Seasonal variations in the composition of mine drainage-contaminated groundwater in Dalarna, Sweden

Journal of Geochemical Exploration 90 (2006) 197 – 214 www.elsevier.com/locate/jgeoexp

Seasonal variations in the composition of mine drainage-contaminated groundwater in Dalarna, Sweden Roger B. Herbert Jr. ⁎ Department of Earth Sciences, Uppsala University, Villavägen 16, S - 752 36 Uppsala, Sweden Received 5 July 2005; accepted 27 December 2005 Available online 28 February 2006

Abstract Groundwater down-gradient from a mine rock dump in Dalarna, Sweden was sampled from the onset of snowmelt runoff (April) until October in order to investigate seasonal variations in groundwater composition. The results demonstrate that considerable variation in solute concentration (Al, Cu, Fe, SO2− 4 , Zn) and acidity occurs in groundwater; the greatest change in solute concentrations occurs during the melting of the snow cover, when sulfide oxidation products are flushed from the rock dump. During this period, groundwater flow is concentrated near the soil surface with an estimated velocity of 1 m/day. Groundwater acidity varied by a factor of four closest to the rock dump during the sampling period, but these variations were attenuated with distance from the rock dump. Over a distance of 145 m, groundwater pH increases from 2.5 to 4.0 and acidity decreases from 3–13 to 0.8–1.1 meq/L, which is the combined effect of ferric iron precipitation and aluminosilicate weathering. As a result of flushing from the upper soil horizons, peaks in total organic carbon and ammonium concentrations in groundwater are observed at the end of snowmelt. In soils impacted by acidic surface runoff, the sequential extraction of C horizon soils indicates the accumulation of Cu in well-crystallized iron oxyhydroxides in the upper C horizon, while Cu, Fe, Ni and Zn accumulate in a well-crystallized iron oxyhydroxide hardpan that has formed 2.5m below the ground surface. Surface complexation modeling demonstrates that SO2− 4 and Cu adsorb to the abundant iron oxyhydroxides at pH b 4, while Zn adsorption in this pH range is minimal. © 2006 Elsevier B.V. All rights reserved. Keywords: Snowmelt; Acidity; Pyrrhotite; Oxidation; Surface complexation; Sequential extraction

1. Introduction A major environmental threat from mining wastes is the transport of sulfide oxidation products into surrounding terrestrial and aquatic ecosystems. Pyrite (FeS2) is a common sulfide mineral in mine wastes, and its oxidation results in the release of sulfate, ferrous iron, and acidity (as protons): þ FeS2 ðsÞ þ 3:5O2 þ H2 O→Fe2þ þ 2SO2− 4 þ 2H

ð1Þ

⁎ Fax: +46 18 55 11 24. E-mail address: [email protected]. 0375-6742/$ - see front matter © 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.gexplo.2005.12.002

Other sulfide minerals may also be present in the mine wastes, such as sphalerite (ZnS) or galena (PbS), but the oxidation of these minerals does not necessarily release protons (cf. reaction 1). Following the atmospheric oxidation of pyrite and other iron sulfides in mine waste deposits, some of the released ferrous iron precipitates within or outside the mine waste deposit as hydrated Fe(II) sulfates (e.g. melanterite, FeSO4·7H2O; see Nordstrom and Alpers, 1999) or upon oxidation to ferric iron (reaction 2) as basic Fe(III) sulfates and Fe(III) oxyhydroxides (for example, goethite, α-FeOOH; reaction 3): Fe2þ þ 0:25O2 þ Hþ →Fe3þ þ 0:5H2 O

ð2Þ

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Fe3þ þ 3H2 O ¼ α  FeOOHðsÞ þ 3Hþ

ð3Þ

As can be seen in reaction 3, dissolved Fe3+ (and Al , not shown) represents significant acidity in acid waters. Variability in the discharge of sulfide oxidation products may be of critical importance to aquatic organisms that inhabit impacted ecosystems, since the episodic release of metals and acidity may be harmful to many organisms that are adapted to low acidity and metal concentrations. Such an episodic release is exemplified in the dissolution of hydrated iron sulfates. During dry periods of net evaporation from waste deposits, the formation of hydrated Fe(II) sulfates is enhanced, and provides a temporary sink for sulfate, Fe, and other heavy metals (e.g. Cu, Zn) that may substitute for Fe (Nordstrom and Alpers, 1999). These sulfates, however, are very soluble; rainfall events and snowmelt runoff from the mine waste results in the dissolution of the sulfates and consequently flushes sulfate, Fe, and heavy metals into the underlying groundwater and surrounding water bodies (see e.g. Alpers et al., 1994; Cravotta, 1994). These surges of acidic, metal-laden waters have been observed in various mine drainage systems (e.g. King, 1995; Kimball, 1999) and are occasionally associated with toxic effects on the microand macrofauna in the receiving aquatic ecosystems (e.g. fishkills; Filipek et al., 1987; Younger et al., 2002, and references therein). In many cases, pulses of sulfide oxidation products are transported first through the subsurface before discharge to surface water, and transport through the subsurface may have a significant retarding effect on acidity and solute release. In order to predict the impact of temporal variations in acidity and solute release from mine waste deposits on surface water ecosystems, it is necessary to understand the physical and biogeochemical processes occurring in the subsurface prior to leachate discharge to surface water. Thus, this study investigates the seasonal changes in groundwater composition during and after snowmelt runoff from a mine rock dump, with the aim of better understanding the relationship between the seasonal variations in catchment hydrology, solute release, and groundwater chemistry. This study is unique in that it presents an interpretation of temporal and spatial changes in groundwater chemistry down-gradient from a mine waste deposit, both during and after snowmelt. Many previous studies have examined acid pulses and dilution events in surface water during snowmelt and the spring flood (e.g. Laudon et al., 2000, 2001, 2005), but very few studies from mining environments have followed 3+

the progression of these events in groundwater prior to discharge to surface water. 2. Site description 2.1. Historical background The site for this study is an abandoned nickel mine, Rudolfsgruvan, near the city of Rättvik in the province of Dalarna, Sweden (60°48′N, 15°15′E). Rudolfsgruvan is one of a series of mines in the Slättberg mine field, which is situated along an amphibolite dike intruded into granitic gneiss. The sulfide ore is associated with the amphibolite. In 1805, Slättberg was worked for copper (Löfstrand, 1904), but by 1850 nickel was the primary metal of interest. Mining in Slättberg ceased in the 1880s, when the price for Ni dropped. Slättberg's sulfide ores were too poor to be exploited under normal conditions and, even during times of crisis with abnormally high metal prices, it was not possible to stimulate large-scale mining operations (Tegengren, 1924). Only for brief periods during the 1900s, prior to and during the two world wars, were mining operations temporarily resumed at the site. Rudolfsgruvan is a very small mine within the larger Slättberg field, and available data indicate that it consists only of a shallow shaft (ca. 30 m deep) opening into a ca. 17 m tunnel. Based on production data from 1917 (Tegengren, 1924) and the size of the mine, it is estimated that more than half of the total rock mass from Rudolfsgruvan was mined in 1917. The ore was, therefore, probably mined rather quickly over the course of several years. In 1969, the mining company Stora Kopparberg AB (hereafter “Stora”) began investigating the feasibility of resumed operations in the Slättberg mine field, and installed 20 boreholes in the mineralized zone. According to their calculations (Stora, archive material), the ore body near Rudolfsgruvan consists of an estimated 300,000 kg of ore with an average compositions of 0.5% Cu, 0.6% Ni, and 0.06% Co. At this time, Stora mined small amounts of ore from the Slättberg field, including Rudolfsgruvan, in order to investigate the metal yield after milling and floatation. Based on their investigations, Stora concluded that further exploitation of these ore bodes was uneconomical (Stora, archive material). According to the above information, the bulk of the waste rock derived from the excavation of the mine is probably no greater than 90 years old, while a smaller portion of the material (estimated to b20 tons) was deposited 30–40 years ago.

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2.2. Field site and previous investigations Waste rock derived from the excavation of the mine has been dumped directly adjacent to the main shaft, and covers an area of about 1300 m2 (see Fig. 1). The waste rock consists of rock material varying in grain size from sand to coarse gravel, and contains gangue minerals (mostly silicates) and sulfides such as nickel-rich pyrrhotite (Fe7S8), pyrite, chalcopyrite (CuFeS2), and sphalerite (ZnS). The rock dump rests on a glacial till hillslope; the glacial till has generally a sandy–silty texture, varies in thickness from 1 to 5 m across the site, and carries groundwater throughout the year. The glacial till is derived from granitic gneiss and has developed a podzolic soil profile. The till mineralogy is characterized by quartz, potassium and plagioclase feldspars, hornblende, and small quantities of micas and clay minerals (Herbert, 1995a). At the time of this study, the entire site was covered by a cultivated pine and spruce forest (approximately 50 years old) with the exception of the rock dump. The average annual precipitation (1961–1990) and potential evaporation for this area (data from city of Falun) are 684 and 485 mm, respectively, with a mean annual temperature of 3 °C.

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The atmospheric oxidation of Ni-bearing pyrrhotite, pyrite, and lesser quantities of chalcopyrite and sphalerite, combined with the percolation of water through the unsaturated rock dump, has generated a leachate characterized by low pH and high concentrations of Fe, Cu, Ni, Zn, and sulfate. Within the rock dump, large quantities of Fe oxyhydroxides, elemental sulfur, and jarosite [KFe3(SO4)2(OH)6] have precipitated (Lin and Herbert, 1997), while rozenite (FeSO4· 5H2O) and melanterite (FeSO4·7H2O) form as efflorescence precipitates on the rock fragments during the dry summer months (Herbert, 1995b). In 1988, a limestone company attempted to neutralize the acidity produced by sulfide oxidation in the deposit by covering the rock dump with limestone. A 5 cm layer of crushed limestone was added to the surface, and the deposit was irrigated with water in order to aid in the infiltration of carbonaterich waters. Because of the high rate of acid generation in the deposit, there is no indication that this treatment has had an effect on groundwater acidity or metal release. However, high Ca concentrations in the groundwater near the rock dump are probably the result of limestone dissolution in the rock dump cover (see below).

Fig. 1. Map of Rudolfsgruvan field site, Dalarna, Sweden.

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Down-gradient from the rock dump, the groundwater level and composition is monitored by a series of monitoring wells installed in the glacial till hillslope (B1, B2, B3A, B3B, B4; where “B” stands for borehole; see Fig. 1). Well R1 (where “R” stands for reference; Fig. 1) is an uncontaminated reference well and does not lie along the flow path from the rock dump. The groundwater discharges to Kambamyran, a transitional fen, which lies 145 m topographically and hydraulically down-gradient from the mine deposit (see Fig. 1). The observed spatial variability in the groundwater chemistry has been discussed by Herbert (1994, 1995b), but the temporal variability has not been reported. In addition to the groundwater investigations, previous studies have examined metal retention in the soils (Herbert, 1997b) and the mineralogical composition of soils and groundwater precipitates (Herbert, 1996, 1997a). Soil sampling pits are indicated on Fig. 1 as PR and P2 (where “P” stands for pit). X-ray diffractometry and 57Fe Mössbauer spectroscopy have indicated that poorly crystalline goethite, lepidocrocite (γ-FeOOH), and jarosite are precipitating from the Fe-rich groundwater at the water table, and the formation of these Fe phases is most likely responsible for the large decreases in Fe concentrations in the groundwater observed a relatively short distance from the rock dump (Herbert, 1994, 1996).

surface (2.9, 1.1, and 3.2 m depth, respectively) while wells P2 and PR were installed to the base of pits P2 and PR, excavated to depths of 2.2 and 2.5 m, respectively. Wells B1, B4 and R1 were constructed of Plexiglas and were screened along 0.4 m at the base of each casing; wells P2 and PR were constructed of HDPE and were screened along the lower 1 m of each casing. It should be noted that wells P2 and PR are relatively shallow, and that both wells were dry during the summer and part of the autumn. Therefore, no results are reported from this period for these wells. Prior to groundwater sampling, the groundwater level was measured in each well. Groundwater samples were collected after pumping three well volumes of water from the wells and allowing the water table to recover. During this period, the pH and temperature of the water stabilized. All samples (except for samples for total organic carbon, TOC) were vacuum filtered through 0.45 μm membrane filters, whereupon samples for cation analysis were preserved by acidifying with concentrated nitric acid to a pH less than 2; samples for anion, ammonium, and TOC analyses were left untreated. After collection, the samples were stored at 4 °C and transported directly to the laboratory.

3. Methodology

Variations in groundwater chemistry after the onset of snowmelt were monitored by measuring pH, redox potential (Eh), TOC, sulfate, ammonium (NH4+), nitrate (NO3− ), and the concentrations of various metals. Nitrate, ammonium, and TOC are determined as these components are derived from the upper soil horizons; the presence of these components in groundwater indicates the mixing of soil water with groundwater. The pH and temperature of groundwater samples were measured in the field with a combination pH (Orion) electrode. The redox potential was obtained from the electrode potential of a platinum electrode coupled with a Ag/AgCl reference cell (Reagecon), and has been corrected to the standard hydrogen electrode. On all reported sampling occasions, Al, Cu, Fe, sulfate and Zn concentrations in groundwater have been analyzed in all wells; on two occasions (days 172, 204), Ca, K, Na, and Mg concentrations were also determined. Metal concentrations were determined by ICP-AES or atomic absorption spectrometry. In most samples, sulfate was determined turbidimetrically on a HACH spectrophotometer using a modification of the barium sulfate method. However, samples from well B1 contained high levels of ferric iron which partially interfere with this technique, so sulfate in B1 was measured by ion

3.1. Groundwater sampling In order to study the seasonal changes in groundwater composition following snowmelt, groundwater sampling commenced in April 1995, at the onset of snowmelt. A cold spring and much snow in early May resulted in a delayed and intense snowmelt runoff in late May, providing good conditions for observing the effects of snowmelt on groundwater chemistry. Subsequent samples were collected on eight occasions from May to October 1995 (950427, 950522, 950609, 950621, 950710, 950723, 950822, 950914, and 951026). Sampling dates will be referred to as a day number from the beginning of the year, so that sampling on 950427 is given as day 117. Sampling during the winter was not possible as the groundwater freezes in the monitoring wells. For this study, the analytical results from wells B1, PR, P2, B4 and R1 will be discussed (Fig. 1); wells B2, B3A, and B3B are not discussed as they were not continuously monitored during this period. Details about well installation and depth are described in Herbert (1994, 1995a). Briefly, wells B1, B4 and R1 were installed directly above the bedrock

3.2. Groundwater analyses

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chromatography. Nitrate, ammonium, and TOC were measured in wells B1 and R1. Nitrate and NH4+ were determined on a Technicon autoanalyser by the Cd reduction and the indophenol-blue methods, respectively. TOC was analyzed by catalytic combustion and IR detection of released CO2. 3.3. Soil sampling In May 1993, the podzol soils from the Rudolfsgruvan site were sampled from pits P2 and PR (Fig. 1). Following sampling, sequential extraction techniques were used to evaluate the partitioning of various heavy metals within the contaminated soil profiles and between different soil fractions. The results from the A, E, B and upper C horizons have been previously published (Herbert, 1997b) and will not be discussed in this current paper; instead, results from the lower C horizon, which is in contact with groundwater during a much greater part of the year, will be presented. The pits P2 and PR were excavated using a backhoe to depths of 2.2 and 2.5 m, respectively. At the time of excavation, the depth to water in well B1 was 2.3 m below the ground surface. Soil samples were collected using a stainless steel trowel from 3 to 4 depths in the two pits. Soil sampling was carried out by collecting soil from several locations in the pit, at the same depth and horizon, and then mixing the samples together to form bulk samples with a total volume of 1–3 dm3. In general, the C horizon soils were dull yellow orange in color. At the base of each pit, there was evidence for iron oxide precipitation. In PR, massive iron oxide precipitates had cemented together soil particles and formed a hardpan; this hardpan was sampled and is labeled PR 250⁎ (“250” for depth in cm). In P2, such massive precipitates were not present, but the soil was mottled indicating iron precipitation; this sample is labeled P2 220. After sampling, soils were transported to the laboratory where they were dried at room temperature for 48 h. The samples were subsequently disaggregated gently and then dry-sieved for grain-size analysis. The fraction less than 63 μm is used in this investigation, which accounts for about 40% of the total sample mass that passes through a 2 mm mesh. Within this fraction, the clay fraction (b2 μm) accounts for less than 10%. 3.4. Soil analyses Extractions were performed to liberate metals bound to (1) cation exchange sites, (2) organic matter and sulfides, (3) poorly crystalline iron oxides, (4) well-

201

crystallized iron oxides, and (5) residual primary minerals. The extraction procedures are summarized in Table 1, which is a sequential extraction scheme similar to that described by Tessier et al. (1979). Because of the nature of the mine drainage waters (i.e. acidic, Fe-rich), it was assumed that metal partitioning in the Fe oxide fraction would be important in the polluted soils, where the metals would be retained by either specific adsorption to the oxide surfaces or by occlusion in the Fe oxide lattices. In the extraction scheme by Tessier et al. (1979), metals bound to organic matter are traditionally extracted after metals bound to iron oxides. However, since it was a concern that the acidic reagents used for the extraction of iron oxides would also partially extract metals from organic matter, the step for the extraction of organic matter and sulfides has been moved prior to the step for the extraction of iron oxides (Table 1). All extractions, with the exception of the final extraction for residual cations, were performed in 50 mL polyethylene centrifuge tubes. A one gram sample was used for the first step in the extraction procedure. After the extraction with a particular extraction agent, the sample was centrifuged at 4000 rpm for 30 min, whereupon the solution was decanted. The samples were subsequently washed with 10 mL de-ionized water for 20 min in a centrifuge at 4000 rpm. This volume of water was decanted, whereupon the next extractant was added. Separate one gram samples were digested in aqua regia and were subsequently filtered into 100 mL volumetric flasks. After extraction, samples were analyzed for Fe, Cu, Ni, Pb, and Zn by flame atomic Table 1 Sequential extraction procedures used in this investigation Metal-bound fraction

Extractant and methodology

Exchangeable

10 mL 1 M MgCl2 (pH 7) for 1 h at room temperature with continuous shaking. 10 mL 30% hydrogen peroxide (H2O2) in 0.05 M HNO3 for 5 h in a water bath at 85 °C, with intermittent agitation. After cooling, 5 mL 3.2 M ammonium acetate in 20% HNO3 and 5 mL de-ionized water were added, whereupon the sample was continuously shaken for 30 min at room temperature. 25 mL 0.25 M NH2OH·HCl in 0.25 M HCl for 30 min in a water bath at 50 °C, with occasional agitation. 20 mL 0.20 M NH2OH·HCl in 25% acetic acid for 18 h in a water bath at 90 °C, with occasional agitation. Hot dissolution in 15 mL aqua regia (3 : 1 solution of concentrated HCl : HNO3) for 1 h.

Organic matter

Poorly crystalline Fe oxides Well crystallized Fe oxides Residual

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absorption spectrometry (AAS). Analytical precision of AAS based on repeated measurements is estimated at ± 5%. Standard solutions for AAS analysis were made up in the same reagent solutions as were used in each respective extraction step. Reagent blanks and deionized water blanks were used to account for any reagent or procedural contamination, and were analyzed with every set of samples. The analysis of these blanks indicated negligible contamination. It should be noted that these extraction steps are operationally defined, and there is considerable overlap in the extraction in each step. For example, a portion of the metals bound to poorly crystalline iron oxides may in fact be extracted in the step designed for metals bound to organic matter. For a discussion of these particular extraction steps and further details on the analytical procedures, the reader is referred to Herbert (1997b). All samples were analyzed in triplicate for quality control assessment. 3.5. Geochemical modeling Aqueous speciation, mineral equilibria calculations and adsorption modeling were conducted using PHREEQC and the MINTEQ database (Parkhurst and Appelo, 1999). When base cation results were not available (see above), the speciation calculations yielded a large negative charge imbalance in the sample; charge balance was attained in these samples by assuming that all of the missing cationic charge was present as calcium. This is a fairly reasonable assumption (see Results, below). From the speciation calculations, groundwater acidity was calculated: Acidity ¼ ½Hþ  þ 3½Fe3þ  þ 2½FeOH2þ  3þ 2þ þ ½FeðOHÞþ 2  þ 3½Al  þ 2½AlOH  þ þ ½AlðOHÞ2 

ð4Þ

In this equation, brackets denote molar concentrations. Because of the low groundwater pH (see Results) and high metal concentrations, contributions to acidity from other hydrated Fe and Al species (e.g. Fe(OH)30, Fe(OH)4−) and organic acids are negligible. Iron redox speciation (i.e. Fe(II) and Fe(III)) was determined using the assumption that Fe2+ and Fe 3+ are in redox equilibrium and that the redox potential measured using a platinum electrode is poised by the Fe2+/Fe3+ redox couple. For acid mine waters (i.e. low pH) with high iron concentrations, this assumption is generally valid (Nordstrom et al., 1979; Langmuir, 1997). In order to investigate the influence of adsorption reactions on groundwater composition at the investigat-

ed site, the specific adsorption of Cu2+, Zn2+, Fe2+, H+ and SO42− on iron oxyhydroxides was simulated using PHREEQC. Together with the results of the groundwater and soil sampling, the surface complexation modeling is used to interpret the adsorption behavior of Cu, Zn, Fe, sulfate, and protons down-gradient from the rock dump. A two-site, diffuse double-layer model was used according to surface complexation reactions and constants that are presented in Dzombak and Morel (1990) and Appelo et al. (2002), which are incorporated in the PHREEQC database. The reactions and equilibrium constants for the most significant complexation reactions on iron oxyhydroxide (referred to as “hydrous iron oxide” in Dzombak and Morel, 1990) are shown in Table 2. The diffuse double-layer model requires the input of adsorbent concentration, adsorbent surface area, and number of adsorption sites; these parameters will be discussed later in the paper as they are derived from the results of the soil analyses. For the modeling, the adsorption of Zn2+, Cu2+, Fe2+, SO42− and protons to both weak sites and strong sites on an iron oxyhydroxide surface was simulated as batch reactions at constant pH (i.e. adsorption at a specific pH value was independent of the results from other pH values), where the pH range 2 to 7 was investigated. All these ions were included in the simulations so as to determine the degree of competition that exists for binding sites. A constant ionic strength of 0.1 M was maintained in all solutions. Since groundwater composition varies during the year, the simulations were conducted for two different cases corresponding to high ion concentrations in well B1 and low ion concentrations in well PR. Concentrations of 1 mM Cu, 1 mM Zn, 50 mM Fe, and 100 mM SO42− were used for the high Table 2 Reactions and equilibrium constants used in surface complexation simulations logKint, logKint, weak sites strong sites

Adsorption reaction ≡FeOH þ Hþ ¼ ≡FeOHþ 2 −

≡FeOH ¼ ≡FeO þ H ≡FeOH þ Zn

2þ

þ þ

¼ ≡FeOZn þ H

þ

≡FeOH þ Cu2þ ¼ ≡FeOCuþ þ Hþ ≡FeOH þ Fe

2þ

þ

¼ ≡FeOFe þ H

þ

7.29

7.29

−8.93

− 8.93

−1.99

0.99

0.6

2.89

−2.98

þ − ≡FeOH þ SO2− 4 þ H ¼ ≡FeSO4 þ H2 O

7.78

2− ≡FeOH þ SO2− 4 ¼ ≡FeOHSO4

0.79

− 0.95

All intrinsic equilibrium constants (Kint) are from Dzombak and Morel (1990), except those for Fe2+ which are taken from Appelo et al. (2002).

R.B. Herbert Jr. / Journal of Geochemical Exploration 90 (2006) 197–214

concentration case, and 0.03 mM Cu, 0.03 mM Zn, 0.3 mM Fe, and 6 mM SO42− were used for the low concentration case. These are approximately equivalent to the maximum and minimum concentrations measured near the rock dump during the sampling episodes during 1995 (see Results). 4. Results 4.1. Snowmelt and groundwater levels At this site, partial or complete snow cover was present until about day 140 (mid May), when the snow cover had completely melted; during this period, more than half the days had mean daily temperatures above 0 °C (SMHI, 1995, Backa field station). At the end of snowmelt (ca. day 140), melt water had collected into small pools around wells B1, P2, and B2, and water was observed seeping along the ground surface from B1 towards P2; the total volume of water in the pools is estimated to 5 m3. On day 142, this acidic surface runoff near well B1 had pH 2.4 and 3300 mg/L SO42−, 433 mg/L Fe, 38 mg/L Al, 15 mg/L Cu and 8 mg/L Zn. This ponded surface water had infiltrated into the hillslope by the beginning of June. The groundwater table by the end of snowmelt coincided with the ground surface in wells B1, P2, B2, and R1, and indicates that the hillslope was completely water saturated in many locations (Fig. 2). After snowmelt, the water table steadily falls during the summer months. As shown in Fig. 2, the water table elevation in B1, B2 and R1 decreased by more than 1.5 m from the end of May until the end of August, but begins to recover during the autumn months. In

Fig. 2. Groundwater table elevation in wells B1, B2, B4, and R1 during sampling episodes. Numbers above curve indicate day of the year.

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comparison, the water table elevation in well B4 shows little variation as the phreatic surface is largely controlled by the water level in the nearby wetland. If the hydraulic gradient is considered over the entire hillslope from well B1 to B4, then the hydraulic gradient decreases by about 35% (from ca. 0.039 on day 142 to ca. 0.026 on day 234) during the period from snowmelt runoff to late summer; however, for a hydraulic gradient calculated in the vicinity of the rock dump (from B1 to B2), the gradient increases by 17% (from ca. 0.031 on day 142 to ca. 0.038 on day 234). Previous geophysical measurements (Herbert, 1995b) have indicated that the leachate plume is relatively narrow and extends from the rock dump, past B1, P2, and B2, to B4 and the wetland. Although it is not entirely apparent on Fig. 1, well PR is situated topographically up-gradient from well P2. Wells B1 and P2 lie much closer to the centerline of the plume than well PR, which lies on the edge of the leachate plume and is less affected by mine leachate (see below). 4.2. Groundwater chemistry The analytical results indicate that metal and sulfate concentrations are generally highest and groundwater pH lowest closest to the rock dump (Fe concentrations N 900 mg/L, sulfate concentrations N 3500 mg/L, pH b 3), with lower concentrations and higher pH detected with increasing distance from the rock dump. Metal and sulfate concentrations are low in well R1 and represent background concentrations for this mineralized area. The groundwater analyses from this study are in agreement with previous studies (Herbert, 1994, 1996). Variations in pH and in metal and sulfate concentrations during the sampling episodes in 1995 are shown in Figs. 3 and 4, respectively. Selected results for these wells and also reference well R1 are presented in Table 3. In general, groundwater pH is highest during snowmelt runoff and decreases toward the end of this period (days 130–150; Fig. 3). In well B1, two pH minima are observed in the time series, with the first coinciding with the end of snowmelt runoff. Two minima are not observed in the time series for well P2 or PR, as these wells were dry after days 170 and 200, respectively. Only one minimum is observed in well B4. Iron and sulfate concentrations varied by up to a factor of five during the monitoring period. Close to the rock dump (i.e. wells B1 and P2), relatively low solute concentrations are detected during snowmelt (i.e. prior to day 130; Fig. 4), but towards the end of the period of snowmelt runoff (days 130–150), a significant peak in

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Fig. 3. Groundwater pH during the sampling episodes.

metal and sulfate concentrations is observed. In B1, this peak is observed on day 142, while in P2 it is observed on day 172. In addition to the first peaks, a second maximum in metal and sulfate concentrations is measured in well B1 on day 172. These solute peaks correspond to the pH minima mentioned above. Solute concentration peaks are not observed for well PR (Fig. 4), although this is probably because the time series for

well PR is rather short. Metal and sulfate concentrations also vary in well B4 during the sampling period (Fig. 4), but these variations are not as great as observed in B1, P2, and PR. Near the rock dump, low pH and high Al and Fe concentrations coincide with high groundwater acidity. Since the pH is quite low in wells B1 and P2 (2.2– 2.5), the proton concentration is the primary component in the acidity calculation (see Eq. (4)). Therefore, as shown in Fig. 5, the acidity peaks occur at the same time as low pH (Fig. 3). Acidity in well PR is, on two sampling occasions, less than the acidity in well B4 despite the close proximity of PR to the rock dump. This supports previous investigations (Herbert, 1995b) that indicate that well PR lies on the edge of the leachate plume and is significantly affected by mixing with uncontaminated groundwater. Saturation index calculations with PHREEQC indicate that groundwater in wells B1, P2, PR and B4 is generally close to saturation or oversaturated with respect to the iron oxyhydroxides goethite and lepidocrocite throughout the sampling period. The groundwater is consistently undersaturated with respect to ferrihydrite, a poorly crystalline iron oxide. For the two sampling occasions when base cation analyses

Fig. 4. Concentrations of sulfate, Al, Cu, Fe, and Zn in groundwater wells B1, PR, P2, and B4 during the sampling episodes in 1995.

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Table 3 Concentrations of various constituents in groundwater for four sampling episodes, corresponding with occurrences of maximum concentrations in B1 (day 142), two sampling episodes where base cations were also analyzed, and the sampling occasion when the water table was at its lowest level (day 234) Eh

SO2− 4

Al

Ca

Cu

Fe

K

Mg

Na

Zn

1995-05-22, 142 days B1 2.11 P2 2.57 PR 4.09 B4 4.14 R1 6.00

n.a. n.a. n.a. n.a. n.a.

13000 2300 37 575 n.a.

216 79 1.8 35 0.62

n.a. n.a. n.a. n.a. n.a.

101 14 0.12 4.0 0.016

4670 295 2.0 0.15 0.54

n.a. n.a. n.a. n.a. n.a.

n.a. n.a. n.a. n.a. n.a.

n.a. n.a. n.a. n.a. n.a.

68 8.3 0.15 3.3 0.058

1995-06-21, 172 days B1 2.42 P2 2.27 PR 3.36 B4 4.11 R1 5.39

676 706 650 549 324

8800 4750 575 628 9.0

179 143 43 37 0.14

280 183 51 97 5.3

74 32 2.9 4.4 0.014

3200 950 15 0.08 b0.01

1.2 0.60 2.8 1.4 0.70

70 49 11 24 0.73

7.2 5.8 3.5 4.6 1.9

41 19 2.6 3.5 0.089

1995-07-23, 204 days B1 2.12 P2 2.56 B4 4.07 R1 5.41

688 686 576 478

6200 2900 550 8.0

153 165 32 0.11

265 200 84 4.0

53 28 4.1 0.007

2220 793 0.08 0.13

0.80 0.70 1.1 0.40

69 49 20 0.53

6.3 5.8 4.2 1.8

34 16 3.1 0.018

1995-08-22, 234 days B1 2.49 B4 4.01 R1 5.32

702 580 488

3800 410 15

125 n.a. 0.17

n.a. n.a. n.a.

34 4.0 0.027

1250 0.47 0.45

n.a. n.a. n.a.

n.a. n.a. n.a.

n.a. n.a. n.a.

24 2.8 0.12

Well

pH

All concentrations are given in mg/L except pH and Eh (mV). Well P2 was dry on day 234 and well PR was dry on days 204 and 234, so no data are shown for these dates. n.a. = not analyzed.

were available, the groundwater in wells B1, P2, PR and B4 was close to saturation or slightly undersaturated with respect to gypsum, and was oversaturated with respect to jarosite in B1, P2 and PR. Aluminum hydroxides were consistently undersaturated in all groundwater samples. As was observed for metal and sulfate concentrations, there is a maximum in TOC and ammonium concentrations in well B1 at the end of snowmelt (day 132), with concentrations decreasing after this point (see Fig. 6). Dissolved organic carbon (DOC) was also measured in some samples (results not shown), and indicated that most of the TOC is in fact DOC. Nitrate concentrations are generally much lower than ammonium concentrations in B1, except between days 170 to 200, when nitrate concentrations are ca. 2.7 mg/L and exceed ammonium levels. In contrast, TOC concentrations in reference well R1 are low (b 2 mg/L), and ammonium concentrations are generally below detection limits (b 0.02 mg/L). Nitrate in the reference well is present above detection limits (0.005 mg/L) on only a few occasions, with a maximum concentration of 1.3 mg/L nitrate measured on day 172.

4.3. C horizon chemistry The results of the sequential extraction of C horizon soils and precipitates are presented in Fig. 7 and Table 4. The total Fe concentrations in the C horizons of the P2 and PR profiles are very similar. The total Ni and Zn concentrations in profile P2 appear to be somewhat greater than in PR, but this difference may not be significant in light of the variation among samples because of soil heterogeneity. Total Cu concentrations in the C horizon of profile P2 are substantially greater than in PR. Metal concentrations were the greatest at the base of each pit where contact time with groundwater is greatest; total Fe, Cu and Ni concentrations were high in the hardpan at the base of PR. The partitioning of Fe, Cu, Zn and Ni in different geochemical fractions is depicted in Fig. 7 and Table 4, and shows that the metals are primarily extracted in the well-crystallized iron oxide and residual fractions in the C horizon samples. While the residual fraction is generally considered to consist of resistant silicates, it may also in this case be composed of well-crystallized iron oxyhydroxides that were not completely extracted with the hydroxylamine hydrochloride/acetic acid

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Fig. 5. Groundwater acidity during the sampling episodes in 1995. Note different y-axis scales in the upper and lower diagrams. Al estimated for last 3 sampling episodes in well B4.

solution (Table 1). Partitioning in the poorly crystalline iron oxide fraction was not as significant relative to the other fractions. For the PR hardpan (sample PR 250⁎), relatively large fractions of Cu and Ni are also extracted as exchangeable cations and in the organic fraction. 4.4. Surface complexation modeling The model PHREEQC has been used to simulate the adsorption of Cu2+, Zn2+, Fe2+, SO42− and protons to

iron oxyhydroxide surfaces in the soils at the Rudolfsgruvan site. As indicated by the C horizon analyses (see above) and mineralogical data (Herbert, 1997a), there are iron oxyhydroxide surfaces available in the soil for ion adsorption. The total Fe concentrations are similar in pits P2 and PR in the C horizon above 2 m depth, so there is no clear indication of excessive Fe accumulation in these horizons as a result of mining activities. However, there has been a large accumulation of Fe as iron oxyhydroxides at depths greater than 2 m below the ground surface (see Fig. 7), where groundwater occurs for much of the year (cf. Fig. 2). This appears to be the region of greatest Fe precipitation. Therefore, the parameters required for the diffuse double-layer model in PHREEQC are acquired from the deepest sample from the C horizon in pit P2. The adsorbent concentration was calculated from the concentration of iron in poorly crystalline iron oxyhydroxides (e.g. ferrihydrite), as these phases have much greater surface areas than well-crystallized iron oxyhydroxides (Cornell and Schwertmann, 1996) and will therefore contribute to ion adsorption to a much greater extent. For an iron concentration of ca. 3000 mg/kg (Table 4), together with an estimated bulk density of 1.25 kg/dm3 and 40% porosity (data adapted from Nyberg, 1995), this gives in an Fe concentration of 0.17 mol Fe/L water and an adsorbent concentration of 15 g iron oxyhydroxide/L water (formula weight of 89 g iron oxyhydroxide/mol Fe). The adsorption site concentration was calculated according to the relation 0.2 mol weak sites per mol Fe and 0.005 mol strong sites per mol Fe (Dzombak and Morel, 1990), resulting in weak and strong site concentrations of 3.4 × 10− 2 and 8.4 × 10− 4 mol/L water, respectively. The adsorbent surface area was assigned a value of 600 m2 /g (Dzombak and Morel, 1990). The results of the surface complexation modeling are presented in Fig. 8 for Cu, Zn and SO42−. Compared with these three components, Fe2+ adsorption is negligible

Fig. 6. Nitrate, ammonium, and TOC concentrations in well B1 and well R1.

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207

Table 4 Results of sequential extraction of C-horizon samples Exch.

Org.

PC FeOx

WC FeOx

Residual

462 555 1721 340 231 368 863

999 760 5151 861 609 769 3599

3717 3791 22 754 3845 4203 4330 8890

4038 4286 10658 3494 5350 5091 6394

Iron PR 100 PR 240 PR 250⁎ P2 105 P2 125 P2 165 P2 220

0.66 1.6 234 2.21 1.15 1.37 39.23

0.36 0.20 3.2 0.40 0.33 0.47 1.6

Org.

PC FeOx

WC FeOx

Residual

0.46 0.73 54.5 5.1 5.6 7.2 5.4

0.50 0.50 9.2 2.5 2.9 3.4 3.7

5.9 7.2 152 11.2 14.6 18.8 27.2

1.9 3.1 b0.01 2.7 5.6 7.1 4.0

0.73 0.74 32.6 1.54 1.19 1.47 3.0

0.25 0.33 2.5 0.58 0.91 0.84 3.3

3.8 3.4 41 5.2 5.3 5.8 12.1

4.6 4.4 15.9 4.3 5.4 6.8 14.7

Copper

Zinc PR 100 PR 240 PR 250⁎ P2 105 P2 125 P2 165 P2 220

Exch. b0.01 b0.01 14.3 2.9 1.9 2.2 5.5 Nickel

0.79 0.80 4.2 1.4 1.3 1.4 4.5

0.82 1.0 2.0 1.0 1.4 1.3 1.5

10.4 12.0 45.5 13.2 16.9 14.9 14.8

4.0 5.6 3.3 2.4 4.8 3.7 4.2

0.43 0.27 47.5 0.17 0.00 0.00 1.3

All concentrations in units of mg/kg dry weight. Numbers after P2 or PR indicate sampling depth in cm. “Exch.” = exchangeable cation fraction, “Org.” = organic fraction, “PC FeOx” = poorly crystalline iron oxide fraction, “WC FeOx” = well crystallized iron oxide fraction.

and is therefore not shown. These observations are qualitatively in agreement with many other studies that have demonstrated high sulfate adsorption at low pH, and that adsorption isotherms for Cu adsorption to iron oxides lie at lower pH than Zn (cf. Langmuir, 1997). At low ion concentrations and low pH, all the available SO42− is complexed to weak surface sites as ≡FeSO4−, while ≡FeOHSO42− surface complexes are quantitatively more important as pH approaches 7; this change in surface speciation reflects the predominance of protonated ≡FeOH2+ surface sites at low pH and uncharged ≡FeOH surface sites at neutral pH. At high ion concentrations, less than 20% of the dissolved SO42− is adsorbed to the iron oxyhydroxide surfaces, which is expected since there are not enough weak binding sites (0.034 mol/L) to completely adsorb the dissolved sulfate at high sulfate concentrations (i.e. 0.1 mol/L). For the conditions of the PHREEQC simulations used in this study, Cu2+ adsorbs to strong binding sites on iron oxyhydroxide surfaces at pH ∼3, while Zn2+ adsorption is not significant until pH ∼4 (Fig. 8). In the groundwater pH range measured in this study (i.e. 2.1 b pH b 4.1; Table 3), Cu and Zn are primarily adsorbed to strong binding sites as ≡FeOCu+ and ≡FeOZn+ , respectively. At low pH, protons outcompete Cu2+ and Zn2+ for strong binding sites, while protons and sulfate dominate the weak binding sites over the pH range 2–7. The degree of Cu2+ and Zn2+ adsorption is greater at lower ion concentrations

since there is less competition with SO42− for ≡FeOH complexation sites (see Table 2). 5. Discussion 5.1. Biogeochemical processes controlling metal and sulfate mobility As is indicated by the results of the groundwater investigation, Al, Cu, Fe, Zn and SO42− concentrations decrease during transport in the groundwater from the rock dump, while pH increases. Since the major cations Ca, Mg, K, and Na were only analyzed in two samples, this discussion will be limited to Al, Cu, Fe, SO42−, and Zn. Part of this change in composition and pH can be attributed to physical dilution with uncontaminated groundwater and percolating rainwater, and part can be attributed to biogeochemical processes including iron oxidation and precipitation, aluminosilicate weathering, and metal adsorption. In addition to these processes controlling metal and sulfate concentrations, the results also indicate that organic matter decomposition and nitrification are important processes in terms of the mobilization of carbon and nitrogen in the hillslope; these processes are discussed later in this section. The decrease in Fe concentrations in groundwater with distance from the rock dump, as indicated by previous mineralogical studies (Herbert, 1997a) and the chemical analysis of soil samples (Table 4, Fig. 7), and as suggested by the saturation index calculations, is the

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result of the precipitation of Fe oxyhydroxides (e.g. goethite, lepidocrocite) and jarosite from the groundwater. Jarosite is only stable at pH b 3 (Nordstrom and Alpers, 1999), and its formation is only expected close to the rock dump. The sequential extraction of C horizon soils shows that most of the iron in the soils is present in well-crystallized iron oxyhydroxides or the residual fraction (e.g. amphiboles), with most of the iron accumulation below a depth of 220–250 cm (Fig. 7). However, it is likely that iron initially precipitates as poorly crystalline goethite (ferrihydrite is consistently undersaturated in the groundwater), and later transforms to a more stable, well-crystallized goethite upon aging. This is a common phenomenon among iron oxides, as the lower interfacial free energy of poorly crystalline iron oxides promotes the nucleation and crystal growth of these phases relative to iron oxides with well-ordered crystal structures (Cornell and Schwertmann, 1996). Much Fe is probably precipitated before the groundwater reaches well B1, as suggested by the Fe / S molar ratios (data not shown) in well B1 that generally lie between 0.35 and 0.65. Since pyrrhotite (Fe7S8) is the major iron sulfide in the rock dump, the rock dump leachate has probably a Fe / S ratio closer to 0.875, but this ratio will decrease with distance from the rock dump as Fe precipitates from the groundwater. Since Fe(II) is the dominant redox state of Fe in the groundwater, as indicated by the PHREEQC calculations, then Fe precipitation must be preceded by Fe(II) oxidation (reaction 2). Although dissolved oxygen (DO) was not measured in this study, it is likely that molecular oxygen enters the subsurface by either the percolation of DOcontaining water through the soil to the water table or by the diffusion of gaseous oxygen through the unsaturated soil column. The transport of DO with percolating water can be expected during and after snowmelt runoff, as significant volumes of melt water percolate into the soils, while oxygen diffusion can be expected throughout the year and should dominate when the degree of water saturation in the soil is relatively low. Close to the rock dump, sulfate has been observed to be precipitating from the groundwater as jarosite (Herbert, 1997a). However, the effect of jarosite precipitation on sulfate concentrations is probably small relative to the changes in concentration due to dilution (see below). Groundwater is generally undersaturated with respect to gypsum, so this is not expected as a sulfur sink. Aside from dilution and to a lesser degree jarosite precipitation, adsorption is likely the major process attenuating sulfate concentrations in the groundwater. Previous studies (Herbert, 1996) have indicated that the iron oxyhydroxide precipitates in pit

PR contain up to several percent sulfate. Sulfate is probably adsorbed and then encapsulated in the growing iron oxyhydroxide precipitates; this is supported by the results of the surface complexation simulations (Fig. 8) that indicate a high degree of sulfate adsorption to iron oxyhydroxides in the soil C horizon. However, the complete adsorption of sulfate on iron oxyhydroxides at lower ionic concentrations, as shown by the surface complexation simulations, is not in agreement with the groundwater measurements as high sulfate concentrations are continuously present in the groundwater (Fig. 4). There are a number of possible explanations for this disagreement: the concentration of weak binding sites may be overestimated, the concentration of iron oxyhydroxides in the soil C horizon may be overestimated, or the rate of sulfate flux from the rock dump may exceed the rate of sulfate adsorption to oxide surfaces. In any of these cases, it is evident that further studies would be needed to better address the inadequacies of the geochemical model. Aluminum is present at relatively high concentrations in the groundwater (Fig. 4), such that Al significantly contributes to groundwater acidity (Fig. 5). Although Al concentrations decrease with distance from the rock dump, Al is not expected to precipitate as does Fe, since Al oxides and basic Al sulfates are unstable at the Al concentrations and pH levels (pH b 4) that dominate in the groundwater at the site. The decrease in Al concentration is attributable to dilution, while it is likely that aluminosilicate weathering under acidic conditions in the glacial till hillslope provides a source of Al (and base cations Ca, Mg, K, Na) in groundwater down-gradient from the rock dump; this has been proposed in other studies as well (Herbert, 1995a,b). The heavy metals Cu, Zn, and Ni (data not shown for Ni) are present in well B1 in concentrations greater than 20 mg/L. During transport to the wetland, the metal concentrations decrease by approximately an order of magnitude (cf. Table 3, Fig. 4). The results from the sequential extraction of C horizon soil indicate that all these metals are accumulating to some degree in the deeper soil horizons. For C horizon soils above the base of the pits (i.e. excluding samples P2 220 and PR 250⁎; Fig. 7), there are higher total concentrations of Cu in soils from pit P2 compared with pit PR, but there does not appear to be a large difference in Ni and Zn concentrations. Higher Cu concentrations in the soils from P2 can be attributed to the infiltration of the acidic surface runoff near pit P2 but not near pit PR, and to the higher metal concentrations in well P2 relative to well PR.

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209

Fig. 7. Metal content in each operationally defined fraction of soils collected from the C horizon of pits P2 and PR, as determined by sequential extraction and analysis. Numbers after P2 and PR along y-axis indicate depth in cm. Error bars represent the sum of the standard deviations for each extraction step, which were performed in triplicate. Cu and Ni bars have been truncated for sample PR 250⁎ so that the x-axes would have the same scale (see Table 4 results).

Compared to the upper C horizon samples, samples from the base of the pits show a clear enrichment in all the studied metals (i.e. Cu, Fe, Ni and Zn); the greatest metal accumulation is observed at the base of pit PR where the groundwater pH was in the range 3.4–4.1 during this study. In the iron hardpan from pit PR, Zn is less enriched compared with Cu and this is in agreement with the surface complexation modeling. Although Zn adsorption at pH b 4.1 is minor compared to Cu, it is not negligible. Indeed, the PHREEQC simulations demonstrated that 0.02% and 1.5% of dissolved Zn2+ is adsorbed to iron oxyhydroxide surfaces at pH 3 and pH 4, respectively (data not shown); the flux of Zncontaminated groundwater through the C horizons soils over the course of many decades has resulted in the accumulation of Zn, despite its very low level of adsorption in this pH range. While Cu, Ni and Zn are

primarily associated with the well-crystallized iron oxyhydroxide fraction and the residual fraction in the C horizon soils, it is likely that the metals first adsorbed to poorly crystalline iron oxyhydroxides that have transformed to more stable, well-crystallized forms upon aging (see above). 5.2. Normalization of groundwater concentrations In order to facilitate the interpretation of changes in groundwater chemistry, it is useful to normalize metal and sulfate concentrations to the concentration of a conservative ion in groundwater. Since advection and dispersion (i.e. dilution) are the only processes that significantly affect the concentration of a conservative ion in groundwater, this would be a reasonable approach to separate the effects of dilution and biogeochemical

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processes on solute concentrations. Anions such as chloride or bromide are traditionally considered conservative ions in most groundwater environments, but these anions have not been measured in this study, and do not have a source in the rock dump. In this study, none of the analyzed components can be considered a conservative solute, but Zn is least impacted by attenuation processes at low pH, among the solutes discussed in this study, and there is no known source for Zn in the hillslope. The surface complexation modeling indicates that zinc concentrations should not be significantly attenuated in groundwater below pH 4 (Fig. 8); as the pH approaches 4 (e.g. wells PR and B4), the PHREEQC simulations indicate that less than 5% of the dissolved Zn will be adsorbed, while N 50% of the dissolved Cu and SO42− will be adsorbed (Fig. 8, low ion concentration model). Furthermore, the precipitation of secondary Zn minerals is not expected, according to the saturation index calculations with PHREEQC. While it is recognized that Zn is not a conservative tracer, it should nevertheless be minimally affected by retention processes in the close vicinity of the rock dump where the pH is less than 4 (e.g. wells B1, P2, PR). Therefore, in order to provide an indication of the relative effects of biogeochemical processes and physical dilution, SO42−,

Fig. 8. Fraction of total Cu, Zn or SO2− 4 concentration adsorbed to iron oxyhydroxide surface. “High” refers to high concentration case that was simulated (1 mM Cu, 1 mM Zn, 50 mM Fe, 100 mM SO2− 4 ) and “low” refers to low concentration case (0.03 mM Cu, 0.03 mM Zn, 0.3 mM Fe, 6 mM SO2− 4 ). Hatch area between adsorption isotherms indicates intermediate concentrations, and bold lines within the Zn and Cu isotherms represent fraction of Zn and Cu adsorbed when only these metals are considered in the model (i.e. no competition exists with other ions). The following parameters for the diffuse double-layer model were used: adsorbent concentration = 15 g iron oxyhydroxide/L water; weak adsorption site concentration = 3.4 × 10− 2 mol/L water; strong adsorption site concentration = 8.4 × 10− 4 mol/L water; adsorbent surface area = 600 m2/g.

Fe, Al, and Cu concentrations have been normalized to Zn concentrations (Fig. 9). According to the assumptions presented above, a constant component / Zn ratio (Fig. 9) with distance from the rock dump indicates that dilution and/or biogeochemical processes are affecting the component and Zn concentrations to the same extent; an increasing ratio suggests ion release from the hillslope, and a decreasing ratio indicates retention relative to Zn. The S / Zn molar ratios in B1 and P2 increase during snowmelt, and then decline to approximately the same level after snowmelt (Fig. 9a). The high S / Zn ratio observed in well P2 on day 160 is very similar to the S / Zn ratio of the accumulated surface runoff near wells B1 and P2 (see Results); this acidic runoff had collected in pools on the ground surface after snowmelt, and then slowly infiltrated into the subsurface. This suggests that sulfate and Zn have passed unattenuated through the soil profile with the percolating runoff. The S / Zn ratio in B1 is much less than that in the surface water and may reflect the precipitation of jarosite and sulfate adsorption during percolation to the groundwater table. After the sampling episode on day 191, the S / Zn ratio converges to similar values for wells B1 and B4 and indicates that attenuation processes in the deeper horizons of the hillslope, during this period of time, are of the same magnitude for sulfate and Zn. The removal of Fe relative to Zn is apparent in Fig. 9b, with decreasing Fe / Zn ratios with distance from the rock dump. Compared with the S / Zn ratio in well P2 on day 160 (Fig. 9a), the Fe / Zn ratio on day 160 is substantially less than the Fe / Zn ratio in the surface water (i.e. Fe / Zn = 63), indicating retention of iron in the soil profile during the downward percolation of this surface water. The accumulation of Fe in the upper soil horizons has been documented in Herbert (1997b). After the period of snowmelt, the Fe / Zn ratio in well P2 rapidly increases and approaches the Fe / Zn ratio observed in well B1. As indicated in Fig. 9c, the Al / Zn ratio increases with distance from the rock dump (B1 b P2 b B4), indicating that Al is released from the hillslope soils. An increasing Al / Zn ratio could also suggest the removal of Zn from the groundwater, but this interpretation is not supported by the surface complexation modeling for low pH ranges (e.g. in well P2). The Al / Zn ratio in well B1 on day 160 is similar to the Al / Zn ratio in the accumulated surface runoff near B1, suggesting that the Al concentration relative to Zn changes very little during percolation into the soil at the low pH conditions near well B1. This is somewhat surprising, since Al release from aluminosilicate

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211

Fig. 9. Variations in S, Fe, Al, and Cu concentrations relative to changes in Zn concentrations, presented as molar ratios for (a) S / Zn, (b) Fe / Zn, (c) Al / Zn, and (d) Cu / Zn in wells B1, P2, PR and B4. Note different y-axis scales in the upper and lower diagrams. The straight dashed line on each diagram is the component / Zn ratio of the acidic surface runoff near well B1 on day 142.

weathering is expected at the low pH of the infiltrating runoff, such as seen for well P2 where the Al / Zn ratio is greater than that of the surface water. However, the soils have not been investigated near well B1, and it is possible that most aluminosilicates are coated with iron oxyhydroxides in the soil profile near B1, inhibiting mineral weathering. For wells P2 and B4, the Al / Zn ratios approach the same value during the summer (day 204) when the groundwater table is close to its lowest level during the year (Fig. 2); since Al data was not obtained for wells P2 and B4 after this date, it is not possible to draw any conclusions about the coincidence of these Al / Zn ratios. The molar Cu / Zn ratio (Fig. 9d) in groundwater from wells B1 and P2 during the infiltration of the accumulated surface water (up to day 160) is less than the ratio in the surface runoff, which could be the result of Cu accumulation in the soils. In addition, the Cu / Zn ratio exhibits a general decrease with distance from the rock dump, suggesting that Cu is retained in the hillslope relative to Zn. As described above, Cu

accumulation has been observed in the C horizon soils from pit P2 (Fig. 7). 5.3. Organic matter and nitrogen leaching In forested boreal catchments, nitrogen and organic carbon leaching from upper soil horizons is commonly observed in the spring during snowmelt runoff (Laudon et al., 2000, 2001). This leaching is the result of the limited nitrogen uptake and aerobic respiration that occurs during the cold winter months below the snow cover. Decaying organic matter is incompletely mineralized, nitrification proceeds much more slowly, and little inorganic nitrogen is taken up by plants during their winter dormancy; hence, organic carbon and inorganic nitrogen accumulate in the soils during the winter and are subsequently flushed from the soils during snowmelt. Nitrate is generally present in very low concentrations in groundwater from forested catchments since these systems are often nitrogen limited; this is confirmed in the nitrogen leaching that

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is observed in forested catchments after clear-cutting (Nohrstedt et al., 1994). Organic carbon and nitrogen leaching are common processes in forested catchments and the Rudolfsgruvan mine site is no exception, as peaks in TOC and ammonium concentrations are measured in wells B1 and R1 at the end of snowmelt. In well R1, the maximum TOC and nitrate concentrations after snowmelt are less than 2 mg/L (Fig. 6), while ammonium is generally present below detection limits. This suggests that nitrification results in the rapid conversion of ammonium to nitrate in the soils, and that most of this nitrate is taken up by plant growth during the growing season. In contrast, the concentrations of TOC, ammonium, and nitrate are much higher in well B1 than in reference well R1. Ammonium in B1 is detected at concentrations in excess of 1 mg/L throughout the sampling campaign, while maximum nitrate concentrations of ca. 2.7 mg/L are measured after snowmelt. The high TOC and ammonium concentrations suggest that certain microbial processes such as carbon mineralization and nitrification are inhibited in the soils near B1. Nitrification is sensitive to soil pH and is inhibited at low pH (Persson and Wirén, 1995), which would explain the high ammonium concentrations in well B1. The appearance of elevated nitrate concentrations after snowmelt indicates nevertheless that nitrification is occurring, but the continued presence of nitrate in the groundwater throughout the growing season demonstrates that plant uptake must be limited in the vicinity of well B1. 5.4. Migration of solute pulse from the rock dump In general, groundwater pH is high and solute concentrations are low during snowmelt; this is due to the downward percolation of relatively uncontaminated melt water that mixes with the leachate-contaminated groundwater. By the end of snowmelt (days 130–150; Figs. 3 and 4), however, pH has decreased considerably in wells B1 and P2 and metal and TOC concentrations are high. This first high acidity pulse in well B1 (cf. Fig. 5, day 142) is likely the arrival of sulfide oxidation products that have been flushed from the rock dump with the melting snow cover; the oxidation products may have been originally present in the rock dump as water-soluble metal sulfates which then dissolved with the percolation of melt water into the dump. This initial metal-rich pulse from the rock dump mixes with percolating snowmelt containing high TOC and ammonium concentrations. After the arrival of the first pulse to well B1, the pulse further propagates to well P2 and is observed on day 172. The observed decrease in

concentration on day 161 followed by an observed increase in concentration on day 172 in both B1 and P2 may be caused by the arrival of the less acidic surface runoff near B1 and P2 that has percolated to the well screens. This interpretation is supported by the similarity between the S / Zn ratio in P2 and the surface runoff (Fig. 9a). The decrease in groundwater acidity with distance from the rock dump (Fig. 5) is evidence for acid neutralization reactions (i.e. aluminosilicate weathering) occurring in the hillslope. Close to the wetland in well B4, sulfate and Al concentrations exhibit some variation, but Cu, Fe, and Zn concentrations are rather stable (Fig. 4). Because of the distance from the rock dump to well B4, it is not expected that the increase in sulfate concentration on day 172 in B4 (Fig. 4) or the small acidity pulse observed in B4 to occur at the end of snowmelt (day 160; Fig. 5) is related to the solute peak observed in P2 on the same day or to the earlier peak in B1, but rather to the flushing of oxidation products from the soils during this period with a high groundwater table. While groundwater acidity close to the rock dump is primarily derived from the hydrogen ion concentration, acidity near B4 is mostly attributed to the Al3+ concentration, which is relatively constant at ca. 0.25 mM. The resulting groundwater acidity near well B4 is ca. 1 meq/L. This is about double the acidity that has been estimated for the spring flood in a heavily acidified stream the Czech Republic during the 1980s (Laudon et al., 2005). In catchment areas that have not been strongly impacted by anthropogenic acidification, episodic acid and aluminum pulses are observed as a result of natural acidification processes. In northern Sweden, for example, a pH decline in stream water is often observed during the spring flood. The most important factors driving the natural pH decline are the increase in concentration of natural organic acids in combination with dilution of the acid neutralization capacity (Laudon et al., 2000, 2001). In addition, acidity pulses can be attributed to the oxidation of sedimentary pyrite (Brown et al., 1983; Cameron et al., 1998) in the catchment area. For this study, the discharge of acidic groundwater from Ruldolfsgruvan (i.e. near well B4) is likely limited to a small region of the wetland where the fen water is naturally acidic and there is not a direct threat to aquatic organisms. In another setting, however, this acidic discharge could be harmful to aquatic ecosystems. The arrival of the acidity and solute peaks in wells B1 and P2 can be used to roughly estimate the groundwater velocity. The distance between these wells is ca. 30 m, and the elapsed time between peaks is 30 days, resulting in an estimated average velocity (v) of 1 m/d. For an average hydraulic gradient (i) of 0.03 (observed between

R.B. Herbert Jr. / Journal of Geochemical Exploration 90 (2006) 197–214

wells B1 and B2; Fig. 2), a porosity (ϕ) of 0.4, and according to the relation v = (Ks · i / ϕ), this velocity corresponds to a saturated hydraulic conductivity (Ks) of 1.5 × 10− 4 m/s. This is a relatively high hydraulic conductivity for unaltered Swedish glacial till (i.e. podzolic C horizon), which typically has a saturated hydraulic conductivity below 10− 6 m/s (cf. Espeby, 1990; Bishop, 1991; Nyberg, 1995). However, hydraulic conductivities on the order of 10− 4 m/s have been measured close to the ground surface (Espeby, 1990; Bishop, 1991; Nyberg, 1995) where soil-weathering processes result in relatively high hydraulic conductivity. It is therefore likely that most of the solute transport at the relatively high groundwater velocity calculated for this study (i.e. 1 m/d) occurs close to the ground surface. A higher lateral discharge will thus occur in the upper soil horizons during periods after snowmelt when portions of the hillslope are nearly completely water-saturated. The relationship between lateral discharge and metal flux (e.g. mass metal per year), however, is not as obvious since metal concentrations are probably not uniform with depth; the highest metal fluxes may not necessarily coincide with the greatest lateral discharges in the soil profile. In order to quantify total metal fluxes from the rock dump, it would be necessary to acquire depth profiles of metal concentration in the groundwater. 6. Conclusions This study demonstrates that considerable variations in solute concentration and acidity occur in groundwater as a result of the episodic releases of oxidation products from mine waste deposits. These releases are apparently greatest following snowmelt runoff, when sulfide oxidation products are flushed from the rock dump and major nutrients (i.e. nitrate, ammonium) are flushed from soils. The period of snowmelt runoff was associated with a high groundwater table and surface runoff, such that acidity and heavy metals were rapidly mobilized through the upper permeable soil horizons and on the ground surface. While considerable acid neutralization occurs in the glacial till hillslope when the water table has receded, most of this acidity is likely conserved during rapid transport through the upper soil horizons and on the ground surface during snowmelt runoff. The rapid mobilization of sulfide oxidation products to surface water bodies during snowmelt runoff supports the common observation that metal concentrations increase in streams during spring runoff and rainfall events in mining areas (e.g. Paschke et al., 2001; Nagorski et al., 2003; Desbarats and Dirom, 2005) while solute concentrations decrease in streams in uncontam-

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inated forested systems because of dilution that arises from the increase in discharge (cf. Laudon et al., 2000). Soil sampling and geochemical modeling demonstrated that surface complexation is likely the primary process contributing to the removal of SO42−, Cu and Zn from the groundwater. At pH b 4, Cu adsorbs more strongly to iron oxyhydroxide surfaces than Zn, which is indicated by the modeling results and by the greater accumulation of Cu relative to Zn in the C horizon of a contaminated soil profile. Despite the consistently low pH of the groundwater down-gradient from the rock dump (pH b 4.2), the metals Cu, Ni and Zn have accumulated in well-crystallized iron oxyhydroxides at depths greater than 2 m below the ground surface, suggesting that low levels of metal adsorption over long periods of time has led to a relatively significant accumulation of Cu, Ni, and Zn in the subsurface. Acknowledgments Funding for the sampling episodes during 1995 and for various chemical analyses was provided by research grants from Stiftelsen Oscar och Lili Lamms Minne, Uppsala University, and the Royal Swedish Academy of Sciences. Stora Kopparberget AB (Falun) is gratefully acknowledged for providing archive materials on the historical background of the site. Two anonymous reviewers are thanked for their comments that have greatly improved this paper.

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