Geochemical processes in mill tailings deposits: modelling of groundwater composition

Applied Geochemistry 19 (2004) 1–17 www.elsevier.com/locate/apgeochem

Geochemical processes in mill tailings deposits: modelling of groundwater composition S. Ursula Salmona,*, Maria E. Malmstro¨ma,b a

Department of Land and Water Resources Engineering, Royal Institute of Technology (KTH), Brinellv.32, 100 44 Stockholm, Sweden b Department Chemical Engineering and Technology, Div. Industrial Ecology, Royal Institute of Technology (KTH), Osquars Backe 7, 100 44 Stockholm, Sweden Received 10 September 2001; accepted 20 May 2003 Editorial handling by J.S. Herman

Abstract A general model is presented for geochemical processes occurring in the unsaturated zone of a carbonate-depleted, pyritic tailings deposit. Quantification of slow geochemical reactions, using published, empirical rate laws from smallscale experiments on monomineralic samples, and geochemical equilibrium reactions successfully reproduced the relative rates of field processes in the case study, Impoundment 1 in Kristineberg. Reproduction of absolute rates was achieved by scaling down all laboratory-derived mineral weathering rates by two orders of magnitude. The sensitivity of the modelled groundwater composition and pH to rates of pH-buffering processes and redox reactions indicated that inclusion and accurate quantification of all dominant geochemical processes on the field scale is necessary for reliable prediction of groundwater composition and pH. # 2003 Elsevier Ltd. All rights reserved.

1. Introduction As a result of mining over hundreds of years, a large number of sulfidic mill tailings deposits exist around the world. These tailings deposits contain sulfide minerals that can oxidize and thereby cause release of acidic, metal-laden leachates, commonly referred to as acid mine drainage (AMD), over extended periods of time. The timing of the potential onset and duration of AMD generation at a particular site depends on the physical and (bio)geochemical characteristics of the site. These, in turn, determine the rate of, and the balance between, sulfide oxidation and natural attenuation of contaminants, such as acidity and metal and metalloid ions, within the deposit. Ability to understand and to predict the time evolution of drainage quality, such as is needed for assessment of the environmental impact of tailings deposits and evaluation of various remediation * Corresponding author. Fax: +46-8-790-8689. E-mail address: [email protected] (S.U. Salmon).

options, thus requires quantification of a number of physical and (bio)geochemical processes occurring in the deposit, and understanding of their complex feedback mechanisms. Many of the previous modelling studies of AMD from tailings deposits focus on the oxidant supply through O2 diffusion in combination with a simplified representation of the geochemistry (e.g., Jaynes et al., 1984; Elberling et al., 1994; Wunderly et al., 1996; Werner and Berglund, 1999; Bain et al., 2000). These models provide valuable insight into the coupling of different physical processes and oxidation of iron sulfide minerals, as well as feedback mechanisms. However, with the exception of iron sulfide oxidation by O2, geochemical processes that are slow and limited by (bio)geochemical kinetics are either negelected or quantified through simplifying mass-balance or local equilibrium. In sulfidic mill tailings, such slow reactions include sulfide oxidation by Fe(III), oxidation of dissolved Fe(II), and aluminosilicate dissolution. As these reactions are associated with iron and acidity attenuation within the

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S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

deposit, they may have a large influence on groundwater pH and/or composition. Reliable interpretative and predictive modelling of the quality of leachate from tailings deposits thus potentially requires detailed assessment of reaction kinetics for such reactions. Scharer et al. (1994), Lichtner (1996) and Mayer et al. (1999, 2000, 2003) consider geochemically limited rates of aluminosilicate weathering and/or Fe(II) oxidation in tailings deposits. These studies show that geochemical limitation on process rates can indeed be important for the composition of AMD, however, a limited number of processes are quantified in the models and the relative importance of different (bio)geochemical reactions are not assessed. Thus, whereas previous models have yielded extensive insight into the coupling and feedback between different physical processes and iron sulfide mineral oxidation and indicate the importance of (bio)geochemical limitations on process rates, a more comprehensive model assessement of the (bio)geochemical processes and their feedback mechanisms is currently not available in the literature. In this communication, the authors focus on geochemical processes that contribute to concentrations of major ions in the groundwater and that determine the proton balance within pyritic tailings deposits with negligible carbonate mineral content. Building from first principles, all geochemical processes are quantified with kinetic rate laws or equilibrium expressions. Specific objectives are to: (1) develop conceptual and mathematical models for the geochemical processes occurring in the unsaturated zone of a carbonate-depleted, pyritic tailings deposit; (2) investigate the relative importance of geochemical processes contributing to the groundwater composition in a tailings deposit; and

(3) test the sensitivity of the modelled groundwater composition and process dominance to uncertainties in selected key parameters and processes. In order to provide for generality, the authors quantify the slow, kinetically controlled reactions using published, empirical rate laws determined in small-scale, abiotic, laboratory experiments. Comparison of model results to the field-scale case study, a mill tailings deposit at the Kristineberg site in northern Sweden, enables assessment of model performance. The commonly observed orders-of-magnitude scale-dependence of process rates (such as mineral dissolution rates, e.g., White and Peterson, 1990; Malmstro¨m et al., 2000; and references therein) and capacities (such as sorption capacity, e.g., Brown et al., 1998), for instance, implies that such comparisons are necessary to indicate whether models that build on independent, small-scale observations are relevant for field-scale processes. This is a prerequisite for making further assessment of model implications meaningful and for predictive use of models.

2. Overview of geochemical processes in mill tailings deposits Mill tailings from base metal sulfidic ores consist of fine grained material containing economically undesirable minerals, such as iron sulfides, carbonates and aluminosilicates, as well as lesser amounts of ore minerals, such as chalcopyrite and sphalerite. Oxidation of iron sulfides, such as pyrite (FeS2) and pyrrhotite (Fe1
xS), by O2 (aq) is a key process, releasing protons, SO4, and Fe(II) (Table 1, Reaction 1; see review by Evangelou and Zhang, 1995). Oxidation of the released Fe(II) produces Fe(III) (Reaction 10 in Table 1), which may precipitate with release of protons, e.g.:

Table 1 Slow geochemical processes used in the modelling of the unsaturated zone of Impoundment 1 in this study Process and reaction stoichiometry Pyrite oxidation (oxygen path): FeS2 ðsÞ þ H2 O þ 72 O2 ðaqÞ ! Fe2þ þ 2SO42
þ 2H þ Pyrite oxidation (ferric iron path): FeS2 ðsÞ þ 14Fe3þ þ 8H2 O ! 15Fe2þ þ 2SO42
þ 16H þ Chalcopyrite oxidation (oxygen path): CuFeS2 ðsÞ þ 4O2 ðaqÞ ! Fe2þ þ Cu2þ þ 2SO42
Chalcopyrite oxidation (ferric iron path): CuFeS2 ðsÞ þ 16Fe3þ þ 8H2 O ! 17Fe2þ þ Cu2þ þ 2SO42
þ 16H þ Sphalerite oxidation (oxygen path): ZnSðsÞ þ 2O2 ðaqÞ ! Zn2þ þ SO42
Sphalerite oxidation (ferric iron path): ZnSðsÞ þ 8Fe3þ þ 4H2 O ! Zn2þ þ SO42
þ 8Fe2þ þ 8H þ a Chlorite (chlinochlore) weathering:
 III þ Mg4:5 FeII Fe Al AlSi ! 4:5Mg2þ þ 0:2Fe2þ þ 0:2Fe3þ þ 2Al3þ þ 3SiO2 ðsÞ þ 12H2 O 3 O10 ðOHÞ8 ðsÞ þ 16H 0:2 0:2 a 8. Muscovite weathering: K0:8 Na0:2 ðFe0:1 Al1:9 ÞAlSi3 O10 ðOHÞ2 ðsÞ þ 10H þ ! 0:8K þ þ 0:2Naþ þ 0:1Fe3þ þ 2:9Al3þ þ 3SiO2 ðsÞ þ 6H2 O 9. Plagioclaseb weathering: Na0:75 Ca0:25 Al1:25 Si2:75 O8 ðsÞ þ 5H þ ! 0:75Naþ þ 0:25Ca2þ þ 1:25Al3þ þ 2:75SiO2 ðsÞ þ 2:5H2 O 10. Ferrous iron oxidation: Fe2þ þ 14 O2 ðaqÞ þ H þ ! Fe3þ þ 12 H2 O 1. 2. 3. 4. 5. 6. 7.

a b

Composition based on reported mineralogy of aluminosilicates in the Kristineberg mine (du Rietz, 1953). Oligoclase; use of this plagioclase composition resulted in a solid phase composition that most closely resembled that in the field.

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

Fe3þ þ 3H2 O $ FeðOHÞ3 ðsÞ þ 3Hþ . However, abiotic oxidation of dissolved Fe(II) is slow at low pH. With increasing importance of hydrolysis species of Fe(II), and faster oxidation of these species than the Fe2+ specie, the overall oxidation rate increases with increasing pH (e.g., Wehrli, 1990). The produced dissolved Fe(III) is a powerful oxidant that may oxidize, for example, pyrite (Reaction 2 in Table 1; see Wiersma and Rimstidt, 1984; McKibben and Barnes, 1986; Moses and Herman, 1991). Through Reactions 2 and 10 in Table 1, dissolved iron may, under favourable conditions, hence undergo internal redox cycling in the presence of dissolved molecular O2 and sulfide minerals. Release of hazardous heavy metals and metalloids occurs through oxidation of other, less abundant sulfide minerals, such as sphalerite, ZnS(s), and chalcopyrite, CuFeS2(s), or dissolution of iron sulfide minerals containing trace amounts of such elements. As for the iron sulfides, oxidation of minor sulfides may occur through the dissolved O2 and/or the Fe(III) reaction path (Reactions 3–6 in Table 1; see Rimstidt et al., 1994; Scharer et al., 1994). At low pH, dissolution of monosulfide minerals, such as sphalerite and pyrrhotite, may also occur without oxidation (e.g., Janzen et al., 2000). The rates of the above mentioned redox reactions (Reactions 1–6 and 10 in Table 1) may all be greatly increased by the action of S or iron oxidising bacteria (see review by Nordstrom and Southam, 1997, and short summary in Section 3.4.4). Natural attenuation of released acidity within the tailings occurs through dissolution of carbonate minerals, e.g., calcite: CaCO3(s)+H+$ Ca2++HCO
3, if available (see Blowes and Ptacek, 1994), as well as weathering of aluminosilicate minerals (see Table 1, Reactions 7–9; Stro¨mberg and Banwart, 1999; Banwart and Malmstro¨m, 2001; Salmon and Malmstro¨m, 2001). Secondary processes occurring in tailings deposits, including the relatively fast precipitation and dissolution of secondary phases, such as hydroxide (Blowes and Ptacek, 1994) and sulfate minerals, may also control and/or be affected by the pH and groundwater composition (see review by Nordstro¨m and Alpers, 1999).

3. Model development 3.1. Conceptual model It is generally accepted that diffusion of O2 through the pore space in the unsaturated zone is the dominant form of O2 transport into tailings deposits (see Elberling and Nicholson, 1996, and references therein). As the effective diffusion coefficient of O2 in porous media is highly dependent upon water saturation, with several orders of magnitude lower diffusivity at full water

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saturation than at low water saturation (Nicholson et al., 1989), O2 is more readily available above the groundwater level. As dissolved molecular O2 is an important oxidant for sulfides, sulfide oxidation is expected to occur predominantly in the unsaturated part of the deposit. Dissolution of minerals that buffer pH are also expected to occur predominantly in the unsaturated part of the deposit, as these weathering reactions are expected to be more intense close to the proton source, where pH is lower. In this modelling study, the unsaturated zone is thus considered to be the major reactive zone in tailings deposits and the dominant source of dissolved constituents in the groundwater. For sulfide minerals, this is supported by depletion above the groundwater level, as is commonly observed in tailings deposits, including Impoundment 1 in the case study. 3.2. Case study: mill tailings deposit ‘‘Impoundment 1’’ in Kristineberg, northern Sweden The case study used in this modelling study is a mill tailings deposit called ‘‘Impoundment 1’’ in Kristineberg, northern Sweden. The authors considered preremediation conditions, for which field data from sampling campaigns in the 1980s and 1990s have been compiled by Malmstro¨m et al. (2001). A detailed site description is beyond the scope of this study, which was carried out as a desk-top study, using the previously published site characteristics; a short introduction to the main features of the deposit prior to remediation is given in this section and further details, including field data acquisition, are found in Appendix. Table 2 lists the main characteristics and parameter values that were used in this modelling study. 3.2.1. Physical and hydrological characteristics The Kristineberg mine site is located in the Skellefte district 175 km SW of Lulea˚, Sweden. The average annual temperature at the site is 1  C. The tailings deposit called Impoundment 1 covers an area of approximately 0.11 km2 and is on average 5 m deep. The extent of the unsaturated zone, that is, the average depth to the groundwater, was 1 m prior to remediation. The total water flow through the deposit, consisting of effective infiltration and recharge from moraine slopes surrounding the deposit, was 47 000 m3 a
1. Fig. 1 summarises the important flows contributing to the water balance of the deposit, along with the position of the groundwater table and the water sampling plane. The tailings consist of a finely ground material (characterised as silty fine sand or fine sand). The specific surface area of the tailings was found to be in the range 0.2–10.1 m2 g
1, with an average value of 2.96 m2 g
1 (detemined in this study; see Appendix for further details).

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S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

Table 2 Parameter values used in the model of the unsaturated zone of Impoundment 1a Average volumetric fractions (g) of minerals Sulfidesb

Silicatesc 

Pyrite Pyrrhotite Chalcopyrite Sphalerite

0.04 0.002 0.001

Chlorite Talc Muscovite Plagioclase Quartz

 0.45 0.15 0.10 0.25

Parameters

h q v c  As,tot

Parameter

Value

Units

Unsaturated zone depth Water flowrate Porosityd Compact densitye Water content Tailings surface area

1 1.35 10
8 0.25 3.3 106 0.15 2.96

m m s
1 m3 m
3 g m
3 m3 m
3 m2 g

Source 9 = ; 

Axelsson et al. (1986) Malmstro¨m et al. (2001)

see Appendix

a

Changes since a preliminary model assessment of Impoundment 1 (Salmon, 2000; Salmon and Malmstro¨m, 2000, 2002): i) The 5pt BET surface area is used, see Appendix; ii) feldspars (plagioclase) and pyrrhotite are included, after a detailed post-remediation mineralogical study (Holmstro¨m et al., 2001), and iii) recharge from adjoining till slopes, in addition to the effective infiltration, is considered in the water flowrate. b Based on Qvarfort (1983), Axelsson et al. (1991), and Holmstro¨m et al. (2001). c Based on Qvarfort (1983) and Holmstro¨m et al. (2001). d Estimated from literature data by Axelsson et al. (1986). e Density, =c(1–v).

3.2.2. Mineralogical and chemical characteristics of the tailings The tailings within the deposit are pyritic and carbonate-depleted; the main minerals are chlorite, talc, muscovite, plagioclase, quartz and pyrite. Sphalerite, chalcopyrite and pyrrhotite are also present. The tailings in the upper part of the deposit, on average down to the groundwater table, had been partly depleted in S, Fe, Zn and Cu, suggesting weathering of sulfides in the unsaturated zone. 3.2.3. Groundwater quality The groundwater, which had been sampled 1.5 m below the groundwater table, had an average pH of 4.9; redox potentials are not available in the literature. The groundwater contained high concentrations of dissolved 2+ SO2
, Zn2+, Ca2+ and Al3+, as well as 4 , Fe, Mg lower concentrations of Na+, Cu2+ and K+. Saturation index calculations (this study, see Appendix) indicated that the groundwater was close to saturation with respect to gypsum (CaSO4.2H2O(s)), amorphous silica (SiO2(s)), and an amorphous ferric hydroxide phase (Fe(OH)3(s)). 3.2.4. Site specific conceptual model For Impoundment 1, the partial depletion of sulfides in the unsaturated zone and the limited variation in

groundwater composition with depth, given a small vertical component of water flow in the deposit, supports the hypothesis that the unsaturated zone was the major source of dissolved species and thus reactions occurring in the saturated zone were less important for the groundwater quality. Given the low spatial and temporal resolution of the available field data, the

Fig. 1. Schematic illustration of the hydrological situation in the case study, ‘‘Impoundment 1’’. The upper marked plane shows the average position of the water table, as observed by Axelsson et al. (1986) in 16 piezometers during 2 a (see Appendix). The lower marked plane indicates the plane, 1.5 m below the groundwater surface, from which aqueous samples for the published groundwater quality data (see Fig. 3) used in this study had been collected.

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

authors, as a first approximation, modelled the groundwater composition by assessing the unsaturated zone only; this zone was conceptualised as a single, completely-mixed flow-through reactor (box). The modelled leachate leaving the unsaturated zone was compared with the available data for groundwater quality from 1.5 m below the groundwater level (see Fig. 1). Solute transport is considered to occur via advection; as model results were compared with concentrations from 1.5 m below the groundwater level, the average total flow of water through the deposit (47 000 m3 a
1, see Appendix) was considered in the model. In this modelling study, the yearly average of the total water flowrate through the deposit was used and all physico-chemical properties, such as the water content, O2 concentration, and mineral abundance, were assumed constant over the modelled zone and in time. This approach is consistent with the limited spatial and temporal resolution of available site data for Impoundment 1 (Malmstro¨m et al., 2001) and for mine waste sites in general (cf. Banwart and Malmstro¨m, 2001). The geochemistry was furthermore assumed to be at quasisteady state (cf. Furrer et al., 1989). 3.3. Geochemical processes The site-specific conceptual model for the geochemical processes occurring in Impoundment 1 is shown in Fig. 2. Based on the mineral abundance (Table 2) and the composition of the aqueous phase, the main primary sources of groundwater solutes were considered to be the slow oxidation of sulfide minerals (pyrite, sphalerite and chalcopyrite) by O2(aq) (Reactions 1, 3 and 5 in Table 1) and Fe(III) (Reactions 2, 4 and 6 in Table 1)

Fig. 2. Conceptual model for geochemical processes in the unsaturated zone of a mill tailings deposit. Kinetically controlled processes (single headed arrows) include weathering of aluminosilicate minerals (chlorite, muscovite, and plagioclase), oxidation of sulfide minerals (pyrite, sphalerite, and chalcopyrite) by both dissolved molecular O2 and Fe(III), and oxidation of aqueous Fe(II). Fast, equilibrium controlled processes (double headed arrows) include aqueous speciation, Henry’s law equilibrium between the aqueous solution and pore gases, and dissolution/precipitation of secondary minerals.

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and the slow weathering of aluminosilicates (chlorite, muscovite and plagioclase; Reactions 7–9 in Table 1). Although Fe(III) is the thermodynamically favoured iron redox species in the presence of O2(aq), the high iron concentrations in the groundwater suggested dominance of Fe(II), as Fe(III) is highly insoluble at the pH in the groundwater (4.9). This implied a kinetic limitation on the oxidation of Fe(II) to Fe(III). Aqueous Fe(II) oxidation by O2 (aq) was thus conceptualised as a relatively slow, kinetic process (cf. Stro¨mberg and Banwart, 1994). Relatively fast, reversible geochemical processes included aqueous speciation, solubility equilibrium between the aqueous solution and secondary mineral phases (gypsum, ferrihydrite and amorphous silicate; see solid phases in Table 3), and Henry’s law equilibrium between aqueous solution and pore gases in the unsaturated zone (see gas phase in Table 3). 3.4. Process quantification 3.4.1. Oxygen availability Diffusion of O2 into the unsaturated zone has in a number of previous tailings leachate modelling studies been assumed to be the limiting factor in oxidation rates (e.g., Wunderly et al., 1996) and the intrinsic surface reaction rate has been neglected. In this modelling study, where the focus is on the interactions between different geochemical processes, the authors investigated the impact of both full and limited O2 availability on the modelled groundwater composition. In the model, the aqueous concentration of O2 was assumed to be at equilibrium with the O2 partial pressure, PO2, in the pore space in the box. Each simulation was performed with a fixed partial pressure of O2, with the steady-state condition defining the associated total O2 flux into the deposit (FO2). Surface reaction control of the overall O2 flux was simulated by fixing the PO2 to the atmospheric partial pressure (0.2 atm); this resulted in the maximum rates of oxidation, and, hence, the maximum FO2, for each set of parameter values tested. Mixed or diffusion control of the overall O2 consumption rate in tailings deposits results in lower values of PO2, due to limitation on FO2; this was simulated through fixing the PO2 in the modelled zone at lower levels. It is noted that the use of an alternative approach, as applied by Brown et al. (2000), in which FO2 is fixed and the values of PO2 and rates of oxidation change to satisfy mass balance considerations, resulted, with otherwise identical parameter values, in identical modelled groundwater composition and PO2. 3.4.2. Equilibrium processes Compared to the average residence time of water in the unsaturated zone, the geochemical processes in the conceptual model occur on different time-scales. The authors treated reversible geochemical processes that

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S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

Table 3 Equilibrium model and equilibrium constants used in the modelling of the unsaturated zone of Impoundment 1 in this study Reaction

Log K(T,I)a

are generally much faster than the residence time of water by assuming chemical equilibrium and quantifying the processes through the use of mass-action laws: Y ð1Þ CðiÞ ¼ KðiÞ CðjÞaði;jÞ j

Aqueous phase H2O
H+ () OH

H++SO2
4 () HSO4
Na++SO2
( ) NaSO 4 4
K++SO2
4 () KSO4 Mg2++H2O
H+ () MgOH+ Mg2++SO2
4 () MgSO4(aq) Ca2++H2O
H+ () CaOH+ Ca2++SO2
4 () CaSO4(aq) + Ca2++H++SO2
4 () CaHSO4 Cu2++H2O
H+ () CuOH+ Cu2++2H2O
2H+ () Cu(OH)2(aq) Cu2++3H2O
3H+ () Cu(OH)
3 Cu2++SO2
4 () CuSO4(aq) Zn2++H2O
H+ () ZnOH+ Zn2++2H2O
2H+ () Zn(OH)2(aq) Zn2++SO2
4 () ZnSO4(aq) 2
Zn2++2SO2
4 () Zn(SO4)2 Fe2++H2O
H+ () FeOH+ Fe2++2H2O
2H+ () Fe(OH)2(aq) Fe2++SO2
4 () FeSO4(aq) + Fe2++H++SO2
4 () FeHSO4 Fe3++H2O
H+ () FeOH2+ Fe3++2H2O
2H+ () Fe(OH)+ 2 Fe3++3H2O
3H+ () Fe(OH)3(aq) Fe3++4H2O
4H+ () Fe(OH)
4 + Fe3++SO2
4 () FeSO4 2+ Fe3++H++SO2
4 () FeHSO4
Fe3++2SO2
4 () Fe(SO4)2 Al3++H2O
H+ () AlOH2+ Al3++2H2O
2H+ () Al(OH)+ 2 Al3++3H2O
3H+ () Al(OH)3(aq) Al3++4H2O
4H+ () Al(OH)
4 + Al3++SO2
4 () AlSO4 2+ Al3++H++SO2
4 () AlHSO4
Al3++2SO2
( ) Al(SO ) 4 4 2 2H++CO2
3 () H2CO3(aq)
H++CO2
3 () HCO3 + Ca2++H++CO2
3 () CaHCO3 2
2+ + Mg +H +CO3 () MgHCO+ 3 + Fe2++H++CO2
3 () FeHCO3 Solid phase Fe(OH)3(s) () Fe3++3H2O
3H+ CaSO4.2H2O(s) () Ca2++SO2
4 +2H2O SiO2(s) () H4SiO4(aq)
2H2O Gas phase O2(g) () O2(aq) CO2(g)() CO2(aq)


14.6b 1.16b,c 0.05b
d 0.13b
d
11.7b,c 0.92b,c
13.1b,c 1.04b,c 1.66c
8.3c,d
14.0c,d
26.9c,d
1.07c,d
10.1c,d
17.2c,d 1.13c,d 2.12c,d
10.6b
d
22.7c,d 0.89b
d 1.66b,c
3.44b,c
7.64b
d
15.0b,c
24.2b
d 2.05b,c 2.77b,c 2.77b,c
6.32b,c
12.7b
d
20.3b,c
26.0b,c 1.15b,d 0.75b,c 2.42b,d 15.96b 9.98b,c 10.33d 10.42b,c 11.4b,c 5.76d,e
3.41b,c,f
2.92b,c,g
2.84c
17.32b

a Equilibrium constants (K) have been corrected, where Hr values were available, for 1  C using the Van’t Hoff equation, and for ionic strength (I =0.27) with the Davies equation. Constants were taken from, in order of preference: b Nordstrom et al. (1990), c Ball and Nordstrom (1991), d Allison et al. (1991), e Ferrihydrite, f Gypsum, g Amorphous silica.

where CðiÞ is the free concentration of species i (mol dm
3), KðiÞ is the conditional stability constant of species i, and aði; jÞ is the stoichiometric coefficient of component j in species i. The equilibrium model is given in Table 3, with constants corrected for field temperature and ionic strength. The choice of components reflected the major elements reported for the groundwater in the deposit. Redox speciation of dissolved iron was considered by defining both Fe(II) and Fe(III) as components, thereby allowing disequilibrium between all redox couples in the model. Mass transfer between the two redox states of iron was controlled by the rates of the involved redox reactions (see Table 1; see also Section 3.3). As none of the major processes in the conceptual model produced or consumed CO2, the concentration of CO2(aq) was assumed to be in equilibrium with the partial pressure of CO2 in the atmosphere (3 10
4 atm). 3.4.3. Abiotic kinetic processes Relatively slow or irreversible processes (Table 1) were conceptualised as kinetic processes and quantified through empirical rate laws: Y rl ¼ kl CðiÞnðl;iÞ ð2Þ i

where rl is the rate of process l normalised to the mineral surface area for heterogeneous reactions, or to the aqueous solution volume for homogeneous reactions; kl is the rate constant and nðl; iÞ is the reaction order of chemical species i in the rate law for process l. For this study, empirical rate laws were selected from laboratory scale, abiotic weathering experiments of pure mineral samples reported in the literature (see brief review in Salmon and Malmstro¨m, 2002). This implies that the basis for quantification of all kinetic processes was independent of site observations. Through consideration of the site specific properties of the unsaturated zone, that is, the available mineral surface area (for heterogeneous reactions) or the water volume, as calculated from the water content (for homogenous reactions), the rate laws were converted to rate expressions, R (mol dm
2 s
1), normalised to the surface area of the deposit: Y Rl ¼ rl PðmÞ ð3Þ m

where PðmÞ is the value of parameter m. Table 4 lists the final rate expressions, together with associated rate constants that have been adjusted for field temperature,

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S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

using published activation energies. Other parameter values for the rate expressions in Table 4 are given in Table 2. Rate laws were not found in the literature for oxidation of chalcopyrite and sphalerite by dissolved molecular O2; analogy was thus made with the rate law for oxidation of pyrite and rate constants were calculated from reported (Scharer et al., 1994) abiotic rates of oxidation of these minerals by O2(aq). 3.4.4. Microbially mediated kinetic processes The rate of pyrite oxidation is reported to be up to a factor 35 greater in microbiological laboratory experiments than in abiotic controls (Olson, 1991; Fowler et al., 2001; Yu et al., 2001; see also review in Nordstrom and Southam, 1997), implying that abiotic rate laws may greatly underestimate the rate of this process. Microbial mediation has also been reported to accelerate sulfide oxidation on the field or near field scale, though only by a factor of 1.5–5 (Stro¨mberg and Banwart, 1999; Edwards et al., 2000; Elberling et al., 2000).

Discussions in the literature (Singer and Stumm, 1970; Sand et al., 2001; Fowler et al., 2001; Yu et al., 2001) indicate that the dominant mechanism for microbial acceleration of sulfide mineral oxidation may be via acceleration of oxidation of aqueous Fe(II), followed by oxidation of the mineral by Fe(III) (Reactions 10 and 2 in Table 1), rather than via the O2 pathway (Reaction 1 in Table 1). The rate of microbially mediated Fe(II) oxidation has been found to exceed the abiotic rate by up to 4–6 orders of magnitude in laboratory experiments (Lacey and Lawson, 1970; Singer and Stumm, 1970; Nemati and Webb, 1997; see also reviews by Nordstrom and Southam, 1997 and Nemati et al., 1998); a similar magnitude of acceleration has also been seen in the field (Kirby and Elder Brady, 1998, and references therein). Abiotic acceleration of Fe(II) oxidation subsequent to sorption onto surfaces of e.g. ferric oxy(hydr)oxides (Tamura et al., 1976; Wehrli, 1990) is also described in the literature. Although acceleration of these redox reactions is thus extensively reported in the literature, the rates of the

Table 4 Model quantification of the slow geochemical processes in Table 1 Processa

Rate expressionb

Activation energyc (kJ mol
1)

Parameterd

Valuee

(1)

Rpyo ¼ Xr hApy kpyo ½O2 ðaqÞ0:5 ½Hþ 
0:11  0:62 Rpyf ¼ Xr hApy kpyf Fe3þ

56.9

kpyo

8.7 10
12

mol0.61 dm
0.83 s
1

92

kpyf


10

8.1 10

mol0.38 dm
0.14 s
1

20

kcpo

4.0 10
13

mol0.5 dm
0.5 s
1

63

kcpf

1.2 10
11

mol0.57 dm
0.71 s
1

kspo


14

mol0.5 dm
0.5 s
1


10

(2) (3) (4) (5)

0:5

Rcpo ¼ Xr hAcp kcpo ½O2 ðaqÞ  0:43 Rcpf ¼ Xr hAcp kcpf Fe3þ

0:5

(9)

Rspo ¼ Xr hAsp kspo ½O2 ðaqÞ  0:58 Rspf ¼ Xr hAsp kspf Fe3þ

 Rch ¼ Xr hAch kch1 ½Hþ 0:50 þkch2

 Rmu ¼ Xr hAmu kmu1 ½Hþ 0:40 þkmu2

 0:45 Rpl ¼ Xr hApl kpl1 ½H þ  þkpl2

(10)


   Rfe ¼ h½O2 ðaqÞ kfe1 Fe2þ þ kfe2 ½FeOH þ  þ kfe3 FeðOHÞ2

Inflowf:

Rin ¼ qCin

(6) (7) (8)

a

Outflowg:Rout ¼ qCout

21

Units

5.4 10

27

kspf

3.9 10

mol0.42 dm
0.26 s
1

30

kch1 kch2 kmu1 kmu2 kpl1 kpl2 kfe1 kfe2 kfe3 Xr

2.6 10
13 4.6 10
15 2.9 10
14 3.6 10
16 6.1 10
14 6.5 10
16 6.8 10
6 2.1 101 6.8 106 0.02

mol0.5 dm
0.5 s
1 mol dm
2 s
1 mol0.6 dm
0.8 s
1 mol dm
2 s
1 mol0.55 dm
0.65 s
1 mol dm
2 s
1 mol
1 dm4 m
1 s
1 mol
1 dm4 m
1 s
1 mol
1 dm4 m
1 s
1 (see Section 4.1)

22 80.3 70

Numbers correspond to reactions in Table 1. Rate given in units of mol dm
2 s
1; concentration in mol l
1. References: (1) Williamson and Rimstidt (1994); (2) Rimstidt and Newcomb (1993; with O2 present); (3–5) modified from Scharer et al. (1994); (4,6) Rimstidt et al. (1994); (7) derived from data in Malmstro¨m et al. (1995); (8) Knauss and Wolery (1989); (9) Oxburgh et al. (1994); (10) Wehrli (1990). c References: (1) McKibben and Barnes (1986); (2) Wiersma and Rimstidt (1984); (3,5) Scharer et al. (1994); (4,6) Rimstidt et al.(1994); (7) Analogy with biotite, White and Brantley (1995); (8) kmu1 from Nagy (1995), kmu2 by analogy with Nagy (1995); (9) Sverdrup (1990); (10) Average of relevant values in Lowson (1982). d Other parameters given in Table 2. e Rate constants have been corrected for temperature of 1  C using the Arrhenius equation and activation energies. f The inflowing water was assumed to be pure rainwater containing some sulfuric acid acidity, such that CH2 SO4 =1 10
5 M, and concentrations of other components were zero. g Cout is the concentration in outflow (groundwater composition). b

8

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

alternative reactions paths have seldom if at all been systematically quantified in a way that is useful for modelling (Edwards et al., 2000). The few rate laws that are available (e.g. Scharer et al., 1994; Pesic et al., 1989) often require site-specific information that generally is not available for mill tailings deposits, and furthermore, there is no consensus in the literature as to the best rate laws (Kirby et al., 1999). In this study, the authors therefore used abiotic rate laws, and tested the impact of enhanced rates of redox reactions, as due to, for example, microbial mediation, in a sensitivity analysis (see Section 6). 3.4.5. Mineral surface area The specific mineral surface area, to be used in conjunction with surface area normalised empirical rate laws in quantification of weathering rates (see Table 4), was for the ith mineral estimated by allocation of a fraction of the total specific surface area of the tailings (As,tot) in proportion to the volumetric fraction,  i, of the mineral: Ai ¼ i As;tot

ð4Þ

This modelling approach is based on the assumptions that weathering rate laws for monomineralic samples can be applied to mixtures of minerals, such as in tailings, and that the specific mineral surface area in the mixture is proportional to the volumetric fraction of the mineral. Preliminary laboratory investigations of these assumptions, using batch weathering experiments on tailings from Impoundment 1, have indicated that they are reasonable first-hand approximations (Salmon and Malmstro¨m, 2001). Volumetric fractions ( i) for the case study are given in Table 2. The weathering of magnesium silicates (chlorite and talc) and iron sulfides (pyrite and pyrrhotite) was for the purpose of modelling represented by weathering of chlorite and pyrite, respectively. This was due to the lack of specification of individual volumetric fractions for the minerals in the reported field data, that the same weathering tracers were to be used for both minerals in each pair, and that no weathering rate laws are available in the literature for talc or pyrrhotite. A reported weathering rate for talc (White and Brantley, 1995) indicates that this approach overestimated the Mg2+ release rate by at most 35%. Similarly, based on a reported rate for pyrrhotite oxidation by dissolved molecular O2 (Janzen et al., 2000; PO2=0.2 atm), this approach was expected to underestimate aqueous SO2
4 and iron concentrations by at most 20 and 40%, respectively, and overestimate pH by at most 0.2 pH units. 3.5. Model implementation The mathematical model was implemented through STEADYQL (Furrer et al., 1989), which allows a numerical, geochemical quasi-steady state box modelling approach that previously has been successfully applied to waste rock dumps and underground mines

(Stro¨mberg and Banwart, 1994; Brown and Lowson, 1997; Brown et al., 2000). For any component, the steady state balance between chemical processes and the outflow is achieved through: " # X X RðlÞsðl; jÞ
q aði; jÞCðiÞ ! 0 ð5Þ l0

i

where l0 represents all processes but the outflow, s(l,j) denotes the stoichiometric coefficient of component j in process l, and q is the water flow rate (dm s
1); the second term in Eq. (5) represents the outflow of components from the box. Processes, such as depletion and accumulation of minerals, that are much slower than the residence time of water can be considered through successive steady state simulations (eg., Salmon, 2000).

4. Model results 4.1. Groundwater composition Fig. 3 shows a comparison of modelled concentrations of components (Po2=0.2 atm; open symbols) and the composition of the groundwater in Impoundment 1 at 1.5 m depth below the groundwater table. Concentrations of components released by aluminosilicate weathering (Mg2+, Al3+, Na+, K+, and Ca2+) were approximately two orders of magnitude greater in the model than in the field. The exception to this was the concentration of Ca2+, which in the model was affected by solubility equilibrium between gypsum and the aqueous phase, and which was similar between the model and the field. The modelled pH and concentrations of com2+ ponents released by sulfide weathering (Fe, SO2
4 , Cu 2+ and Zn ) were sensitive to the availability of O2, as determined by the chosen value of PO2 for a given simulation. With lower values of PO2, the modelled pH increased above the value observed in the field while iron and SO2
4 concentrations were still greater than in the field; hence, no value of PO2 could simultaneously reproduce the iron and SO2
4 concentrations and pH observed in the field. For conditions of full O2 availability (PO2=0.2 atm), the modelled pH was close to the value of 4.9 reported for the field; however, modelled concentrations of most major components were approximately two orders of magnitude greater than in the field (Fig. 3). This is consistent with the commonly observed scale-dependence of mineral weathering rates (White and Peterson, 1990; see also Malmstro¨m et al., 2000, and references therein). In order to allow for this in the model, a factor, Xr (Xr <1) was applied to all weathering rate expressions (Expressions 1–9 in Table 4). In the quantification of the magnitude of Xr, the concentration of Mg2+ was used as a tracer for the field weathering rates, as Mg2+ originates from the dominant aluminosilicates, the weathering

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

rates of which do not depend (directly) upon the dissolved molecular oxygen concentration. A value of Xr in the order of 10
2 resulted in a Mg2+ concentration corresponding to that reported for the field. For all major components, allowance for scaledependence of mineral weathering rates in the model, through application of the same factor (Xr=0.02) to all weathering rates, resulted in modelled concentrations that were close to those reported for the field (filled symbols in Fig. 3). At the same time, the modelled pH was still close to the field value. Remaining discrepancies between results of the calibrated model and field observations are considered small, given the uncertainty in, e.g., the mineral content in the unsaturated zone, which is associated with analysis of solid phase from a small number of observation points, and the specific surface area of the tailings, as demonstrated by the large variability in the value of this parameter, as well as in the water flow and the depth of the unsaturated zone. The only exception to the good agreement between results of the calibrated model and field observation was the Zn2+ concentration, where the modelled concentration was approximately 3 orders of magnitude lower than the field concentration. Possible reasons for this include that the assumption of an analogous form of rate laws for sphalerite and pyrite weathering (Section 3.4.3), in the absence of a rate law for sphalerite oxidation by O2(aq) in the literature, is inappropriate. Alternatively, other sources of dissolved Zn2+ than sphalerite oxidation, which are not accounted for in the model, may be important in the field. 4.2. Dominant processes The calibrated model was analysed for dominant processes that contributed to the proton balance, O2

9

consumption, redox cycling of Fe, and composition of the groundwater in the deposit (Fig. 4). For full O2 availability (PO2=0.2 atm; filled bars in Fig. 4), the major processes contributing to the proton balance were pyrite oxidation by dissolved molecular O2 (proton release) and chlorite dissolution (proton consumption). This explains why the pH was not greatly affected by application of Xr (see Fig. 3); the rates of the reactions dominating proton production and consumption were scaled by the same factor. As well as being the major source of protons, oxidation of pyrite by dissolved molecular O2 was found to be the major source of Fe(II) and SO2
4 and the major sink of dissolved molecular O2 (Fig. 4). Gypsum dissolution was the source of approximately one quarter of the released SO2
4 . Pyrite oxidation by Fe(III) was relatively low (Fig. 4), due to low solubility of Fe3+ at pH  5. Ferrous iron oxidation was also low; Fe(II) was found to be the dominant redox form of Fe, with a Fe(II) to Fe(III) ratio of  105 in solution, and a modelled pe of approximately 7. As well as being a major sink for protons (Fig. 4), chlorite dissolution was also the source of Mg2+, the main source of Al3+ and Si, and a minor source of Fe(II) and Fe(III). Dissolution of muscovite and plagioclase had only a minor effect on the proton balance (not shown), but were sources of K+, Na+, and Ca2+, as well as of Al3+ and Si. 4.3. Modelled turnover times of minerals Site-specific characteristic turnover times for primary minerals in the unsaturated zone of the deposit were calculated by division of the amount of each mineral (Table 2) by their respective, modelled dissolution rate (see Stro¨mberg and Banwart, 1994) after application of Xr. For full O2 availability (PO2=0.2 atm), pyrite had

Fig. 3. Modelled concentrations of components (this study, PO2=0.2 atm) vs. the arithmetic mean of reported concentrations of components in monthly samples from 3 piezometers at 1.5 m depth below the groundwater table as observed during several years (see Appendix for details). Open symbols indicate initial model results; filled symbols show results after allowance for scale-dependent weathering rates through application of Xr (Xr=0.02). The solid diagonal line indicates ‘‘perfect prediction’’, where modelled and observed concentrations coincide. Vertical bars indicate the range of concentrations reported for the field.

10

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

Fig. 4. Modelled fluxes of components from different processes normalised to the deposit surface area (mmol m
2 s
1), of (a) H+; (b) SO2
4 ; (c) Fe(II); (d) Fe(III); and (e) O2(aq) for a fixed PO2 of 0.2 atm (filled bars) and 0.002 atm (open bars). Processes (i)–(v) are defined in Table 4; Process (vi)–(vii) show the effect of solubility equilibrium between aqueous solution and the given secondary phase, see Table 3.

the shortest turnover time in the order of 100 a. As pyrite is the major source of protons, the characteristic turnover time of pyrite also provides a first hand estimate of the expected duration of AMD production (Banwart and Malmstro¨m, 2001). The characteristic turnover times for the other primary minerals in the model, that is, sphalerite, chalcopyrite, and aluminosilicates, were in the order of 103–104 a. As the mineral turnover times were much longer than the estimated water residence time in the unsaturated zone of 0.5–1 a (Malmstro¨m et al., 2001), these results support the use of a steady-state model in this study.

5. pH-buffering processes: sensitivity analysis As a consequence of natural variability in physicochemical characteristics of the deposit and uncertainty in the experimentally determined kinetic parameters, model input is subject to uncertainty. In the following sections, the sensitivity of model results to uncertainties associated with the major processes and values of key parameters in the calibrated model are tested. In this section, the authors focus on pH-buffering processes and test for model sensitivity to variation in the rate, or the complete removal, of aluminosilicate weathering and Fe(OH)3(s) dissolution. Removal of the forced solubility equilibrium between Fe(OH)3(s) and the aqueous phase, in the presence of aluminosilicate weathering, had little effect on pH and component concentrations, as Fe(OH)3(s) dissolution was not a major process. Removal of both aluminosilicate weathering and Fe(OH)3(s) dissolution from the model resulted in a pH below 2. The impact of uncertainty in the chlorite rate constant and/or chlorite con-

tent was tested by increasing or decreasing the rate expression for chlorite weathering (that is, by manipulating the factor Xr hAch k in Rch in Table 4) by up to an order of magnitude. Increase in Rch by as little as a factor of two led to an increase in pH to around neutral. Further increase in Rch led, as expected, to further increase in pH, as a result of the associated increase in proton consumption, and correspondingly higher concentrations of Mg2+ and Al3+. A decrease in Rch by a factor of two led to a drop in pH by about half a unit. With further decrease in Rch, the influence of the chlorite weathering rate on the proton balance became insignificant. The expected decrease in pH was, however, mitigated to an extent by increased dissolution of Fe(OH)3(s). This kept the pH  4.3 and at the same time implied faster depletion of Fe(OH)3(s). Thus, while field iron and SO2
4 concentrations and pH could be reproduced without aluminosilicate dissolution as long as fast-dissolving minerals were present in the model, accurate representation of the proton balance and the temporal evolution of the geochemistry required consideration of all acidity attenuating processes, including aluminosilicate dissolution. Furthermore, neglect of aluminosilicate weathering reactions excluded the possibility of quantifying the primary sources of components, such as Al3+, Mg2+, Na+, K+ and Ca2+, which are released by these processes, and which may participate in other geochemical reactions subsequent to release.

6. Rates of oxidation: sensitivity analysis Uncertainty in the quantification of the rates of redox reactions includes parametric uncertainty in the

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

employed abiotic rate expressions (Table 4), but also the possibility of other parallel reaction pathways, for example, through microbial mediation, as described in Section 3.4.4. The total reaction rate, r, is the sum of the effect of parallel pathways. This implies that the total reaction rate, thus far in this study represented solely by an abiotic rate law, ra, can contain a component, rb, which quantifies the contribution of the alternative pathway: r ¼ ra þ rb

ð6Þ

In the absence of a rate law explicitly defining the rate of such parallel reactions (see Section 3.4.3), a scalar c (c =1+rb/ra) was applied to the abiotic rate law, according to: r ¼ ð1 þ rb =ra Þra ¼ cra

ð7Þ

Although the rates of the parallel reactions are not mechanistically described, this approach allows assessment of the magnitude of the rate of a parallel reaction that is required in order to have an impact on the modelled system. By applying c to the rate expressions for pyrite oxidation by O2(aq) (Rpyo in Table 4; c=O2 ) and aqueous Fe(III) (Rpyf; c =Fe ), as well as to the oxidation of aqueous Fe(II) (Rfeo; c= ) and varying the values of Fe , O2 , and over a range of values that corresponds to the effect of microbial mediation under favourable laboratory conditions, as reported in the literature (see Section 3.4.4), the sensitivity of model results was tested to rates of redox reactions. As these rates also depend upon O2 availability, the impact of a fixed, lower PO2 was also tested. As conditions of lower O2 availability are often the main aim of tailings deposit remediation, where measures are taken to limit the O2 flux, the simulations with low PO2 also indicate the effect of successful remediation. With PO2=0.2 atm and O2 = =1, a value of Fe of up to 35 had little effect on the groundwater composition, other than a slight increase in pH (not shown). In contrast, a value of O2 of as little as 3 (with Fe = =1) led to higher SO2
4 and iron concentrations and lower pH (Fig. 5a and b). For all tested combinations of values of O2 and Fe , pyrite oxidation was dominated by the dissolved molecular O2 pathway (Reaction 1 in Table 1). The higher rate of proton production due to O2 > 1 was not compensated by an increase in the chlorite dissolution rate, but instead led to increased importance of Fe(OH)3(s) dissolution for the proton balance. The associated increase in Fe(III) release also led to increased importance of the oxidation of pyrite by Fe(III) (Reaction 2 in Table 1), which then contributed up to 40 and  50% of the total proton and Fe(II) release, respectively. For a given value of O2 >1, removal of the condition of solubility equilibrium between Fe(OH)3(s) and solution led to lower pH

11

(e.g., crosses in Fig. 5b; O2 =3.5), as then determined by the balance between pyrite oxidation and aluminosilicate dissolution only. Under these conditions, the contribution by the Fe(III) pathway for pyrite oxidation was insignificant, as there was no readily available source of Fe(III). Increase in alone by over 3 orders of magnitude, and the associated increase in Fe(III) production, led to an increase in pyrite oxidation by Fe(III), an increase in SO2
concentrations (Fig. 5a), and a decrease in pH 4 (Fig. 5b). Already for 100, the importance of pyrite oxidation by Fe(III) as a source of protons, Fe(II), and SO2
increased, as did the importance of aqueous 4 Fe(III) oxidation as a sink for protons; for =104, these processes were dominant, even with higher values of O2 and/or in the absence of solubility equilibrium between solution and Fe(OH)3(s). Increase in Fe slightly enhanced the impact of for > 100 (not shown), as the rate of iron cycling was further increased. Limited O2 availability, as simulated by a fixed partial pressure of 0.002 atm, led to lower release rates of 2+ protons, SO2
(open bars in Fig. 4, 4 , and Fe O2 ¼ Fe = = 1). It also resulted in precipitation of Fe(OH)3(s) and, despite the associated additional proton production (Fig. 4a), higher pH (Fig. 5d). For this low PO2, increase in by as much as 104 had little impact on the SO2
4 concentration (Fig. 5c); there was, however, an impact on the proton balance, as pH decreased by up to one unit (Fig. 5d). In contrast, modelled concentrations of iron and SO2
4 , as well as pH, were sensitive to O2 even at a fixed, low PO2 (Fig. 5c and d). Different combinations of values of O2 , Fe , , and PO2 that resulted in the same value of FO2 produced similar concentrations of dissolved components, as exemplified by SO2
4 in Fig. 5e. For interpretative modelling of field sites, the interdependence of PO2, FO2, and the rates of sulfide oxidation implies that assessment of process dominance and values of O2 and would be facilitated by determination of PO2 as well as FO2 and/or the SO2
4 flux. The modelled pH, on the other hand, was not necessarily constant for a given FO2 (Fig. 5f); the pH was sensitive to the presence or absence of solubility equilibrium between aqueous solution and Fe(OH)3(s), as well as, even at low and approximately constant FO2, the pyrite oxidation pathway dominance, as the different pathways had different effects on the proton balance.

7. Discussion and conclusions In this study, a model is presented for assessment of the geochemistry in carbonate mineral depleted pyritic tailings deposits. The model is built from first principles with regards to biogeochemical processes and specifically includes kinetic representations of slow processes,

12

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

which have often been simplified or neglected in previous modelling studies of tailings deposits. The model is not comprehensive, it is rather a geochemically detailed but otherwise simple model that nonetheless provides valuable insight into modelling of the leachate composition in tailings deposits on the field scale, and in addition, highlights geochemical processes to be included in future, improved models for AMD. Prediction of the groundwater pH, which is determined by the balance between proton producing and proton consuming processes, is particularly important in the modelling of tailings leachate composition, as pH exerts critical control on metal ion mobility and bioavailability after release from primary minerals. The model results indicated that pyrite oxidation through the O2 path is the dominant proton releasing process in pyritic, carbonate mineral depleted mill tailings. These

results are consistent with earlier modelling studies of sulfidic mill tailings deposits, which also showed or assumed that oxidation of iron sulfides, e.g., pyrrhotite (e.g., Bain et al., 2000) or pyrite (e.g., Elberling et al., 1994; Wunderly et al., 1996), was the dominant proton source. In addition, the results indicate increasing importance of proton release from Fe(III) precipitation with decreasing O2 availability, due to the coupled overall increase in pH and increase in the rate of Fe(II) oxidation (Fig. 4; open bars). In the model, important proton consuming processes at full O2 availability were aluminosilicate weathering and, to a lesser extent, dissolution of a Fe(III) (oxy)hydroxide. Due to both the high abundance of chlorite and talc and their relatively fast dissolution compared to other aluminosilicates in the tailings (muscovite and feldspars), weathering of Mg bearing silicates was the


2
1 Fig. 5. Modelled SO2
s ) (panel e and f) with Fe =1, 4 concentrations and pH as a function of (panels a–d) and FO2 (mol m where and Fe represent the magnitude of acceleration of Fe(II) oxidation and the Fe(II) pathway for pyrite oxidation, respectively. Filled symbols indicate PO2=0.2 atm, open symbols indicate PO2=0.002 atm. Symbol shapes indicate the value of O2 : * =1, & =3.5, ^ =10,  =35. Crosses indicate simulations without solubility equilibrium between aqueous solution and Fe(OH)3(s) (O2 =3.5, PO2=0.2 atm).

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

dominant sink for protons in the model. These results agree with earlier conceptual models for mine waste deposits, suggesting that acidity is consumed by dissolution of a sequence of carbonate and (hydr)oxide minerals (Blowes and Ptacek, 1994; Bain et al., 2000) and aluminosilicates (e.g., Banwart and Malmstro¨m, 2001), where the relative rates of aluminosilicate weathering and sulfide mineral oxidation determine the pH once the fast-reacting buffering capacity has been depleted. The results furthermore indicated that fastreacting acidity consuming minerals, such as (oxy)hydroxides and carbonates, in the model represented by Fe(OH)3(s), dissolve to an extent that is determined by the difference between proton release rate and the rate of proton consumption from aluminosilicate weathering. This implies that aluminosilicate weathering plays an important role in prolonging the lifetime of fastreacting buffer capacity in the deposit, and therefore influences the time during which pH is buffered at relatively high levels. For the case study, Impoundment 1 in Kristineberg, northern Sweden, the uncalibrated model, using published empirical rate laws from small-scale, laboratory studies of abiotic weathering of monomineralic samples, successfully reproduced the relative concentrations of major components and the proton balance in the deposit. This implies that the relative reactivity of the various minerals is consistent between the field and laboratory scales. Model reproduction of absolute concentrations in the field was achieved through scaling down all weathering rates with a factor Xr=0.02 as the only calibration, thereby allowing for the commonly observed scale-dependence of mineral weathering rates in the model. This magnitude of scale-dependence is approximately the same as reported in other studies. For example, Stro¨mberg and Banwart (1994), in their model assessment of mine waste rock heaps at the Aitik site, had to scale down all mineral weathering rates by a factor of 0.006 to arrive at agreement between model results and field observations. For tailings from Impoundment 1 in the specific case study, laboratory weathering rates were observed (Salmon and Malmstro¨m, 2001) that were 1 to 5 orders of magnitude greater than those previously estimated for the field (Banwart and Malmstro¨m, 2001). In a detailed study, Malmstro¨m et al. (2000) showed that the scale-dependence of mineral weathering rates in waste rock from the Aitik site to a large degree could be accounted for through quantitative consideration of differences in temperature, pH, mineral content, hydrological flow paths, and particle size distribution between the observations scales. In this modelling study, the authors explicitly accounted for the effects of mineral content, temperature, and pH on the mineral weathering rates in the employed rate expressions; hence, the calibration factor used in the study, Xr, quantified any

13

remaining scale-dependence. Further quantitative assessment of the discrepancy between mineral weathering rates determined under controlled conditions in the laboratory and those reported for the field is thus important in order to improve the basic understanding of dominant weathering processes on the field scale. As a consequence of uncertainty in experimentally determined kinetic parameters as well as of natural variability in physico-chemical characteristics of tailings deposits in combination with generally sparse mine waste site data (Banwart and Malmstro¨m, 2001), model input is subject to uncertainty. The model results were found to be sensitive to oxidation rates and pathways, demonstrating complex feedback mechanisms between pH, kinetic and equilibrium processes, and O2 availability. Interestingly, at full O2 availability, a direct increase of the rate of oxidation of pyrite via the Fe(III) oxidation pathway, which recent publications indicate to be the main processes involved in bacterial mediation of pyrite oxidation (Sand et al., 2001; Fowler et al., 2001; Yu et al., 2001), had little effect on the modelled groundwater composition. In fact, the indirect effect of an increase in the rate constant for pyrite oxidation by O2 or for Fe(II) oxidation gave a more pronounced effect on the rate of pyrite oxidation by Fe(III) than did a direct increase in the rate constant of this process. The complexity of feedback mechanisms implied by these results indicates the need for further systematic quantification of sulfide and Fe(II) oxidation reactions rates, including the effect of microbial mediation. The modelled pH and groundwater quality were relatively insensitive to a decrease in the rate of aluminosilicate weathering, as long as Fe(OH)3(s) was present and dissolving. However, model results were sensitive to an increase in the aluminosilicate weathering rate, with increased concentrations of dissolved base cations and Al3+ and approach to alkaline pH in the model. The sensitivity of model results with respect to the rate of aluminosilicate weathering, in combination with uncertainty in site specific kinetic parameter values, furthermore implies uncertainty in the appropriate value of the calibration factor, Xr, that accounts for scaledependence in process rates. The demonstrated complex interdependence between dominant processes in tailings deposits in combination with the generally large uncertainty in model input imply that interpretative and predictive model assessments of mill tailings geochemistry requires thorough consistency tests between model results and field observations in order to be reliable.

Acknowledgements We gratefully acknowledge Dr. Erik Carlsson, Lulea˚ Technical University, Sweden, for providing tailings

14

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

samples for BET analysis and Professor Janet Herman (Associate Editor) and 3 anonymous reviewers for valuable comments on the manuscript during the peerreview process. This research was funded by the Swedish Foundation of Strategic Research (MISTRA) through the research programme ‘‘Mitigation of the environmental impact from mining waste’’ (MiMi).

Appendix Site description—tailings deposit ‘‘Impoundment 1’’ in Kristineberg, northern Sweden Background The Kristineberg mine is owned and operated by Boliden Mineral AB. The tailings deposit referred to as ‘‘Impoundment 1’’ was in use from the 1940s to the 1950s for deposition of tailings from the Zn–Cu–Pb sulfide ore mine at Kristineberg, as well as from other mines in the area. Impoundment 1 was remediated in 1997 by combined dry cover application and raised groundwater level (Lindvall et al., 1999). In this study, the authors considered pre-remediation conditions, for which field data (e.g., du Rietz, 1953; Qvarfort, 1983; Axelsson et al., 1986, 1991; Ekstav and Qvarfort, 1989) from sampling campaigns in the 1980s and 1990s have been compiled in English by Malmstro¨m et al. (2001; available at www.mimi.kiruna.se); a summary of results relevant for this modelling study is provided in this Appendix. Holmstro¨m et al. (2001), Werner et al. (2001), Corre´ge et al., (2001), and Carlsson et al. (2002, 2003) are referred to for further post-remediation characterisation of the deposit. Physical and hydrological characteristics Impoundment 1 covers an area of approximately 0.11 km2 and is on average approximately 5 m deep. The annual average air temperature at the site is 1  C, with 5 months of average temperature below 0  C. The deposit was formed by pumping tailings as a slurry into an impoundment; when deposition ceased, the decrease in the water level led to the formation of an unsaturated zone. The average depth to the groundwater, as observed in 16 piezometers over 2 a, is reported to have been 1 m, with average annual fluctuations of 1 m (Axelsson et al., 1986). In addition to the effective infiltration of 27 000 m3 a
1, as estimated from precipitation and evapotranspiration at the site, recharge from the surrounding moraine slopes led to a total flow of water through the deposit of 43 000–51 000 m3 a
1 (Malmstro¨m et al., 2001). Most of the discharge was collected in ditches, allowing the estimate of the water flow through the deposit to be made; based on a detailed water balance for the site, it was also estimated that 10– 20% of the discharge from the deposit passed to under-

lying fractured rock (Axelsson et al., 1986), which implies that there was a (small) vertical component of water flow within the deposit. The tailings consist of a finely ground material (characterised as silty fine sand or fine sand), as determined from sieve analysis of 26 samples from 8 different locations and several depths within the deposit (e.g., Axelsson et al., 1986; Qvarfort, 1983). The particle size distribution varied between samples, with 20–85 wt.% of the particles being smaller than 100 mm. Neither concentrations of individual pore gases, the influx of O2, the water content of the unsaturated zone, nor the O2 diffusivity has been reported for the preremediation conditions. Specific surface area of the tailings In this study, the specific surface area of the tailings was assessed using 5-point, N2 gas adsorption data evaluated through the BET equation, a method commonly used in kinetic mineral weathering studies. The specific surface area of 8 samples from 7 locations within the deposit ranged from 0.2 to 10.1 m2 g
1, with an average value of 2.96 m2 g
1. The large variability in specific surface area is consistent with the reported highly variable particle size distribution (see above). Mineralogical and chemical characteristics of the tailings Based on determination of the chemical composition of samples from drill cores from 6 locations within the deposit, the upper part of the deposit, on average down to the groundwater level, had been partly depleted in S, Fe, Zn and Cu (Qvarfort, 1983; Axelsson et al., 1991), suggesting weathering of sulfides in the unsaturated zone. The main minerals in the tailings, as deduced from the reported mineralogical composition of ore and gangue material at the Kristineberg mine (du Rietz, 1953; Qvarfort, 1983), were chlorite, talc, muscovite, quartz and pyrite. Sphalerite and chalcopyrite were also present. More detailed post-remediation characterisations of the tailings confirmed this mineralogy and also indicated the presence of feldspars and minor amounts of pyrrhotite, galena and arsenopyrite (Holmstro¨m et al., 2001; Gleisner et al., 2003). The average pH of 4.9 and aqueous Mg2+ and Ca2+ concentrations (see below), in combination with a reported low carbonate content, as determined for tailings samples from 5 depths at 4 of the 6 different locations, and low to negative neutralising potential (NP), as determined by standard back-titration method for the same samples (see summary in Malmstro¨m et al., 2001), indicated that the reported 10 vol.% carbonate content of the waste rock fraction from the Kristineberg mine (Qvarfort, 1983) was likely to have been effectively depleted during the 30–50 a between deposition and characterization of the tailings. The net neutralizing potential (NNP), as determined from estimated acid

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

producing potential (AP), based on sulfide content, and experimentally determined NP, was negative for all the above mentioned samples. Identification and quantification of secondary minerals have not been performed, other than qualitative observations of the weathered zone of the tailings, which indicate that there were layers of yellow-brown material present in surficial parts of the deposit (Qvarfort, 1983), suggesting the presence of a secondary Fe(III) phase. It is also noted that no hardpan was visible (Qvarfort, 1983). Groundwater quality Determination of the composition of groundwater at depths of 1.5 and 7 m below the groundwater table had been performed using 3–5 a of monthly samples from piezometers at 3 locations within the deposit (Ekstav and Qvarfort, 1989). For comparison to model results in this study, the authors used the arithmetic mean of data from Piezometers 3, 4, and 4H, at 1.5 m below the groundwater surface over the sampling occasions (see Fig. 1). The groundwater contained high concentrations 2+ of SO2
, Zn2+, Ca2+ and Al3+, as well as 4 , Fe, Mg lower concentrations of Na+, Cu2+, and K+ (see Fig. 3). The concentrations of dissolved ions varied between the sampling locations and depths and over the year, however, the solute concentrations differed more between different sampling points than between the two sampling depths at the same location. Neither redox potential nor redox speciation of iron in the groundwater has been reported in the literature. Saturation index calculations performed for this study with the PHREEQC code (Parkhurst, 1995) in conjunction with the WATEQ4F database (Ball and Nordstrom, 1991) indicated that, of phases likely to be controlling aqueous concentrations in AMD environments (Nordstrom and Alpers, 1999), the groundwater was close to saturation with respect to gypsum (CaSO4.2H2O(s)), amorphous silica (SiO2(s)), and for pe values of 5–7, an amorphous ferric hydroxide phase (Fe(OH)3(s)). References Allison, J.D., Brown, D.S., Novo-Gradac, K.J., 1991. MINTEQA2/PRODEFA2, A geochemical assessment model for environmental systems; Version 3.0 User’s Manual. EPA/ 600/3-91/021, Environmental Research Laboratory, Office of Research and Development, US Environmental Protection Agency, Athens, Georgia. Axelsson, C.-L., Karlqvist, L., Lintu, Y., Olsson, T., 1986. Gruvindustrins restproduktupplag- fa¨ltunderso¨kningar med vattenbalansstudie i Kristineberg. Uppsala Geosystem AB, Uppsala, Sweden (in Swedish). Axelsson, C.-L., Ekstav, A., Jansson, T., 1991. Provtagning av sand och grundvatten i sandmagasin 1, 1B och 2 Kristine-

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berg - Fa¨ltrapport. Rapport 917-1687, Golder Geosystems AB, Uppsala, Sweden (in Swedish). Bain, J.G., Blowes, D.W., Robertson, W.D., Frind, E.O., 2000. Modelling of sulfide oxidation with reactive transport at a mine drainage site. J. Contam. Hydrol 41, 73–94. Ball, J.W., Nordstrom, D.K., 1991. User’s manual for WATEQ4F, with revised thermodynamic database and test cases for calculating speciation of major, trace and redox elements in natural waters. US Geol. Surv. Open-File Rep. 91-183 (Revised and reprinted August 1992.). Banwart, S., Malmstro¨m, M.E., 2001. Hydrochemical modelling for preliminary assessment of minewater contamination. J. Geochem. Explor 74, 73–97. Blowes, D.W., Ptacek, C.J., 1994. Acid-neutralization mechanisms in inactive mine tailings. In: Jambor, J.L., Blowes, D.W. (Eds.), Environmental Geochemistry of Sulphide MineWastes. Min. Assoc. Can, Nepean, Canada, pp. 271–292. (Chapter 10).. Brown, K.G., Bassett, R.L., Glynn, P.D., 1998. Analysis and simulation of reactive transport of metal contaminants in ground water in Pinal Creek Basin, Arizona. J. Hydrol 209, 225–250. Brown, P.L., Lowson, R.T., 1997. The use of kinetic modelling as a tool in the assessment of contaminant release during rehabilitation of a uranium mine. J. Contam. Hydrol 26, 27– 34. Brown, P.L., Ritchie, I.M., Bennett, J.W., 2000. Geochemical kinetic modelling of acidic rock drainage. In: Proc. ICARD 2000, 5th Internat. Conf. Acid Rock Drainage, Denver, Colorado, USA, 21–24 May, 2000, vol. 1, pp. 289–296. Carlsson, E., Thunberg, J., O¨hlander, B., Holmstro¨m, H., 2002. Sequential extraction of sulfide-rich tailings remediated by the application of till cover, Kristineberg mine, northern Sweden. Sci. Total Environ 299, 207–226. Carlsson, E., O¨hlander, B., Holmstro¨m, H., 2003. Geochemistry of the infiltrating water in the vadose zone of a remediated tailings impoundment, Kristineberg mine, northern Sweden. Appl. Geochem. 18, 659–674. Corre´ge, O., Carlsson, E., O¨hlander, B., 2001. Geochemical investigation of the groundwater in sulphide-bearing tailings remediated by applying till cover, Kristineberg, northern Sweden. In: Proc. Securing the Future, Internat. Conf. Mining and the Environment, Skelleftea˚, Sweden, 25 June-1 July, 2001, pp. 97–114. Edwards, K.J., Bond, P.L., Druschel, G.K., McGuire, M.M., Hamers, R.J., Banfield, J.F., 2000. Geochemical and biological aspects of sulfide mineral dissolution: Lessons from Iron Mountain, California. Chem. Geol. 169, 383–397. Ekstav, A., Qvarfort, U., 1989. Metallbalans Kristineberg. Quaternary Geology, Dept. Earth Sciences, Uppsala universitet, Uppsala, Sweden. (in Swedish). Elberling, B., Nicholson, R.V., 1996. Field determination of sulphide oxidation rates in mine tailings. Water Resour. Res. 32, 1773–1784. Elberling, B., Nicholson, R.V., Scharer, J.M., 1994. A combined kinetic and diffusion model for pyrite oxidation in tailings: a change in controls with time. J. Hydrol 157, 47–60. Elberling, B., Schippers, A., Sand, W., 2000. Bacterial and chemical oxidation of pyritic mine tailings at low temperatures. J. Contam. Hydrol 41, 225–238. Evangelou, V.P., Zhang, Y.L., 1995. A review: pyrite oxidation

16

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17

mechanisms and acid mine drainage prevention. Crit. Rev. Environ. Sci. Tech. 25, 141–199. Fowler, T.A., Holmes, P.R., Crundwell, F.K., 2001. On the kinetics and mechanism of the dissolution of pyrite in the presence of Thiobacillus ferrooxidans. Hydrometal 59, 257– 270. Furrer, G., Westall, J., Sollins, P., 1989. The study of soil chemistry through quasi-steady-state models: I. Mathematical definition of model. Geochim. Cosmochim. Acta 53, 595–601. Gleisner, M., Herbert, R., Salmon, S.U., Malmstro¨m, M.E. 2003. Comparison of sulfide oxidation in unweathered pyritic mine tailings, accepted for presentation at ICARD 2003 6th Internat. Conf. Acid Rock Drainage, Cairns, Australia, 12– 18 July 2003. Holmstro¨m, H., Salmon, U.J., Carlsson, E., Paraskev, P., O¨hlander, B., 2001. Geochemical investigations of sulfidebearing tailings at Kristineberg, northern Sweden, a few years after remediation. Sci. Total. Environ 273, 111–133. Janzen, M.P., Nicholson, R.V., Scharer, J.M., 2000. Pyrrhotite reaction kinetics: Reaction rates for oxidation by oxygen, ferric iron, and for nonoxidative dissolution. Geochim. Cosmochim. Acta 64, 1511–1522. Jaynes, D.B., Rodowski, A.S., Pionke, H.B., 1984. Acid mine drainage from reclaimed coal strip mines 1. Model description. Water Resour. Res. 20, 233–242. Kirby, C.S., Elder Brady, J.A., 1998. Field determination of Fe2+ oxidation rates in acid mine drainage using a continuously-stirred tank reactor. Appl. Geochem. 13, 509–520. Kirby, C.S., Thomas, H.M., Southam, G., Donald, R., 1999. Relative contributions of abiotic and biological factors in Fe(II) oxidation in mine drainage. Appl. Geochem. 14, 511– 530. Knauss, K.G., Wolery, T.J., 1989. Muscovite dissolution kinetics as a function of pH and time at 70  C. Geochim. Cosmochim. Acta 53, 1493–1501. Lacey, D.T., Lawson, F., 1970. Kinetics of liquid-phase oxidation of acid ferrous sulfate by the bacterium Thiobacillus ferrooxidans. Biotechnol. Bioeng. 12, 29–50. Lichtner, P.C., 1996. Continuum formulation of multicomponent-multiphase reactive transport. In: Lichtner, P.C., Steefel, C.I., Oelkers, E.H. (Eds.), Reactive Transport in Porous Media. Rev. Mineral., Mineralogical Soc. Am., Washington, DC, USA, pp. 1–81. (Chapter 1). Lindvall, M., Eriksson N., Ljungberg J., 1999. Decommissioning at Kristineberg Mine, Sweden. In: Proc. Sudbury ’99, Conf. Mining and the Environment II, Sudbury, Canada, 13–15 September 1999, pp. 855–862. Lowson, R.T., 1982. Aqueous oxidation of pyrite by molecular oxygen. Chem. Rev. 82, 461-497. Malmstro¨m, M., Banwart, S., Duro, L., Wersin, P., Bruno, J., 1995. Biotite and chlorite weathering at 25  C: The dependence of pH and (bi)carbonate on weathering kinetics, dissolution stoichiometry and solubility; and the relation to redox conditions in granitic aquifers. Tech. Rep. 95-01, The Swedish Nuclear Fuel and Waste Co. (SKB), Stockholm, Sweden. Malmstro¨m, M.E., Destouni, G., Banwart, S.A., Stro¨mberg, B.H.E., 2000. Resolving the scale-dependence of mineral weathering rates. Environ. Sci. Technol. 34, 1375–1378. Malmstro¨m, M., Werner, K., Salmon, U., Berglund, S.,

2001. Hydrogeology and geochemistry of mill tailings impoundment 1, Kristineberg, Sweden: Compilation and interpretation of pre-remediation data. MiMi 2001:4, Mitigation of the environmental impact from mining waste programme (MiMi), Stockholm, Sweden. (available at www.mimi.kiruna.se). Mayer, K.U., Benner, S.G., Blowes, D.W., 1999. The reactive transport model MIN3P: application to acid mine drainage and treatment - Nickel Rim Mine Site, Sudbury, Ontario. In: Proc. Sudbury ’99, Conf. Mining and the Environment II, Sudbury, Canada, 13–15 September 1999, pp. 145–154. Mayer, K.U., Blowes D.W., Frind E.O., 2000. Numerical modelling of acid mine drainage generation and subsequent reactive transport. In: Proc. ICARD 2000, 5th Internat. Conf. Acid Rock Drainage, Denver, Colorado, USA, 21–24 May, 2000, Vol. 1. pp. 135–141. Mayer, K.U., Frind, E.O., Blowes, D.W., 2003. Multicomponent reactive transport modelling in variably saturated porous media using a generalized formulation for kinetically controlled reactions. Water. Resour. Res. 38, 1174. McKibben, M.A., Barnes, H.L., 1986. Oxidation of pyrite in low temperature acidic solutions: rate laws and surface textures. Geochim. Cosmochim. Acta 50, 1509–1520. Moses, C.O., Herman, J.S., 1991. Pyrite oxidation at circumneutral pH. Geochim. Cosmochim. Acta 55, 471–482. Nagy, K.L., 1995. Dissolution and precipitation kinetics of sheet silicates. In: White, A.F., Brantley, S.L. (Eds.), Chemical Weathering Rates of Silicate Minerals. Rev. Mineral., Mineralogical Soc. Am, Washington DC, USA, pp. 173–234. (Chapter 5). Nemati, M., Webb, C., 1997. A kinetic model for biological oxidation of ferrous iron by Thiobacillus ferrooxidans. Biotechnol. Bioeng. 53, 477–486. Nemati, M., Harrison, S.T.L., Hansford, G.S., Webb, C., 1998. Biological oxidation of ferrous sulphate by Thiobacillus ferrooxidans: a review on the kinetic aspects. Biochem. Eng. J. 1, 171–190. Nicholson, R.V., Gillham, R.W., Cherry, J.A., Reardon, E.J., 1989. Reduction of acid generation in mine tailings through the use of moisture-retaining cover layers as oxygen barriers. Can. Geotech. J. 26, 1–8. Nordstrom, D.K., Alpers, C.N., 1999. Geochemistry of acid mine waters. In: Plumlee, G.S., Logsdon, M.J. (Eds.), The Environmental Geochemistry of Mineral Deposits. Reviews in Economic Geology, Soc. Econ. Geologists, Chelsea, USA, pp. 133–160. (Chapter 6). Nordstrom, D.K., Southam, G., 1997. Geomicrobiology of sulfide mineral oxidation. In: Banfield, J.F., Nealson, K.H. (Eds.), Geomicrobiology: Interactions Between Microbes and Minerals. Rev. Mineral., Mineralogical Soc. Am., Washington DC, USA, pp. 361–390. (Chapter 11). Nordstrom, D.K., Plummer, N., Langmuir, D., Busenberg, E., May, H.M., Jones, B.F., Parkhurst, D.L., 1990. Revised chemical equilibrium data for major water-mineral reactions and their limitations. In: Melchior, D.C., Bassett, R.L. (Eds.), Chemical Modelling of Aqueous Systems II, Vol. 416. ACS Symposium Series, Washington DC, USA, pp. 398– 413. (Chapter 31). Olson, G.J., 1991. Rate of pyrite bioleaching by Thiobacillus

S.U. Salmon, M.E. Malmstro¨m / Applied Geochemistry 19 (2004) 1–17 ferrooxidans: Results of an interlaboratory comparison. Appl. Environ. Microbiol. 57, 642–644. Oxburgh, R., Drever, J.I., Sun, Y.-T., 1994. Mechanism of plagioclase dissolution in acid solution at 25  C. Geochim. Cosmochim. Acta 58, 661–669. Parkhurst, D.L., 1995. User’s guide to PHREEQC - a computer program for speciation, reaction-path, advective transport and inverse geochemical calculations. US Geol. Surv. Water-Resour. Invest. Rep. 95-4227. Pesic, B., Oliver, D.J., Wichlacz, P., 1989. An electrochemical method of measuring the oxidation rate of ferrous to ferric iron with oxygen in the presence of Thiobacillus ferrooxidans. Biotech. Bioeng 33, 428–439. Qvarfort, U., 1983. En kartla¨ggning av sandmagasin fra˚n sulfidmalmsbrytning. Naturva˚rdsverkets Rapport SNV PM 1699, The Swedish National Environment Protection Agency, Stockholm, Sweden (in Swedish). du Rietz, T., 1953. Geology and ores of the Kristineberg deposit, Vesterbotten, Sweden. Ser C. 524, A˚rsbok 45(5), Geological Survey of Sweden, Uppsala, Sweden. Rimstidt, J.D., Newcomb, W.D., 1993. Measurement and analysis of rate data: the rate of reaction of ferric iron with pyrite. Geochim. Cosmochim. Acta 57, 1919–1934. Rimstidt, J.D., Chermak, J.A., Gagen, P.M., 1994. Rates of reaction of galena, sphalerite, chalcopyrite and arsenopyrite with Fe(III) in acidic solutions. In: Alpers, C.N., Blowes, D.W. (Eds.), Environmental Geochemistry of Sulfide Oxidation, Vol. 550. ACS symposium series, Washington D.C., USA, pp. 2–13. (Chapter 1). Salmon, S.U., 2000. Biogeochemical processes in mill tailings: Modelling and assessment of remediation effects. Licentiate thesis TRITA-AMI LIC 2053, Water Resources Engineering, Dept. Civil and Environmental Engineering, Royal Institute of Technology, Stockholm, Sweden. Salmon, U.J., Malmstro¨m, M., 2000. Assessing geochemical processes in mill tailings impoundments and the effect of remediation. In: Proc. ICAM 2000, 6th Internat. Conf. Applied Mineralogy, Go¨ttingen, Germany, 13–21 July, 2000, pp. 671–673. Salmon, S.U., Malmstro¨m, M.E., 2001. Quantification of mineral weathering rates in tailings from Impoundment 1, Kristineberg, northern Sweden. In: Proc. Securing the Future, Internat. Conf. Mining and the Environment, Skelleftea˚, Sweden, 25 June–1 July, 2001, pp. 747–756. Salmon, S.U., Malmstro¨m, M., 2002. Steady state, geochemical box model of a tailings impoundment: Application to Impoundment 1, Kristineberg, and prediction of effect of remediation. MiMi 2002:2, Mitigation of the environmental impact from mining waste programme (MiMi), Stockholm, Sweden. (available at www.mimi.kiruna.se). Sand, W., Gehrke, T., Jozsa, P.-G., Schippers, A., 2001. (Bio)chemistry of bacterial leaching—direct vs. indirect bioleaching. Hydrometallurgy 59, 159–175. Scharer, J.M., Nicholson, R.V., Halbert, B., Snodgrass, W.J., 1994. A computer program to assess acid generation in pyri-

17

tic tailings. In: Alpers, C.N., Blowes, D.W. (Eds.), Environmental Geochemistry of Sulfide Oxidation. ACS symposium series, Washington DC, USA, pp. 132–152. (Chapter 11). Singer, P.C., Stumm, W., 1970. Acid mine drainage: the ratedetermining step. Science 167, 1121–1123. Stro¨mberg, B., Banwart, S., 1994. Kinetic modelling of geochemical processes at the Aitik mining waste rock site in northern Sweden. Appl. Geochem. 9, 583–595. Stro¨mberg, B., Banwart, S., 1999. Weathering kinetics of waste rock from the Aitik copper mine, Sweden: Scale-dependent rate factors and pH controls in large column experiments. J. Contam. Hydrol 39, 59–89. Sverdrup, H.U., 1990. The Kinetics of Base Cation Release Due to Chemical Weathering. Lund University Press, Lund, Sweden. Tamura, H., Goto, K., Nagayama, M., 1976. The effect of ferric hydroxide on the oxygenation of ferrous ions in neutral solutions. Corrosion Sci. 16, 197–207. Wehrli, B., 1990. Redox reactions of metal ions at mineral surfaces. In: Stumm, W. (Ed.), Aquatic Chemical Kinetics. Wiley, New York, USA, pp. 311–336. (Chapter 11). Werner, K., Berglund S., 1999. Effects of spatial variability on oxygen transport in sulphidic mining waste deposits. In: Proc. Sudbury ’99, Conf. Mining and the Environment II, Sudbury, Canada, 13–17 September, 1999, pp. 243–252. Werner, K., Carlsson, E., Berglund, S., 2001. Oxygen and water fluxes into a soil-cover remediated mill tailings deposit: Evaluation of field data from the Kristineberg mine site, northern Sweden. In: Proc. Securing the Future, Internat. Conf. Mining and the Environment, Skelleftea˚, Sweden, 25 June–July 1, 2001, pp. 896–905. White, A.F., Brantley, S.L., 1995. Chemical Weathering Rates of Silicate Minerals. Rev. Mineral.. Mineralogical Soc. Am., Washington DC, USA. White, A.F., Peterson, M.L., 1990. Role of reactive-surface area characterization in geochemical kinetic models. In: Melchoir, D.L., Bassett, R.L. (Eds.), Chemical Modelling of Aqueous Systems II, vol. 416. ACS symposium series, Washington D.C., USA, pp. 461–475. (Chapter 35). Wiersma, C.L., Rimstidt, J.D., 1984. Rates of reaction of pyrite and marcasite with ferric iron at pH 2. Geochim. Cosmochim. Acta 48, 85–92. Williamson, M.A., Rimstidt, J.D., 1994. The kinetics and electrochemical rate-determining step of aqueous pyrite oxidation. Geochim. Cosmochim. Acta 58, 5443–5545. Wunderly, M.D., Blowes, D.W., Frind, E.O., Ptacek, C.J., 1996. Sulfide mineral oxidation and subsequent reactive transport of oxidation products in mine tailing impoundments: a numerical model. Water Resour. Res. 32, 3173– 3187. Yu, J.-Y., McGenity, T.J., Coleman, M.L., 2001. Solution chemistry during the lag phase and exponential phase of pyrite oxidation by Thiobacillus ferrooxidans. Chem. Geol. 175, 307–317.