The Diavik Waste Rock Project: Geochemical and microbiological characterization of low sulfide content large-scale waste rock test piles

Applied Geochemistry 65 (2016) 54e72

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The Diavik Waste Rock Project: Geochemical and microbiological characterization of low sulfide content large-scale waste rock test piles Brenda L. Bailey a, *, David W. Blowes a, Leslie Smith b, David C. Sego c a

Department of Earth and Environmental Sciences, University of Waterloo, Waterloo, ON N2L 3G1, Canada Department of Earth and Ocean Sciences, University of British Columbia, Vancouver, BC V6T 1Z4, Canada c Department of Civil and Environmental Engineering, University of Alberta, Edmonton, AB T6G 2W2, Canada b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 12 May 2015 Received in revised form 22 October 2015 Accepted 23 October 2015 Available online 27 October 2015

Two experimental waste-rock piles (test piles), each 15 m in height
60 m
50 m, were constructed at the Diavik diamond mine in Northern Canada to study the behavior of low-sulfide content waste rock, with a similarly low acid-neutralization potential, in a continuous permafrost region. One test pile with an average of 0.035 wt.% S (<50 mm fraction; referred to as Type I) and a second test pile with an average of 0.053 wt.% S (<50 mm fraction; referred to as Type III) were constructed in 2006. The average carbon content in the <50 mm fraction of waste rock in the Type I test pile was 0.031 wt.% as C and in the Type III test pile was 0.030 wt.% as C. The NP:AP ratio, based on the arithmetic mean of particle-size weighted NP and AP values, for the Type I test pile was 12.2, suggesting this test pile was non-acid generating and for the Type III test pile was 2.2, suggesting an uncertain acid-generating potential. The Type I test pile maintained near-neutral pH for the 4-year duration of the study. Sulfate and dissolved metal concentrations were low, with the exception of Ni, Zn, Cd, and Co in the fourth year following construction. The pore water in the Type III test pile contained higher concentrations of SO2 4 and dissolved metals, with a decrease in pH to <4.7 and an annual depletion of alkalinity. Maximum concentrations of dissolved metals (20 mg L1 Ni, 2.3 mg L1 Cu, 3.7 mg L1 Zn, 35 mg L1 Cd, and 3.8 mg L1 Co) corresponded to decreases in flow rate, which were observed at the end of each field season when the contribution of the total outflow from the central portion of the test pile was greatest. Bacteria were present each year in spite of annual freeze/thaw cycles. The microbial community within the Type I test pile included a population of neutrophilic S-oxidizing bacteria. Each year, changes in the water quality of the Type III test-pile effluent were accompanied by changes in the microbial populations. Populations of acidophilic S-oxidizing bacteria and Fe-oxidizing bacteria became more abundant as the pH decreased and internal test pile temperatures increased. Irrespective of the cold-climate conditions and low S content of the waste rock, the geochemical and microbiological results of this study are consistent with other acid mine drainage studies; indicating that a series of mineral dissolutioneprecipitation reactions controls pH and metal mobility, and transport is controlled by matrix-dominated flow and internal temperatures. © 2015 Elsevier Ltd. All rights reserved.

Keywords: Mine drainage Arctic Waste rock Low-sulfide content

1. Introduction 1.1. Background Waste rock is the non-economic grade of rock excavated to gain access to economic ore deposits. The construction of waste-rock piles exposes sulfide minerals to atmospheric oxygen and water, leading to sulfide-mineral oxidation and the release of acid mine

* Corresponding author. E-mail address: [email protected] (B.L. Bailey). http://dx.doi.org/10.1016/j.apgeochem.2015.10.010 0883-2927/© 2015 Elsevier Ltd. All rights reserved.

drainage (AMD). AMD is characterized by high concentrations of Hþ, SO2 4 , and dissolved metals, such as Zn, Fe, Ni, Co, and Cu. The metal loadings released from waters impacted by AMD are known to cause environmental damage; however, the release of AMD from low S content (<0.1 wt.% S) waste-rock stockpiles with low acidneutralization potential is not well understood. Nor is the influence of cold-climate conditions on these types of systems well defined. Waste rock is composed of a variety of grain sizes, ranging from large boulders to fine-grained sand and silt. Deposition of waste rock in large, steeply sloping piles results in segregation, with a coarse rubble zone at the base of the pile and a larger proportion of

B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

the finer fraction remaining near the top of the pile (MEND, 1995; Chi et al., 2013). Water flow through the waste rock at the Diavik diamond mine site in northern Canada, occurs predominantly as continuum flow through the finer-grained matrix rather than as macropore or by-pass flow (Neuner et al., 2013). Seasonal temperature fluctuations influences the thermal profile within waste rock at the Diavik site (Pham et al., 2013); as the ambient air temperature in the region increased in the spring, the temperature within the waste-rock test piles core also increased. Because the greatest proportion of waste-rock surface area is associated with the fine-grained matrix of the waste rock (Stromberg and Banwart, 1999; Hollings et al., 2001; Chi et al., 2013; Smith et al., 2013b; Pham et al., 2013), understanding the interaction between thermal, hydrological and biogeochemical processes at this scale in waste-rock piles is fundamental to quantifying the extent and duration of environmental impacts of waste-rock stockpiles. One of the objectives of the Diavik Waste Rock Project is to investigate the geochemical characteristics of water from wasterock piles of different scale, with low but differing sulfide content over time, and to examine the microbial evolution of the wasterock test piles in terms of type and abundance of different groups of Fe- and S-oxidizing bacteria in a continuous permafrost region. Bailey et al. (2015) discusses the geochemical characteristics and microbial evolution of waste rock at the 2
2
2 m scale in active zone lysimeters (AZL). This current paper focuses on the (bio) geochemical evolution of low-S content waste rock in large-scale (~70,000 tonnes) waste-rock test piles at Diavik over 4 years from 2007 to 2010. Effluent samples were collected and analyzed to evaluate the metal release rates, extent of acidity in the drainage water, and microbial activity. 1.2. Site description Diavik, an open pit and underground diamond mining operation, is located in a continuous permafrost region, 300 km northeast of Yellowknife, Northwest Territories on a 20 km2 island in Lac de Gras (64 290 N, 110180 W, el. 440 m). The field site has a semiarid climate, receiving an average mean annual precipitation of 280 mm from March 1998 through March 2007, of which 65% occurred as snow (Environment Canada, 2010). Rainfall recorded at the Diavik research site for the duration of this study was 93 mm in 2007, 152 mm in 2008, 73 mm in 2009, and 95 mm in 2010. Over a 5 year period, 2007 through 2011, the mean annual rainfall measured at the test piles was 113 mm. The ambient air temperature at Diavik varies from a maximum of 27  C in July to a minimum of 44  C in January/February. The measured mean annual air temperature from 2007 to 2011 was 8.8  C. The temperature range affects the location, rate and duration of sulphide oxidation within the test piles (Sinclair et al., 2015). The economic ore bodies are diamondiferous kimberlite pipes hosted in granite (75%) and granite pegmatite (14%) containing irregular xenoliths of biotite schist (10%). These host rocks are cut by diabase dikes (<1%). The granites contain trace sulfides and consists mainly of potassium feldspar [KAlSi3O8], albite [NaAlSi3O8], and quartz [SiO2], with <5% each of biotite [KMg3AlSi3O10(OH)2] and muscovite [KAl2AlSi3O10(OH)2] (Jambor, 1997). Mineralogical data has shown that granite host rocks at Diavik have carbonate mineral assemblages present in fractures (unpublished data). The biotite schist (<0.42 wt.% S) contains locally disseminated pyrrhotite with lesser amounts of pyrite, sphalerite, and chalcopyrite (Smith et al., 2013b). The biotite schist aluminosilicate mineral assemblage consists mainly of quartz (20e50%), albite (35e55%), and biotite (10e25%; Jambor, 1997). Diavik segregates waste rock into three types based on sulfide content, Type I (<0.04 wt.% S), Type II (0.04 wt.% S to 0.08 wt.% S) and Type III

55

(>0.08 wt.% S) and will generate a waste-rock stockpile up to 120 Mt over the course of its production (Smith et al., 2013b). The waste rock examined in this study originated from the open pit operations. 1.3. Test pile construction and instrumentation Two large-scale test piles (15 m in height and 60 m
50 m at the base; Fig. 1) were constructed with differing S content, a Type I test pile (average of 0.035 wt.% S in the <50 mm fraction) and a Type III test pile (average of 0.053 wt.% S in the <50 mm fraction; Blowes et al., 2006; Smith et al., 2013a,b,c). Mineralogical and chemical analysis of the Diavik waste rock indicated that carbon in the waste rock is present as carbonate and sulfur is in the sulfide form. Total carbon and total sulfur measurements were used to calculate carbonate and sulfide contents of the <50 mm fraction of the waste rock. These concentrations were subsequently used to estimate the overall CO3-NP and AP values of the bulk mass of waste rock (Table 1; Smith et al., 2013b). Data collected during the construction of the Type I and III test piles shows that biotite schist is well mixed within the waste rock piles (Smith et al., 2013b). A detailed study of the distribution of sulfide and carbonate content versus grain size indicates that both the sulfide and carbonate minerals are concentrated in the finer fractions of the waste rock (Smith et al., 2013b). The average NP:AP ratio (based on the arithmetic mean of particle-size weighted NP and AP values) for the Type III test pile was 2.2 and for the Type I test pile was 12.2. Instruments were placed within and at the base of each test pile. Instruments installed at the base of the test piles included basal collection lysimeters, which provided continuous measurements of the volume of water flow and water samples for geochemical and microbiological characterization, thermistors for monitoring internal temperature, and gas sampling tubes, which provided measurements of the pore-gas composition. In addition, instruments were placed along five tip faces in each test pile. Instrumentation on the tip faces included thermistors, gas sampling tubes and pore-gas permeability probes, time-domain reflectometry sensors for determination of moisture content, access tubes for measurement of the thermal conductivity of the waste rock, and soil-water solution samplers (SWSSs) for collection of pore-water samples. The SWSS were distributed to align vertically above the basal lysimeters. The polyethylene tubing that connected the SWSS to the surface was strung through 50 mm flexible PVC protective conduit that contained heat trace to allow samples to be collected from the late spring to the late fall. Various simulated rainfall event were applied to the Type III test pile in 2007 adding 60 mm of precipitation to the Type III test pile above the natural precipitation in 2007 (not discussed herein). The rainfall in subsequent years was derived entirely from natural events. The Type I test pile was not irrigated and reached field capacity based on natural rainfall events. The artificial infiltration experiments resulted in greater effluent discharge from the Type III test pile than the Type I test pile. In addition, the configuration of the basal drains differs between the two test piles. The basal drain in the Type I test pile runs under the center of the test pile, whereas the Type III test

Table 1 Characteristics of the Type I and Type III waste rock piles (after Smith et al., 2013b).

Total S (wt.%) <50 mm fraction Total C (wt.%) <50 mm fraction mNP:mAP (unitless)

Type III

Type I

0.053 (s ¼ 0.037) 0.030 (s ¼ 0.038) 2.2

0.035 (s ¼ 0.019) 0.031 (s ¼ 0.036) 12.2

mNP ¼ arithmetic mean of particle-size weighted neutralization potential. mAP ¼ arithmetic mean of particle-size weighted acid-generating potential.

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pile contains two basal drains, located on the pile margins. These differences may affect the timing and volume of effluent from the test piles. Full details on the design, construction, and instrumentation of the test-piles are provided by Smith et al. (2013c). 2. Methods of investigation 2.1. Waste rock characterization During the construction of the Type I and Type III test piles, waste-rock samples were collected for particle size and S and C analysis (Smith et al., 2013b). A subset of 37 samples from the <40mm fractions (metric sieve sizes 0.16, 0.315, 0.63, 1.25, 2.5, 5, 10, 14, 20, 28, and 40 mm) were collected for whole-rock analysis for this study. Whole-rock analysis was completed using 10-g samples analyzed as powders that were firmly compressed in cups with a polyester X-ray film using an energy dispersive X-ray fluorescence spectrometer (EDXRF; Minipal4; Panalytical, The Netherlands). The instrument was calibrated against granite, granodiorite and granite gneiss certified reference material (CRM) standards. 2.2. Sample collection and analysis 2.2.1. Field methods Water samples for chemical analysis were collected from the Type I basal drain, Type III north and south basal drains, and Type III basal lysimeters (Fig. 1). Effluent from each test-pile basal drain was collected in a series of flow-through cells and directed to a calibrated tipping-bucket flow gauge to measure the flow rate and volume. Water samples were collected from the first cell in the flow-through cell series as described in Smith et al. (2013a).

Pore-water samples from the interior portions of the test piles were collected using in situ SWSSs (model e-127-1920F1L12B02M2; Hoskin Scientific, Canada), installed on two tip faces (Faces 1 and 2) in the Type I test pile and on three tip faces in the Type III test pile (Faces 1, 2, and 3; described in Smith et al., 2013c). Pore-water samples were extracted using SWSSs set with an initial vacuum of 40e50 kPa. Water was purged from the SWSSs into sealed flow-through cells using N2 gas and samples were retrieved with sterile polyethylene (PE) syringes. Water was stored in prewashed HDPE bottles which were triple-rinsed with the sample water. Over the period between 2007 and 2010, samples were obtained from 13 SWSSs, at depths ranging from 1 m to 9 m below the test pile surface in the Type I test pile and from 18 SWSSs, at depths ranging from 2 m to 9 m below the Type III test pile surface. The frequency of pore-water sample collection was irregular because the SWSS did not consistently yield sufficient sample volumes due to variations in the antecedent moisture conditions and temperature. Moreover, at times, there was insufficient water available for the collection of a full sample volume and limited geochemical analysis was completed on these samples. Field and laboratory procedures for analysis of basal-drain effluent and pore-water samples are described in Bailey et al. (2013, 2015) and Smith et al. (2013a). Analysis included the pH, redox potential relative to the standard hydrogen electrode (Eh), EC, temperature, alkalinity, orthophosphate, Fe(II), H2S, NH3, anion concentrations, and total and dissolved cation concentrations. Field measurements were conducted using Hach spectrophotometer DR/ 8400 for the determination of orthophosphate (o-PO4; Ascorbic Acid method), Fe(II) (Phenanthroline method), S (II) (Methylene Blue method) and NH3 (NH3-N; Salicylate method) concentrations  for select samples (SMEWW, 2005). Concentrations of SO2 4 , NO3 ,

Fig. 1. Ariel view of the Diavik Waste Rock Research Facilities showing locations of the Type I test pile, Type III test pile, Covered pile, and AZLs. Dash line represents the HDPE liner. Inset shows the HDPE barrels used in the construction of the AZLs before rock placement.

B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72  NO 2 , and Cl were determined on unacidified, 0.45-mm-filtered samples by ion-chromatography (IC; DX600, Dionex, USA). Concentrations of total and dissolved cations were determined from unfiltered and filtered water samples acidified with HNO3 to a pH < 2. The determination of minor cations were completed using inductively coupled plasma mass spectrometry (ICP-MS; XSeries 2, Thermo Scientific, USA) and major cations, Ca, K, Mg, and Na, using inductively coupled plasma optical emission spectrometry (ICPOES; iCAP 6000, Thermo Scientific, USA). Quality control and quality assurance were assessed by evaluating several standards covering the range of measured concentrations, by incorporating both field replicates and laboratory blanks into the protocol, and through the calculation of ion balances. Speciation calculations, as described in Bailey et al. (2015), were conducted for samples collected from Type I and Type III north and south drains, and Type I and Type III SWSSs with a complete suite of field and laboratory analyses, and these calculations were undertaken to assist in the identification of the potential phases controlling aqueous concentrations. Unfiltered, unpreserved water samples were collected from the Type I, Type III north and south drains, and from selected basal lysimeters for microbial enumerations of neutrophilic S-oxidizers (Thiobacillus thioparus and related species), acidophilic S-oxidizers (Acidithiobacillus thiooxidans and related species), and acidophilic Fe-oxidizers (Acidithiobacillus ferrooxidans and related species), following the methods described in Bailey et al. (2015). Microbial work was not completed on pore water samples because the mesh size of the Nitex nylon screen (<0.45-mm) surrounding the porous cup on the SWSS (Smith et al., 2013c) was potentially smaller than the size of the microbes.

3. Results and discussion 3.1. Whole-rock analysis The Type I and III test piles were constructed with waste rock with an elemental composition dominated by SiO2 and Al2O3, Table 2 Average particle-size weighted elemental composition of waste rock used to construct the Type I (n ¼ 6) and Type III (n ¼ 31) waste-rock test piles determined by XRF. Element

Type III (wt.%)

Type I (wt.%)

SiO2 Al2O3 Fe2O3 TiO2 MnO MgO CaO Na2O K2O

70.9 15.9 2.5 0.27 0.88 0.03 0.95 3.8 4.8

73.7 15.0 1.4 0.12 0.38 0.02 0.94 3.6 5.0

Type III (ppm)

Type I (ppm)

445 123 169 45 123 9.9 495 41 0.80 30 27 5.5

455 109 153 36 95 4.7 114 27 0.03 12 15 4.0

Ba Rb Sr Pb Cr Cu S Zn As V Ni Co

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followed by K2O, Ba, and Na2O (n ¼ 31 for Type III; n ¼ 6 for Type I; Table 2). The Type I and Type III rock types were dominated by granite and granite pegmatite lithologies, similar to the AZLs (Bailey et al., 2015) and previously characterized country rock at the Diavik site (Jambor, 1997). Pyrite, present in trace amounts, is the principal sulfide mineral associated with the granite and pegmatitic granite rock (Jambor, 1997). Pyrrhotite is the most abundant sulfide mineral (ranging from 0.02 to 0.42 wt.% S) associated with the biotite schist, which also contains lower abundances of sphalerite and chalcopyrite. The Type III material had higher proportions of metals than the Type I material (Table 2). The relative proportions of sulfide minerals can be estimated on the basis of the concentrations of these metals. Pyrrhotite is the principal host mineral for Ni and Co. Electron-microprobe analysis indicates that the composition of the pyrrhotite within the biotite schist component of the waste rock is ((Fe0.852Ni0.004Co0.001)P0.857S1.000; Jambor, 1997), indicating a Ni:Co molar ratio of 4:1. Whole-rock analysis shows an average particlesize weighted Ni to Co molar ratio of 4.4 (n ¼ 31) in the Type III waste rock and 3.4 (n ¼ 6) in the Type I waste rock. The main sources of Cu and Zn are chalcopyrite [CuFeS2] and sphalerite [(Zn,Fe)S], respectively (Bailey et al., 2015; Smith et al., 2013a). The mass of Cu in the Type III waste rock was 9.9 ppm (0.16 mmol kg1) and the mass of Zn in the Type III waste rock was 41 ppm (0.63 mmol kg1). Assuming that the mass of S in the Type III waste rock is present as pyrrhotite, chalcopyrite and sphalerite, the difference between the total molar mass of S in the whole rock samples (392 ppm; 12.3 mmol kg1), and the combined contributions of S from chalcopyrite (0.31 mmol kg1; assuming 2 mol of S for every mole of Cu) and sphalerite (1.25 mmol kg1; assuming 2 mol of S for every mole of [Zn0.95,Fe0.05]), reflects the mass of S present as pyrrhotite (residual S ¼ 10.7 mmol kg1; Table 3). The estimated mass of pyrrhotite accounts for the total mass of both Ni and Co present in the Type III waste rock based on the molar ratios from the pyrrhotite formula of (Fe0.852Ni0.004Co0.001)P0.857S1.000 (Table 3). In addition, the mass of Fe from each sulfide mineral was estimated. The total mass of Fe from sulfide minerals was 9.3 mmol kg1, which represents 6.8% of the Fe in the Type III waste rock. Iron can substitute for Mg in the lattice structure of mica group minerals. Jambor (1997) measured the proportion of FeO, MgO and K2O through electron-microprobe analysis in muscovite grains to be 0.77 wt.%, 0.59 wt.%, and 9.5 wt.%, respectively and biotite grains to be 17 wt.%, 10.6 wt.% and 3.0 wt.%, respectively. The mass of K (500 mmol kg1) in the Type III waste rock was assumed to all be from mica group minerals. Therefore, a large proportion of Mg (total mass of 250 mmol kg1) and Fe (total mass of 145 mmol kg1) was assumed to be from mica group minerals and a small fraction from other trace minerals (Tables 2 and 3). 3.2. Sulfide-mineral oxidation: the release and transport of oxidation products The oxidation of sulfide minerals contained in the granite and þ biotite schist lithologies release SO2 4 , H , dissolved Fe and other metals (Smith et al., 2013a; Bailey et al., 2015). The oxidation of pyrrhotite by O2 can be represented by Equation (1).

Fe17 S20ðSÞ þ

77 þ O þ 3H2 O/17Fe2þ þ 20SO2 4 þ 6H 2 2

(1)

Iron(II) can be further oxidized to Fe3þ and under mildly acidic to near-neutral pH conditions, Fe3þ may precipitate as Fe3þ oxyhydroxides, such as ferrihydrite [nominally 5Fe2O3$9H2O] and goethite [aFeOOH]. Sulfide oxidation also releases Hþ and SO2 4 .

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B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

Table 3 Estimates of sulfide minerals in the waste rock used to construct the Type III waste-rock test pile based on XRF results. Average moles per unit mass of metals (mmol kg1) for chalcopyrite, sphalerite, and pyrrhotite. Mass

Moles in chalcopyrite

Moles in sphalerite

Moles S remaining

Moles in pyrrhotite

10.7

10.7

Residual Ni, Co, Fe

mmol kg1

Total Total Total Total Total Total

S Cu Zn Ni Co Fe

n ¼ 31

CuFeS2

(Zn0.95,Fe0.05)S

15 0.17 0.71 0.52 0.11 150

0.31 0.16

1.3

Fe0.852(Ni0.004Co0.001)S0.857S1.000

0.63

0.16

0.031

Sulfate is also released from sulfide minerals through oxidation during blasting. Bailey et al. (2013) measured the concentrations of blasting residues and SO2 4 in freshly blasted rock from the Diavik open pit and determined that blasting residuals provide a significant source of SO4 in the initial flush of water from the test piles. 3.2.1. Type III test pile 3.2.1.1. Pore water. Sample collection from the SWSSs in the Type III test pile began in 2007. The SWSSs collect pore-water samples from a small area of influence (centimeters to tens of centimeters) surrounding the porous ceramic cup (Smith et al., 2013c). Nichol et al. (2005) suggest pore-water samples represent flux-averaged samples of the pore water based upon the flux generated by the applied suction and the geometry of the ceramic cup, rather than volume-averaged concentrations or flux-averaged samples representative of the entire waste-rock pile. The pore-water samples, therefore, are discrete point measurements from within the matrix material, rather than the integrative samples provided by the basal drains and the basal lysimeters. A limited volume of water was available for sample collection during the wetting up period in 2007 and 2008. Neuner et al. (2013) estimate pore water migrates at a rate of <102 to 3
102 m day1 in response to common rainfall events (1.3 mm h1, <2-year recurrence interval) and up 0.7 m day1 in response to high intensity rainfall events (8 mm h1, 35-year recurrence interval). In 2007, the wetting front had migrated to a depth of 7 m below the crest of the test pile and SWSS samples could be collected at the 3m and 5-m depths. Limited water was available for sample collection from the SWSSs at 7-m and 9-m depths until 2008.  The dissipation of blasting residuals, such as NO 3 -N and Cl , can be used to quantify the first flush of the matrix material (Bailey et al., 2013). The application of the LiCl tracer test on the Type III test pile in September 2007 precludes the use of Cl as a resident tracer. The concentration of SO2 4 in pore water during the first flush of the matrix material is influenced by SO2 4 derived from sulfidemineral oxidation during blasting and from in situ sulfide-mineral oxidation (Bailey et al., 2013). After the first flush of the upper matrix material, the majority of SO2 4 is derived from in situ sulfidemineral oxidation or from the dissolution of secondary mineral assemblages (Bailey et al., 2013). Some areas of the test piles were not flushed in the first few years of the study. As these portions of the test pile wet-up and contribute to the outflow at the base of the test pile, the sources of SO2 4 are likely sulfide-mineral oxidation during blasting, in situ sulfide-mineral oxidation, and the dissolution of secondary mineral assemblages. Nitrate, NO2  -N and NH3-N in the pore water from the Type III test pile, derived from blasting residues, increased in 2007 to the maximum observed concentrations at the 3-m and 5-m depths (Fig. 2 and Supplementary Figs S.1 and S.2). Concentrations of N species decreased by the end of 2008 through 2010 (Fig. 2 and Supplementary Figs. S.1 and S.2). The concentration of SO2 4 at the

0.043 0.011 9.1

0.42 0.093 125

3-m and 5-m depths increased through 2007 from a minimum of 10 mg L1 to a maximum of 2600 mg L1. By the end of 2008, the  concentrations of SO2 4 decreased following a similar trend as NO3 N concentrations, suggesting that blasting residuals were the  principal source of SO2 4 (Bailey et al., 2013). Unlike NO3 -N, the concentration of SO2 4 increased through 2009 and 2010 in near surface pore water (upper 5 m of the test pile) indicating SO2 4 was released from in situ sulfide mineral oxidation (Fig. 2). The pore water at the 7-m and 9-m depths showed increasing SO2 4 and N species concentrations during 2007. These concentrations remained elevated through 2008 (Fig. 2 and Supplementary Figs. S.3 and S.4). By 2009, low concentrations of blasting residuals in the pore water were observed at the 7-m depth, whereas the 9-m depth remained high through 2009 and subsequently decreased in 2010. The SO2 4 concentrations at the 7-m and 9-m depths remained elevated through 2009 and 2010. By 2009, water at the 7-m and 9-m depths showed an increased contribution of SO2 4 from the upper portions of the test pile; therefore there are two main sources of SO2 4 that results in increased concentrations over time including: 1) blasting residuals in the vicinity of the SWSS; and, 2) from in situ sulfide-mineral oxidation. In 2010, the matrix material surrounding the SWSSs at the 9-m depth had been flushed of blasting residuals. These observations suggest that by 2010, constituents derived from blasting activities were displaced beyond the 9-m sample depth and solutes present in the matrix pore water were derived principally from in situ chemical reactions (Figs. S.1eS.4). 3.2.1.2. Basal lysimeters. Flow first reported to one 2 m by 2 m and two 4 m by 4 m basal lysimeters in the Type III test pile in 2008. In 2009, flow was observed in an additional 4 m by 4 m basal lysimeter. No flow has reported to the remaining locations between 2007 and 2010. Two SWSSs, located at 3 m and 7 m, are located directly above one 4 m by 4 m basal lysimeter located in the central portion of the test pile under the crest. This is the only basal lysimeter in the Type III test pile discussed in this paper. Water first reported to the basal lysimeter in the Type III test pile in August 2008 and sample collection commenced at that time. In the Type III basal lysimeter, the concentrations of NO 3 -N, NO 2 -N, and NH3-N decreased through 2008, increased in early 2009, and subsequently decreased through late 2009 (Fig. 3). In 2010, concentrations of nitrogen species were variable, but were within the concentration ranges previously observed at this location. Measurable concentrations of blasting residuals in the basal lysimeters suggest the matrix material above this basal lysimeter had not yet been fully flushed. In 2008, SO2 4 concentrations were moderately-well correlated 2 to NO concentrations 3 -N concentrations, but by 2009, SO4  remained constant while NO3 -N concentrations decreased, suggesting that the contribution of SO2 4 from sulfide-mineral oxidation during blasting was less than that from in situ sulfide-mineral

B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

59

Fig. 2. Pore-water chemistry in the Type III test pile at 3 m, 5 m, 7 m, and 9 m depths from June to November, annually. Face symbols differ for each SWSS depth. Dashed lines represent break in time during frozen periods from December through May.

oxidation (Fig. 3). Geochemical equilibrium modeling suggests the effluent was at or near saturation with respect to gypsum, barite, alunite, and, at times, jarosite, potentially limiting the SO2 4

concentrations (Fig. 4).

concentrations in the 3.2.1.3. Basal drainage. In 2007, the SO2 4

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Fig. 3. Pore-water chemistry samples collected from the center of the pile at 3 m and 7 m depths in the Type III test pile and in the 4 m by 4 m basal lysimeter from June to November, annually. Dashed lines represent break in time during frozen periods from December through May.

Type III basal drain are highly correlated to the concentrations of  2  2 NO 3 and Cl (NO3 :R ¼ 0.87; Cl : R ¼ 0.81), suggesting that the blasting residuals were an important source of SO2 4 (Fig. 5). From 2008 onward the correlation between the concentrations of dis  solved SO2 4 and the concentrations of NO3 , and Cl declined as the proportion of SO2 4 derived from in situ oxidation increased (Bailey et al., 2013).

In 2008, SO2 4 concentrations increased early in the summer through the end of the season. During this phase, the majority of SO2 4 was derived from in situ oxidation. An increase in the concentrations of blasting residuals occurred at the end of the field season suggesting an increase in the proportion of water-flow derived from the central portion of the test pile (relative to batter flow), which had not previously contributed to the test-pile

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61

Fig. 4. Time series plots of calculated saturation indices (SI), using MINTEQA2, for effluent from the basal lysimeter the Type III test pile. The dashed black line at 0 represents equilibrium.

drainage. In 2009 and 2010, SO2 4 and other blasting residuals followed a trend that was similar to that observed in 2008 (Fig. 5). Increases in

 concentrations of NO 3 and Cl suggest that portions of the test pile that had not previously contributed flow to the basal drain became active late in the year. During the period between May and

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Fig. 5. Type I test pile and Type III test pile north and south drain water chemistry (After Bailey et al., 2013). Dashed line represents a break in time during frozen periods.

September, variations in SO2 4 concentrations followed trends in internal test-pile temperatures (Pham et al., 2013), with increasing solute concentrations as the test pile warmed. This pattern suggests that the water first reporting to the basal drain was derived from the batters of the test piles, where the flow

path is shortest. In all subsequent years, the pattern of release of the blasting residuals is consistent with the hypothesis that the water first released to the basal drain systems each year was derived from the margins of the test pile, and flow observed later in the year was derived from a greater proportion of the test pile (Bailey et al.,

B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

2013). The configuration of the test piles affects the timing and origin of flow reporting to the basal drains of the test piles. The thermal regime also profoundly impacts the flow system and transport of sulfide mineral oxidation products within the test piles. Although the interior temperatures throughout the test piles fall below 0  C every winter, the cooling process is not uniform (Pham et al., 2013). In the spring, the test piles thaw from the margins inward, with temperatures rising above 0  C at the outer surfaces of the test pile and a thawing front migrating inward and downward to the center and base. In the fall, this process reverses, with cooling from the surface inward, temperatures initially fall below 0  C at the test pile surface, and the freezing front migrates inward (Pham et al., 2013). As a consequence of the test-pile configuration and the annual thermal cycle, flow derived from the test-pile batters dominates the test-pile effluent in the spring of each year, with increasing contributions from the central portions of the test piles through the summer (Bailey et al., 2013). Late in the year, flow from the center of the test pile becomes more dominant as frost migrates inward from the batters (Bailey et al., 2013). Ahonen and Tuovinen (1992) observed that the rate of pyrrhotite oxidization is dependent on temperature following the Arrhenius equation. The low temperatures prevalent in the spring probably result in lower rates of pyrrhotite oxidation, contributing to lower SO2 concentrations. However, the mass loading of 4 oxidation products, such as SO2 4 and dissolved metals, remained within the waste-rock piles over winter as components of secondary minerals or held within highly concentrated pore waters, and were remobilized later in the evolution of the waste-rock pile, such as in the spring when the test piles thaw. The accumulation of weathering products was observed in waste-rock piles at the Aitik mine in Sweden with 0.7e1.7% S (Stromberg and Banwart, 1999). Similarly, at the Doyon mine in Quebec, Canada, maximum concentrations were observed during dry periods, and decreased concentrations were observed during recharge periods (Sracek et al., 2004). The release and transport of blasting residuals and oxidation products have important implications with respect to the management of waste rock at other mine sites and the ability to predict the quality of mine drainage. The total mass of S released from the Type III test pile (cumulative mass of both the north and south drains) from 2007 to 2010 was 129 kg (Table 4). This S mass corresponds to an estimated oxidation rate of 1.1
1010 kg (O2) m3 s1. Oxidation rates at other sites range from 5
1010 kg (O2) m3 s1 to 5
106 kg (O2) m3 s1, with average rates of 5
108 kg (O2) m3 s1 (Ritchie, 2003). The difference in oxidation rates could be a result of the distribution and mass of sulfides in the waste rock (Ritchie, 2003). 3.2.2. Type I test pile 3.2.2.1. Pore water. Unlike the Type III test pile, the Type I test pile was not irrigated through infiltration experiments or through tracer tests. Neuner et al. (2013) suggest there was no net infiltration into the Type I test pile in 2007. Due to limited infiltration, the moisture content of the waste rock was insufficient to allow sample collection from each SWSS every year, and therefore, sampling was sporadic. There were only a few pore-water samples collected from the Type I test pile in 2007. Pore-water samples were collected in 2008 at an increased number of sample locations, and sampling continued through 2010; however, there was an insufficient volume of water from select pore-water samples to complete all geochemical analyses. Elevated concentrations of blasting residuals were observed in the first few samples collected from pore water at the 1-m depth in 2008 and concentrations subsequently (Fig. 6 and Supplementary Fig. S.5) dissipated from June to October 2008, and remained low

63

Table 4 Calculated mass loadings for parameters of interest from the Type I and Type III test piles (2007e2010). Type III test pile

2007

2008

2009

2010

Total

Ca (kg) K (kg) Mg (kg) Na (kg) Al (g) Fe (g) Co (g) Ni (g) Cu (g) Zn (g) S (kg) Flow (
103 L)

8.5 3.0 9.3 3.9 21 28 19 110 5.5 22 17 49

12 6.0 13 4.5 30 18 46 230 46 64 27 130

13 6.2 21 6.6 140 22 89 460 51 110 19 110

29 11 39 13 300 20 170 1000 61 210 66 210

63 26 82 28 491 88 324 1800 164 406 129 499

Type I test pile

2007

2008

2009

2010

Total

Ca (kg) K (kg) Mg (kg) Na (kg) Al (g) Fe (g) Co (g) Ni (g) Cu (g) Zn (g) S (kg) Flow (
103 L)

0.0084 0.0022 0.003 0.0017 0.043 1.8 nd 0.0065 0.46 0.070 0.026 4.0

6.3 2.3 3.6 2.2 15 8.9 0.6 4.6 20 6.2 3.8 76

0.51 0.19 0.32 0.14 1.5 0.62 0.093 1.0 1.5 0.11 0.46 9.1

28 4.4 12 6.7 9.6 5.0 2.5 17 0.74 1.7 7.6 44

35 6.9 16 9.0 26 16.3 3.2 23 23 8.1 11.9 133

for the remainder of the study at the 1-m depth. The exception to this trend was observed in one sample collected in 2010 that had the maximum NO 3 -N concentrations observed at the test piles 1 research site (NO 3 -N ¼ 21,000 mg L ). Sulphate concentrations followed a similar trend to blasting residuals in 2008, however, in subsequent years SO2 4 concentrations increased with decreasing pH suggesting SO2 was derived from in situ sulfide mineral 4 oxidation after 2007. Concentrations of blasting residuals in pore water at the 5-m depth were elevated in 2008 and 2009 (Fig. 6 and Supplementary Fig. S.6), and only one samples was collected in 2010 due to low sample volumes. The concentrations of blasting residuals in the 7m depth progressively increased from 2007 through 2009 and 2010, with maximum concentrations observed in 2010 (Fig. 6 and Supplementary Fig. S.7). Sulfate concentrations were not correlated to N species and the pH in the Face 2 West SWSS decreased below 5, therefore, SO2 4 was likely derived from a combination of sulfide mineral oxidation during blasting and in situ. Limited samples were collected at the 9-m depth, however, in the few samples collect, elevated blasting residuals were measure through 2009 and 2010 (Fig. 6 and Supplementary Fig. S.5) with the maximum observed 1 concentrations of Cl and NO 3 in all the test piles at 2800 mg L 1 and 28,000 mg L , respectively. The pore-water SO2 4 and metal concentrations observed at the top of the test pile at the 1-m depth were lower than at the 7-m depth (Fig. 6 and Supplementary Figs. S.5eS.7) and both depths observed similar minimum pH at 4.7 and 4.6, respectively. Both of these locations had a high frequency of success in sampling pore water and can be assumed to represent the migration of water from the top of the test pile to the 7-m depth in the test pile. The maximum concentration of SO2 4 measured in all Type I test pile pore water samples was 2700 mg L1 at the 7-m depth. Neuner et al. (2013) suggest that drainage through the finergrained matrix dominates the flow mechanisms, rather than macropore or by-pass flow. The progressive increase in concentrations at the base of the test pile was consistent with the migration of water through the matrix material within the Type I

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B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

Type I 1 m and 2 m

Type I 5 m

9

Type I 9 m

Type I 7 m

pH

8 7 6 4 50 40 30 20 10 0 30000 25000 20000 10000 7500 5000 2500 0 3000 2500 2000

2-

-1

SO4 (mg L )

-

-1

NO3 -N (mg L )

-1

Alk. (mg L as CaCO3)

5

1500 1000 500 0

-1

Ca (mg L )

3000 2000

-1

Fe (mg L )

1000 0 3.0 2.5 2.0 1.5 1.0 0.5

8

-1

Zn ( m g L )

-1

Ni ( m g L )

0.0 10 6 4 2 0 14 12 10 8 6 4 2 0

Face 1 East Face 1 West Face 2 East Face 2 West Face 2 West

2007

2008

2009

2010

2007

2008

2009

2010

2007

2008

2009

2010

2007

2008

2009

2010

Fig. 6. Pore-water chemistry in the Type I test pile at 1 m, 2 m, 5 m, 7 m, and 9 m depths from June to November, annually. Dashed lines represent break in time during frozen periods from December through May.

test pile. The effluent chemistry at the base of the test pile is consistent with matrix-dominated flow, rather than macropore or by-pass flow.

3.2.2.2. Basal drain. Due to a defect in the basal drain of the Type I test pile only one water sample was collected in 2007. In 2008 and 2009, the effluent from the Type I test pile contained lower

B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72 1 concentrations of SO2 4 (10e410 mg L ) than those observed in the Type III test-pile effluent (Fig. 5). Sulfate concentrations increased as the flow rate decreased and ceased in the fall of each year. In 2010, SO2 concentrations increased from July to November, 4 reaching a maximum of 1800 mg L1, as a larger portion of the test pile contributed to the outflow, releasing blasting residuals. The maximum SO2 4 concentration correlates to maximum concentrations of other blasting residuals (Fig. 5), suggesting limited contribution of in situ sulfide oxidation. The total mass of S released from the Type I test pile from 2007 to 2010 was 12 kg corresponding to an estimated global oxidation rate of 4.0
1011 kg (O2) m3 s1. The global oxidation rate of the Type I material is slightly lower than that observed in the Type III material, but it is greater than that observed in the AZL tests (Bailey et al., 2015).

3.3. Microbial populations Sulfide-mineral oxidation is catalyzed by bacteria, including Thiobacillus species and Acidithiobacillus species. Three groups of Fe- and S-oxidizing bacteria (Acidithiobacillus ferrooxidans and related species, Acidithiobacllus thiooxidans and related species and Thiobacillus thioparus and related species) were monitored in effluent from the Type I and Type III waste-rock test piles and enumerated (Fig. 7). Neutrophilic S-oxidizing bacteria (nSOB) were observed in both Type III basal drains in 2008. The number of nSOB decreased in late 2009 as the pH decreased (Fig. 7). Populations of nSOB were also

8

Type III test pile north

106

7 6

102

5

100

4 8

Type III test pile south

106

7

104

6

102

5

100

4 8

Type I test pile

106

pH

MPN (bacteria mL-1)

104

7

104

6

2008

2009

Nov

Aug

Nov May

Aug

4

Nov May

100

Aug

5

May

102

2010

pH

nSOB

aFeOB

aSOB

Fig. 7. Populations of aFeOB, nSOB, and aSOB in the Type I test-pile drain and the Type III test-pile north and south drains. Dashed line represents a break in time during frozen periods.

65

present in the Type I basal drain effluent, with populations ranging from 9.6
102 bacteria mL1 to 1.1
106 bacteria mL1 from 2008 through 2010. The pH in the Type I test pile remained near neutral for the duration of the study e ideal for neutrophilic S-oxidizing bacteria. In 2008, acidophilic S-oxidizing bacteria (aSOB) were not detected in the Type III test pile; aSOB were detected in 2009, but not in 2010. Iron-oxidizing bacteria (aFeOB) were observed in the Type I basal drain in June 2009 and in 2010. In the Type III basal drains the populations of aFeOB increased with decreasing pH. A comparison of the population numbers to pH shows the abundance of aFeOB is highest when pH was below 6 and nSOB show higher populations when the pH was above 5.5 (Fig. 7). Sulfur- and Fe-oxidizing bacteria play an important role in the biogeochemical evolution of waste-rock test piles, accelerating the rate of sulfide oxidation and increasing the release rate of sulfate and other typical AMD constituents. These results are consistent with studies conducted on the upper 2 m of waste rock at Diavik (Bailey et al., 2015) and on mine tailings (e.g., Blowes et al., 1995; Benner et al., 2000); which documented low populations of Feand S-oxidizing bacteria where sulfide-mineral oxidation was limited, and that higher population numbers where sulfide-mineral oxidation was occurring and acidic conditions were prevalent. 3.4. Geochemistry of waste-rock test-pile pore water and effluent 3.4.1. Type III test pile 3.4.1.1. pH, alkalinity and acid neutralization. The pH of the Type III test-pile effluent decreased from 7.5 to 5.0 between May and September 2007 and the pH remained at 5.0 until the end of October (Fig. 8). The initial near-neutral pH and the presence of detectable alkalinity suggest the Hþ generated by sulfide oxidation was consumed by the dissolution of carbonate minerals and subsequently by the dissolution of silicate minerals. Average carbon content measurements of Type III waste rock, weighted based on particle size, indicate decreasing carbon content with increasing particle size (Smith et al., 2013b). The carbon content of the <50 mm fraction of Type III waste rock was 0.030 wt.% (s ¼ 0.038). Smith et al. (2013b) estimate the ratio of neutralization potential to acid generating potential (NP:AP) of the Type III waste rock to be 2.2 (Table 1), suggesting an uncertain acid-generating potential. Alkalinity concentrations were highest (35 mg L1 as CaCO3) in early 2007 and decreased with decreasing pH to ~ 5 at the end of the field season (Fig. 5). Similar trends in pH and alkalinity were observed in 2008, 2009 and 2010; however, the pH decreased earlier each year. In 2008 and 2009, the alkalinity decreased from 40 mg L1 as CaCO3 to below the method detection limit (<0.5 mg L1). In 2010, the alkalinity was lower, decreasing from 20 mg L1 as CaCO3 in May to below detection (Fig. 5). Geochemical equilibrium modeling indicates the effluent is undersaturated with respect to all carbonate minerals (Fig. 9). Acid generation in the Type III test pile exceeded the neutralizing capacity available from carbonate minerals at the end of each field season. The pH decrease from near neutral to a secondary pH plateau of 5.0 corresponded to an increase in Al concentrations from 2007 to the maximum concentrations in 2010 (Fig. 10). Dissolved Al is derived from the dissolution of aluminosilicate and aluminum hydroxide minerals. A previous mineralogical study was conducted on weathered samples obtained from humidity-cell experiments conducted on the Diavik waste rock (Blowes and Logsdon, 1998). Samples of biotite schist obtained from humidity-cell experiments that were acidic showed alteration of biotite grains, with depletion of K and Mg and enrichment of Al and Si on the edges of biotite laths. Early each year, the Type III test-pile effluent approached equilibrium with respect to amorphous Al(OH)3, and was

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B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

Fig. 8. Concentrations of dissolved metals in Type I test-pile drain and Type III test-pile north and south drains. Dashed line represents a break in time during frozen periods.

supersaturated with respect to gibbsite and boehmite when the pH was above 5.0 and alkalinity concentrations were above detection (Fig. 9). Near the end of the field season, as the pH falls below 5.0, effluent was undersaturated with respect to these minerals.

Geochemical speciation modeling indicates that effluent from the Type III test pile was at equilibrium with respect to boehmite and basaluminite in 2007, and was supersaturated with respect to both of these minerals for the remainder of the period between 2008

B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

67

Fig. 9. Time series plots of calculated saturation indices (SIs), using MINTEQA2 for the Type I test-pile drain and Type III test-pile north and south drains. The solid black line at 0 represents equilibrium.

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B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

Fig. 10. Major cation concentrations in the Type I test-pile drain and Type III test-pile north and south drains. Dashed line represents a break in time during frozen periods.

and 2010, with the exception of a few select samples (Fig. 9). Dissolved Fe concentrations were less than 1.0 mg L1 throughout the period between 2007 and 2010 (Fig. 8). The Type III test pile effluent was supersaturated with respect to ferrihydrite at the beginning of each year, and became undersaturated with

respect to ferrihydrite at the end of each year as the pH decreased below 5.0. The effluent remained supersaturated with respect to goethite over the entire period between 2007 and 2010. The effluent was occasionally supersaturated with respect to jarosite between 2007 and 2009, and approached saturation with respect to

B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

jarosite in 2010. Mineralogical studies of Type III column charges from humidity tests indicate the presence of patchy, ochreous coatings of jarosite [KFe3(SO4)2(OH)6] (Langman et al., 2014). Mineralogical characterization of the secondary minerals from the test piles indicates that similar phases are present (Langman et al., 2015). These observations suggest that formation of secondary Fe(III) (oxy)hydroxides and hydrosulfates, control the dissolved Fe concentrations from the Type III test pile (Fig. 9). 3.4.1.2. Major cations. The dissolution of carbonate and aluminosilicate minerals releases major ions (Ca, Mg, K, Na and Mn), and can play an important role in the formation of secondary minerals, such as gypsum and jarosite (Blowes and Jambor, 1990). Some K may be released from KClO4 that is used during blasting but the concentrations are minor (Bailey et al., 2013). Calcium, Mg, K and Na concentrations in the Type III pore water were elevated in 2007 at the 3-m and 5-m depths as water from the top of the test pile migrated downward (Fig. 2 and Supplementary Figs. S.1 and S.2). Concentrations subsequently decreased in 2008, corresponding with the decrease in pH, and further decreased each subsequent year, with the exception of slight increases in concentrations at the 5-m depth. Few pore-water samples were collected in the 7-m and 9-m depths in 2007, as the wetting front remained at 7 m by the end of 2007 (Fig. 2 and Supplementary Figs. S.3 and S.4). Concentrations of Ca, Mg, K, and Na were elevated at 7 m and 9 m in 2008, indicating water at the 3-m and 5-m depths from 2007 had migrated downward (Fig. 2 and Supplementary Figs. S.3 and S.4). Concentrations subsequently decreased similar to the trend observed in the upper portions of the test pile, with the exception of a few samples that were elevated in 2009 and 2010. The concentrations of Ca and Mg in the basal lysimeter were elevated at the beginning of each field season, when the pH was near neutral, and subsequently decreased through to the end of the field season with decreasing pH and alkalinity (Fig. 3). Maximum concentrations of Ca, Mg, K, and Na were observed in 2010, suggesting that water from the top of the test pile reached the basal lysimeter through matrix flow. The concentrations of Ca, Mg, K, and Na in the pore water and the basal lysimeters were similar to, or slightly lower than those of the Type III test-pile basal drains, suggesting that solute transport in the waste-rock test pile was matrix dominated. Geochemical equilibrium modeling suggests the pore water was supersaturated with respect to alunite for the duration of the study and jarosite at various times (Supplementary Figs. S.9eS.12). The precipitation of these minerals may provide a limit to dissolved K concentrations. Geochemical equilibrium modeling suggests Ca and SO2 4 concentrations in the pore water may have been limited by the precipitation of gypsum and barite for the duration of the study (Supplementary Figs. S.9eS.12). Calcium concentrations in the Type III north and south drains increased with increasing internal core temperature and decreasing pH. In 2009 and 2010, Ca concentrations were low in May, gradually increased until August and then sharply increased in September through October. The maximum Ca concentrations were observed at the end of 2009 and 2010 (Fig. 10). Geochemical speciation modeling suggests effluent from the Type III test pile was undersaturated with respect to gypsum, with the exception of the last few samples collected at the end of each field season, when the flow rate significantly decreased (Fig. 9). Magnesium and K are likely released by the kinetically limited dissolution of biotite (Bailey et al., 2015). The cumulative ratio of Mg:K concentrations in the Type III basal drain was 3.1, with an average of 2.85, and the ratios for individual samples range from 0.21 to 5.35. The ideal molar ratio of Mg:K in biotite [KMg3AlSi3O10(OH)2] is 3:1. The range of ratios in the Type III basal drain

69

suggests there was another source of Mg, or that K was preferentially removed from the effluent. Geochemical modeling indicates that the effluent from the Type III basal drains was at saturation with respect to jarosite at the beginning of 2008 and 2009, thus the precipitation of jarosite may provide a limit to dissolved K concentrations (Fig. 9). Dissolved Na and Mn are derived from weathering of aluminosilicate minerals. In 2010, the concentration of Na followed a similar trend to Mg and Ca, and significantly increased at the end of the field season (Fig. 10). 3.4.1.3. Metals. Increases in Fe, SO2 4 , and most dissolved metal concentrations coincided with the decrease in pH from the top of the test pile through to the base of the test pile (Fig. 8). Concentrations of dissolved metals were low at all depths in 2007, with the exception of one SWSS at 5 m that had Ni concentrations ranging from 13 mg L1 to 20 mg L1 and Co concentrations ranging from 4 mg L1 to 6 mg L1; however, the pore water at 3 m and 5 m had slightly higher concentrations than the 7-m and 9-m depths (Fig. 2 and Supplementary Figs. S.1eS.4). The metal concentrations at the 3-m and 5-m depth increased slightly in 2008 (Supplementary Figs. S.1 and S.2). The concentrations at the 3-m and 5-m depths further increased through 2010 with high concentrations of SO2 4 (4500 mg L1), Al (79 mg L1), Ni (25 mg L1), Cu (6.6 mg L1), Zn (19 mg L1), and Co (5.5 mg L1) in near-surface pore water (Fig. 2). The concentrations observed in the Type III test-pile pore water were lower than those observed in the effluent from the Type III 2 m by 2 m scale active zone experiments (Bailey et al., 2015). The pore water at the 9-m depth did not exhibit the same decrease in pH or increases in SO2 4 and dissolved metals observed in upper portions of the Type III test pile (Fig. 2 and Supplementary Fig. S.4). Because of the variability in the Type III waste rock and the small zone of influence of the SWSSs, it is possible that the waste rock in the area of the 9-m SWSS may have had a lower S content or higher carbonate content than other regions within the test pile. Effluent reporting to the base of the test piles in the basal lysimeters and the basal drains exhibited decreases in pH, increases in SO2 4 and in most dissolved metals similar to those observed for pore-water samples at 3 m, 5 m, and 7 m. The effluent and porewater concentrations of Al, Ni, Co, Zn, Cd, and Cu increased, corresponding to decreases in pH (Figs. 3 and 8). In 2007, dissolved metal concentrations at the 3-m and 7-m depths were low (Fig. 3). Dissolved metal concentrations increased with decreasing pH from 2008 through 2010 at both depths, with higher concentrations observed at the 3-m depth (Supplementary Fig. S.1) than at the 7-m depth (Supplementary Fig. S.3). The maximum pore-water concentrations at the 3-m depths and 7 m were reported in 2010. This observation is consistent with the hypothesis that solute transport is through matrix flow, and that concentrations deeper in the test pile are likely to increase over time, similar to the upper zone of the pile. The basal-lysimeter effluent remained above pH 5.0 and concentrations of dissolved metals remained low, compared to the 3-m and 7-m depths, and only slightly increased over time (Fig. 3). The concentrations of dissolved metals in the basal-lysimeter effluent increased with decreasing pH, and maximum concentrations of Al, Ni, Co, Cu, Zn, and Cd were observed when the pH decreased below 5.0 in 2010 (Fig. 3). The maximum concentration of Fe was observed in 2009. Geochemical equilibrium modeling suggests the basallysimeter effluent was supersaturated with respect to Fe oxyhydroxides (goethite, lepidocrocite, and, at times, ferrihydrite), suggesting that the accumulation of these secondary minerals may have limited the Fe concentrations (Fig. 4). The maximum concentrations of dissolved metals in the basal lysimeter are expected to increase over time as the water moves from the top of the test pile through to the base, and with a decrease in effluent pH.

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B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

The concentrations of dissolved metals in the north and south Type III basal drains differed in 2007. The north-drain effluent showed a trend of increasing metal concentrations from May through September when flow ceased. The dissolved metal concentrations in south-drain effluent also increased, but reached lower maximum concentrations. The volume of flow was larger for the south basal drain and flow continued until late October. In subsequent sampling seasons, the concentrations and trends were similar for both drains; dissolved metal concentrations increased as the temperature within the test pile increased and more extensive portions of the test pile contributed to the outflow from the basal drains. Maximum concentrations were observed in 2009. The oxidation of pyrrhotite releases dissolved Ni and Co in the test-pile effluent. Trace amounts of sphalerite and chalcopyrite have been observed in the Diavik waste rock, providing potential sources of dissolved Zn and Cu, respectively. Cadmium, Cu, and Mn can be found at trace concentrations in sphalerite. The concentrations of dissolved metals in the Type III basal drains followed consistent seasonal trends in 2009 and 2010, with increased concentrations correlated to increased internal core temperature, and decreases in pH and alkalinity at the end of the field season (Fig. 5). However, the maximum concentrations were observed in 2009, and an increase in the total flow from the Type III basal drains in 2010 (2.10
105 L) relative to 2009 (1.10
105 L), resulted in lower total mass loadings in 2009, approximately half of the 2010 total mass loadings (Table 4). The concentrations of Ni were highly correlated to the concentrations of Co in the Type III south drain (R2 ¼ 0.98) and Type III north drain (R2 ¼ 0.96). The ratio of Ni to Co concentrations in the Type III test-pile effluent averaged 4.9. This value is only slightly higher than the molar ratios observed in the whole-rock analysis (4.63) and the electron microprobe analysis of pyrrhotite grains [(Fe0.852Ni0.004Co0.001)P0.857S1.000] (Jambor, 1997), suggesting pyrrhotite oxidation is the primary source of Ni and Co. Geochemical equilibrium modeling suggests the effluent from the Type III test-pile pore water and basal-drain effluent was undersaturated with respect to all secondary Ni, Zn, Cd, and Cu minerals included in the WATEQ4F database, suggesting that Ni, Zn, Cd, and Cu concentrations were not limited by precipitation of secondary minerals (Figs. 4 and 9, and Supplementary Fig. S.9eS.12). Metals such as Ni, Zn, Cd, Cu and Co, however, have been observed to coprecipitate with, and adsorb to Fe(III)oxyhydroxides at near-neutral pH conditions, and the increase in metal concentrations corresponded to the decrease in pH (Dzombak and Morel, 1990; Ribet et al., 1995; McGregor and Blowes, 2002; Carlsson et al., 2002; Dold and Fontbote, 2002; Moncur et al., 2005; Gunsinger et al., 2006). Dissolved metal concentrations at all sampling locations (pore water and effluent) in the Type III test pile increased with decreasing pH below 5.0 and the maximum dissolved metal concentrations were observed at the lowest pH (pH < 4), suggesting dissolved metal concentrations in the Type III test pile were controlled by coprecipitation or adsorption reactions. The SO2 4 and dissolved metal concentrations observed in the Type III test pile were much lower than well-known AMD-impacted sites, such as pore waters in the Camp tailings impoundment at 1 1 Sherridon, Manitoba (maximum SO2 4 ¼ 280 g L , Cu ¼ 1.6 g L , 1 and Zn ¼ 55 g L ; Moncur et al., 2005) and underground workings in the Richmond mine at Iron Mountain, California (maximum 1 1 1 SO2 4 ¼ 760 g L , Cu ¼ 4.76 g L , and Zn ¼ 23.5 g L ; Nordstrom, 2000). However, the metal concentrations at the base of the Type III I test pile are high compared to Canadian water-quality guidelines for the protection of aquatic life (CEQG, 1999). This suggests that, if not managed appropriately, the waste rock at Diavik has the potential to have an impact on the local environment in terms of

metal loadings, in addition to low pH and high SO2 4 . In choosing a site-specific remediation plan, the contributions from metal loadings must be considered rather than concentrations alone, because the highest concentrations do not always have the greatest impact (Kimball and Runkel, 2009). 3.4.2. Type I test pile 3.4.2.1. pH, alkalinity and neutralization. Flow from the base of the Type I test pile began in 2008. The pH of the effluent ranged from circum-neutral pH (pH > 7) at in the spring of each year to approximately pH 5 later in the year, for the duration of this study (Fig. 5). In early 2008 alkalinity increased from 30 mg L1 as CaCO3 in May to 60 mg L1 as CaCO3 in June, and decreased through the remainder of the field season to <5 mg L1 as CaCO3. Indicating that carbonate mineral dissolution is a principal acid consuming mechanism in the Type I test pile. The alkalinity concentration returned to 50 mg L1 as CaCO3 when flow commenced in 2009 and decreased over the field season to <8 mg L1 as CaCO3. Alkalinity concentrations were elevated in two samples collected in early 2010, but the drain lines were frozen from late May through early July. After flow re-commenced in late July, concentrations had decreased to 15 mg L1 as CaCO3, following a similar trend to 2009, ending the field season at 5 mg L1 as CaCO3. Geochemical equilibrium modeling suggests the Type I test pile effluent was at equilibrium with respect to calcite for a short period at the beginning of 2008, then fell below saturation from 2008 through to 2010 (Fig. 9). The neutral pH of the effluent suggests that the limited neutralization potential of the Type I rock (0.031 wt.% C, s ¼ 0.036; Smith et al., 2013b) was sufficient to consume the Hþ generated by the oxidation of the modest sulfide content of the waste rock. The effluent from the Type I test pile remained undersaturated with respect to siderite suggesting no potential for the formation of secondary siderite. The Type I test pile effluent contained very low concentrations of Al, with maximum concentrations of <1.0 mg L1. From 2007 to 2010 the Type I test pile released a total mass of 26 g Al. Aluminosilicate dissolution provides limited acid neutralization at near neutral pH (Jambor et al., 2007). The low concentrations of Al in the effluent suggests that Al was released via aluminosilicate dissolution and that precipitation of secondary Al-bearing minerals may have limited the concentrations of dissolved Al. The effluent is near saturation or undersaturated with respect to amorphous Al(OH)3 and consistently supersaturated with respect to the more crystalline Al oxides including gibbsite and boehmite (Fig. 9). 3.4.2.2. Major cations. The concentrations of Ca, Mg, K, and Na in the Type I test-pile pore water were variable over time and with depth (Fig. 6 and Supplementary Fig. S.5eS.8). The average concentrations of Ca (590 mg L1), Mg (450 mg L1), K (180 mg L1), and Na (400 mg L1) observed in the Type I test-pile SWSSs from 2007 to 2010 were within the range of concentrations observed in the Type I basal drain (Ca ranged between 11 and 2100 mg L1; K ranged between 2.8 and 230 mg L1; Mg ranged between 4.0 and 1200 mg L1; Na ranged between 1.1 and 640 mg L1). However, the maximum concentration of K (730 mg L1) and Na (2300 mg L1) in the pore water was significantly higher than in the Type I basal drain. The concentrations of Ca and Mg were not significantly different from the range of concentrations observed in the pore water (P ¼ 0.351 for Ca and P ¼ 0.210 for Mg; ManneWhitney rank sum test; SigmaPlot, Systat Software Inc.). Geochemical equilibrium modeling suggests the Type I test-pile pore water was supersaturated with respect to alunite for the duration of the study, with the exception of the 2-m SWSS, and individual samples from the 5-m, 7-m, and 9-m depths (Supplementary Figs. S.13eS.17). The calculated saturation indices

B.L. Bailey et al. / Applied Geochemistry 65 (2016) 54e72

for jarosite were variable at all depths over time, ranging from above to below saturation. Pore water was also below saturation with respect to gypsum for the duration of the study (Supplementary Figs. S.13eS.17). The precipitation of secondary minerals, such as alunite and jarosite, may limit to SO2 4 and/or dissolved K concentrations. The concentrations of Ca, Mg, K and Na in the Type I basal drain effluent generally followed a similar seasonal trend to the Type III basal drain. However, the total mass loadings of Ca, Mg, K and Na from the Type I drain were lower than in the Type III test pile (Table 4). Type I basal drain followed a similar trend to the Type III south drain in 2007 and reached a maximum Ca concentration of 320 mg L1 (Fig. 10). In 2008 and 2009, concentrations also followed a similar seasonal trend to the Type III south drain, but with lower concentrations. In 2010, the Ca concentrations were significantly higher and increased to 2100 mg L1. The alkalinity of the Type I test pile effluent remained >5 mg L1 for the duration of the study and the effluent from the Type I basal drain was undersaturated with respect to gypsum throughout the period from 2007 to 2010. The Type I basal drain had lower Mg and K concentrations than the Type III basal drains through 2008 and 2009, but in 2010 concentrations increased similar to the maximum concentrations observed in the Type III basal drain. The molar ratio of Mg:K in the cumulative Type I basal drain effluent was 2.3, the average Mg:K molar ratio was 1.79, and individual samples ranged from 0.23 to 3.42. These values are slightly lower than the ideal 3:1 molar ratio of Mg:K in biotite. Type I test pile effluent was undersaturated with respect to K-bearing minerals, suggesting that the lower molar ratio may have been a result of an additional source of K, or the molar concentration of Mg in biotite was not ideal (3:1), or Mg was preferentially removed from the effluent. The Type I basal-drain effluent had low concentrations of Na through 2008 and 2009, and concentrations steadily increased in 2010 to 20 mg L1.

3.4.2.3. Dissolved metals. The concentrations of dissolved metals in the Type I SWSSs were highly variable, with maximum concentrations that were greater than the average concentrations observed in the Type III test-pile basal drain (Fig. 6 and Supplementary Figs. S.5eS.8). The average concentrations of dissolved metals in the Type I pore-water samples were 0.35 mg L1 Fe, 1.5 mg L1 Ni, 0.50 mg L1 Co, 0.50 mg L1 Cu, 1.4 mg L1 Zn, and 13 mg L1 Cd. Concentrations of dissolved Ni, Zn, Cu and Cd in the Type I basaldrain effluent were lower than in the Type III basal drains, with maximum concentrations of Fe ¼ 0.99 mg L1, Ni ¼ 1.1 mg L1, Co ¼ 0.13 mg L1, Cu ¼ 1.3 mg L1, Zn ¼ 0.55 mg L1, and Cd ¼ 16 mg L1. The concentrations of Ni were highly correlated to the concentrations of Co in the Type I basal drain (R2 ¼ 0.98), similar to the Type III test pile and the ratios observed in pyrrhotite (Jambor, 1997). The Type I test-pile pore water and effluent was undersaturated with respect to Ni-, Cu-, Cd-, and Zn-bearing phases included in the WATEQ4F database (Fig. 9 and Supplementary Figs. S.13eS.17). No secondary Ni-, Cu-, Cd-, or Zn-bearing minerals were identified in mineralogical studies of waste rock from the Diavik site (Jambor, 1997). Pore water and effluent were supersaturated or near equilibrium with respect to ferrihydrite and goethite, suggesting that the formation of secondary Fe(III) (oxy) hydroxides and hydrosulfates may control the dissolved Fe concentrations (Fig. 9 and Supplementary Figs. S.13eS.17). Maximum metal concentrations were observed at minimum pH, suggesting that metal concentrations were probably controlled by adsorption reactions, and that metals were released through acidneutralization reactions.

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4. Conclusions The effluent and pore water from the Type I waste-rock test-pile (0.035 wt.% S) and the Type III waste-rock test pile (0.053 wt.% S) were monitored from their initial construction and the subsequent four years after waste rock placement. In spite of the low sulfide content of both test piles, significant and distinguishable impacts on water quality were observed. A modest difference in sulfide content between the Type I rock and the Type III rock resulted in substantial differences in effluent water quality. Biotite schist, present in low quantities (<10% of waste rock at Diavik), contributes most of the sulfide minerals in the Type III waste rock test pile. The low carbonate content and kinetically limited acid neutralization potential from aluminosilicate mineral assemblages results in net acidity and metal leaching in the Type III test pile effluent. Due to the low sulfide content of the Type I rock, acid generation by sulfide-mineral oxidation was balanced by acid neutralization through carbonate mineral dissolution (NP:AP ¼ 12.2) and the effluent of the Type I test pile remained near-neutral. The con1 centrations of SO2 4 (<500 mg L ) and dissolved metals remained low (average concentrations of 0.12 mg L1 Fe, 0.16 mg L1 Ni, 0.14 mg L1 Cu, 0.062 mg L1 Zn, 2.2 mg L1 Cd, and 0.026 mg L1 Co). In contrast, the modestly higher sulfide content of the Type III test pile resulted in a lower NP:AP ratio (2.2), lower effluent pH (<4.5), and higher concentrations of SO2 (maximum of 4 2800 mg L1) and dissolved metals (maximum of 20 mg L1 Ni, 2.3 mg L1 Cu, 3.7 mg L1 Zn, 35 mg L1 Cd, and 3.8 mg L1 Co). The average alkalinity of the Type I effluent remained above 16 mg L1 throughout the duration of the study, indicating ongoing dissolution of carbonate minerals, whereas the alkalinity of the Type III effluent decreased to below the detection limit (0.05 mg L1) each year. The low S and C content of the Type III waste rock test piles suggested there was a low potential for AMD from this material; however, the acid-neutralization sequence observed in the Type III waste-rock test pile was similar to other AMD studies. A series of mineral dissolutioneprecipitation reactions controlled pH and metal mobility; carbonate mineral dissolution buffered the acidity generated from sulfide-mineral oxidation at near neutral pH and the dissolution of Al and Fe (oxy)hydroxides buffered the effluent and pore water at pH < 5.0. Pore-water concentrations were elevated compared to the effluent samples in both the Type I and Type III test piles The concentrations of blasting residuals were high in the early time samples in the upper portion of each the test pile. As pore water migrated downwards blasting residuals were flushed and concentrations increased in the deeper SWSS and in the test pile basal drains. The geochemical patterns documented in this study corroborate the findings by Neuner et al. (2013) that conclude flow is matrix dominated, rather than flow through the macropores or by-pass flow. The pronounced variations in the internal temperatures of the waste-rock test piles affected the concentrations of SO2 4 and dissolved metals in effluent by altering the locations within the tests piles contributing flow to the basal drainage system, and by affecting the rate of sulfide mineral oxidation. The waste-rock test piles internal temperatures decreased below 0  C each year and flow ceased; however, sulfide mineral oxidation may have continued below freezing temperatures. The results from this study showed near neutral effluent in the Type I test pile was associated with a population of neutrophilic Soxidizing bacteria, whereas, the evolving effluent in the Type III test pile was accompanied by changes in the microbial populations. Populations of acidophilic S-oxidizing bacteria and Fe-oxidizing bacteria were present each year.

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Acknowledgments Funding for this research was provided by a Natural Sciences and Engineering Research Council of Canada (NSERC) Collaborative Research and Development Grant (CRDPJ 401328 - 10); Diavik Diamond Mines Inc.; a Canadian Foundation for Innovation (CFI) Innovation Fund Award; the Mine Environment Neutral Drainage (MEND) Program, the International Network for Acid Prevention (INAP), and an Ontario Research Fund e Research Excellence Award (#RE05 e 043). Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.apgeochem.2015.10.010. References Ahonen, L., Tuovinen, O.H., 1992. Bacterial oxidation of sulfide minerals in column leaching experiments at suboptimal temperatures. Appl. Environ. Microbiol. 58, 600e606. Bailey, B.L., Blowes, D.W., Smith, L., Sego, D., 2015. The Diavik Waste Rock Project: geochemical and microbiological characterization of drainage from low-sulfide waste rock: active zone field experiments. Appl. Geochem. 62, 18e34. Bailey, B.L., Smith, L.J.D., Blowes, D.W., Ptacek, C.J., Smith, L., Sego, D., 2013. Diavik Waste Rock Project: persistence of contaminants from blasting agents in waste rock effluent. Appl. Geochem. 36, 256e270. Benner, S.G., Gould, W.D., Blowes, D.W., 2000. Microbial populations associated with the generation and treatment of acid mine drainage. Chem. Geol. 169, 435e448. Blowes, D.W., Jambor, J.L., 1990. The pore-water geochemistry and the mineralogy of the vadose zone of sulfide tailings, Waite Amulet, Quebec, Canada. Appl. Geochem. 5, 327e346. Blowes, D.W., Logsdon, M.J., 1998. Diavik Geochemistry Baseline Report. Report to Diavik Diamond Mines Inc.. Blowes, D.W., Al, T., Lortie, L., Gould, W.D., Jambor, J.L., 1995. Microbiology, chemical, and mineralogical characterization of the Kidd Creek Mine tailings impoundment, Timmins Area, Ontario. Geomicrobiol. J. 13, 13e31. Blowes, D.W., Moncur, M.C., Smith, L., Sego, S., Bennett, J., Garvie, A., Linklater, C., Gould, D., Reinson, J., 2006. Construction of two large-scale waste rock piles in a continuous permafrost region. In: 7th International Conference on Acid Rock Drainage, March 27e30, 2006, St. Louis, Mo., USA. Canadian Environmental Quality Guidelines (CEQG), 1999. Canadian Council of Ministers of Environment. Environment Canada, Winnipeg. € €m, H., 2002. Sequential extraction Carlsson, E., Thunberg, J., Ohlander, B., Holmstro of sulfide-rich tailings remediated by the application of till cover, Kristineberg mine. North. Swed. Sci. Total Environ. 299, 207e226. Chi, X., Amos, R.T., Stastna, M., Blowes, D.W., Sego, D.C., Smith, L., 2013. The Diavik Waste Rock Project: implications of wind-induced gas transport. Appl. Geochem. 36, 246e255. , L., 2002. A mineralogical and geochemical study of element Dold, B., Fontbote mobility in sulfide mine tailings of Fe oxide CueAu deposits from the Punta del Cobre belt, northern Chile. Chem. Geol. 189, 135e163. Dzombak, D.A., Morel, F.M.M., 1990. Surface Complexation Modeling: Hydrous Ferric Oxide. John Wiley and Sons, Toronto, Canada. Environment Canada, 2010. Monthly Data Report for Ekati A, Northwest Territories, Bulk Data 1998e2010. Accessible at: http://climate.weatheroffice.ec.gc.ca/ climateData/canada_e.html. Gunsinger, M., Ptacek, C.J., Blowes, D.W., Jambor, J.L., 2006. Evaluation of long-term sulfide oxidation processes within pyrrhotite-rich tailings, Lynn Lake, Manitoba.

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