemgeo.2009.01.024 Foucher, D., Ogrinc, N., Hintelmann, H., 2009. Tracing mercury contamination from the Idrija mining region (Slovenia) to the Gulf of Trieste using Hg isotope ratio measurements. Environ. Sci. Technol. 43, 33–9. Fuller, C.C., Bargar, J.R., 2014. Processes of zinc attenuation by biogenic manganese oxides forming in the hyporheic zone of Pinal Creek, Arizona. Environ. Sci. Technol. 48, 2165–2172. doi:10.1021/es402576f Goldberg, S., Forster, H.S., Godfrey, C.L., 1996. Molybdenum Adsorption on Oxides, Clay Minerals, and Soils. Soil Sci. Soc. Am. J. 60, 425–432. doi:10.2136/sssaj1996.03615995006000020013x Goldberg, S., Forster, H.S., 1998. Factors Affecting Molybdenum Adsorption By Soils and Minerals. Soil Sci. 163, 109–114. doi:10.1097/00010694-199802000-00004 Goldberg, S., Johnston, C.T., Suarez, D.L., Lesch, S.M., 2008. Mechanism of Molybdenum Adsorption on Soil Minerals Evaluated Using Vibrational Spectroscopy and Surface Complexation Modeling, in: Barnett, M.O., Kent, D.B. (Eds.), Developments in Earth & Environmental Sciences. pp. 235–266. Goldberg, T., Archer, C., Vance, D., Poulton, S.W., 2009. Mo isotope fractionation during adsorption to Fe (oxyhydr)oxides. Geochim. Cosmochim. Acta 73, 6502–6516.     261  doi:10.1016/j.gca.2009.08.004 Goldberg, T., Gordon, G., Izon, G., Archer, C., Pearce, C.R., McManus, J., Anbar, A.D., Rehkämper, M., 2013. Resolution of inter-laboratory discrepancies in Mo isotope data: an intercalibration. J. Anal. At. Spectrom. 28, 724–735. doi:10.1039/c3ja30375f Golder Associates, 2018.  https://www.golder.ca/index.php/modules.php?name=Projects&service_id=242&sector_id=0&sort_by=date&sort_dir=desc&page=1&sp_id=27. Visited February 27, 2018. Gomes, M.L., Johnston, D.T., 2017. Oxygen and sulfur isotopes in sulfate in modern euxinic systems with implications for evaluating the extent of euxinia in ancient oceans. Geochim. Cosmochim. Acta 205, 331–359. doi:10.1016/j.gca.2017.02.020 Goto, K.T., Shimoda, G., Anbar, A.D., Gordon, G.W., Harigane, Y., Senda, R., Suzuki, K., 2015. Molybdenum isotopes in hydrothermal manganese crust from the Ryukyu arc system: Implications for the source of molybdenum. Mar. Geol. 369, 91–99. doi:10.1016/j.margeo.2015.08.007 Goumih, A., El Adnani, M., Hakkou, R., Benzaazoua, M., 2013. Geochemical Behavior of Mine Tailings and Waste Rock at the Abandoned Cu-Mo-W Azegour Mine (Occidental High Atlas, Morocco). Mine Water Environ. 32, 121–132. doi:10.1007/s10230-013-0221-0 Greaney, A.T., Rudnick, R.L., Gaschnig, R.M., 2016. Crustal Sources of Molybdenum, in: Goldschmidt Conference, June 2016, Yokohama, Japan. Greber, N.D., Hofmann, B.A., Voegelin, A.R., Villa, I.M., Nägler, T.F., 2011. Mo isotope composition in Mo-rich high- and low-T hydrothermal systems from the Swiss Alps.     262  Geochim. Cosmochim. Acta 75, 6600–6609. doi:10.1016/j.gca.2011.08.034 Greber, N.D., Siebert, C., Nägler, T.F., Pettke, T., 2012. δ98/95Mo values and Molybdenum Concentration Data for NIST SRM 610, 612 and 3134: Towards a Common Protocol for Reporting Mo Data. Geostand. Geoanalytical Res. 36, 291–300. doi:10.1111/j.1751-908X.2012.00160.x Greber, N.D., Pettke, T., Nägler, T.F., 2014. Magmatic–hydrothermal molybdenum isotope fractionation and its relevance to the igneous crustal signature. Lithos 190–191, 104–110. doi:10.1016/j.lithos.2013.11.006 Greber, N.D., Puchtel, I.S., Nägler, T.F., Mezger, K., 2015. Komatiites constrain molybdenum isotope composition of the Earth’s mantle. Earth Planet. Sci. Lett. 421, 129–138. doi:10.1016/j.epsl.2015.03.051 Guilbaud, R., Butler, I.B., Ellam, R.M., Rickard, D., Oldroyd, A., 2011. Experimental determination of the equilibrium Fe isotope fractionation between Feaq2+ and FeS (mackinawite) at 25 and 2°C. Geochim. Cosmochim. Acta 75, 2721–2734. doi:10.1016/j.gca.2011.02.023 Guinoiseau, D., Gélabert, A., Moureau, J., Louvat, P., Benedetti, M.F., 2016. Zn Isotope Fractionation during Sorption onto Kaolinite. Environ. Sci. Technol. 50, 1844–1852. doi:10.1021/acs.est.5b05347 Gustafsson, J.P., 2003. Modelling molybdate and tungstate adsorption to ferrihydrite. Chem. Geol. 200, 105–115. doi:10.1016/S0009-2541(03)00161-X Gustafsson, J.P., Tiberg, C., 2015. Molybdenum binding to soil constituents in acid soils: An     263  XAS and modelling study. Chem. Geol. 417, 279–288. doi:10.1016/j.chemgeo.2015.10.016 Hall, G.E.M., Vaive, J.E., Beer, R., Hoashi, M., 1996. Selective leaches revisited, with emphasis on the amorphous Fe oxyhydroxide phase extraction. J. Geochemical Explor. 56, 59–78. doi:10.1016/0375-6742(95)00050-X Hanesch, M., 2009. Raman spectroscopy of iron oxides and (oxy)hydroxides at low laser power and possible applications in environmental magnetic studies. Geophys. J. Int. 177, 941–948. doi:10.1111/j.1365-246X.2009.04122.x Hannah, J.L., Stein, H.J., Wieser, M.E., de Laeter, J.R., Varner, M.D., 2007. Molybdenum isotope variations in molybdenite: Vapor transport and Rayleigh fractionation of Mo. Geology 35, 703–706. doi:10.1130/G23538A.1 Hayes, J.R., Grosvenor, A.P., Rowson, J., Hughes, K., Frey, R.A., Reid, J., 2014. Analysis of the Mo speciation in the JEB tailings management facility at McClean Lake, Saskatchewan. Environ. Sci. Technol. 48, 4460–4467. doi:10.1021/es404980x Helz, G.R., Bura-Nakić, E., Mikac, N., Ciglenečki, I., 2011. New model for molybdenum behavior in euxinic waters. Chem. Geol. 284, 323–332. doi:10.1016/j.chemgeo.2011.03.012 Helz, G.R., Miller, C. V., Charnock, J.M., Mosselmans, J.F.W., Pattrick, R.A.D., Garner, C.D., Vaughan, D.J., 1996. Mechanism of molybdenum removal from the sea and its concentration in black shales: EXAFS evidence. Geochim. Cosmochim. Acta 60, 3631–3642. doi:10.1016/0016-7037(96)00195-0 Hem, J.D., 1976. Inorganic chemistry of lead in water, in: Lead in the Environment. U.S. Geological Survey Professional Paper 957. Editor: T.G. Lovering. pp. 5–11.     264  Heumann, K.G., Gallus, S.M., Rädlinger, G., Vogl, J., 1998. Precision and accuracy in isotope ratio measurements by plasma source mass spectrometry. J. Anal. At. Spectrom. 13, 1001–1008. Hirsche, D.T., 2012. A field and humidity cell study of metal attenuation in neutral rock drainage from the Antamina mine, Peru. MSc Thesis, University of British Columbia, 152 p. Hirsche, D.T., Blaskovich, R., Mayer, K.U., Beckie, R.D., 2017. A study of Zn and Mo attenuation by waste-rock mixing in neutral mine drainage using mixed-material field barrels and humidity cells. Appl. Geochemistry 84, 114–125. doi:10.1016/j.apgeochem.2017.06.005 Iavazzo, P., Adamo, P., Boni, M., Hillier, S., Zampella, M., 2012. Mineralogy and chemical forms of lead and zinc in abandoned mine wastes and soils: An example from Morocco. J. Geochemical Explor. 113, 56–67. doi:10.1016/j.gexplo.2011.06.001 Jacquat, O., Voegelin, A., Villard, A., Marcus, M.A., Kretzschmar, R., 2008. Formation of Zn-rich phyllosilicate, Zn-layered double hydroxide and hydrozincite in contaminated calcareous soils. Geochim. Cosmochim. Acta 72, 5037–5054. doi:10.1016/j.gca.2008.07.024 Jamieson-Hanes, J.H., Amos, R.T., Blowes, D.W., 2012. Reactive transport modeling of chromium isotope fractionation during Cr(VI) reduction. Environ. Sci. Technol. 46, 13311–6. doi:10.1021/es3046235 Jamieson-Hanes, J.H., Shrimpton, H.K., Veeramani, H., Ptacek, C.J., Lanzirotti, A., Newville, M., Blowes, D.W., 2017. Evaluating zinc isotope fractionation under sulfate reducing     265  conditions using a flow-through cell and in situ XAS analysis. Geochim. Cosmochim. Ac