DOCTORA L T H E S I S

Elsa N

yström Preventing A

cid Rock D

rainage Formation from

Sulfidic Waste R

ock Using Secondary R

aw M

aterials

Department of Civil, Environmental and Natural Resources EngineeringDivision of Geosciences and Environmental Engineering

ISSN 1402-1544ISBN 978-91-7790-825-8 (print)ISBN 978-91-7790-826-5 (pdf)

Luleå University of Technology 2021

Preventing Acid Rock Drainage Formation from Sulfidic Waste Rock

Using Secondary Raw Materials

Elsa Nyström

Applied Geochemistry

Tryck: Lenanders Grafiska, 136575

136575 LTU_Nystrom.indd Alla sidor136575 LTU_Nystrom.indd Alla sidor 2021-05-10 11:042021-05-10 11:04

PREVENTING ACID ROCK DRAINAGE FORMATION FROM SULFIDICWASTE ROCK USING SECONDARY RAW MATERIALS

Elsa Nyström

Department of Civil, Environmental and Natural Resources Engineering

Luleå University of Technology

Luleå, Sweden

Supervisors Lena Alakangas, Andrzej Cwirzen, Christian Maurice

Printed by Lenanders Grafiska AB

ISSN: 1402-1544

ISBN: 978-91-7790-825-8 (print)

ISBN: 978-91-7790-826-5 (electronic)

Luleå 2021

www.ltu.se

Cover image: Secondary precipitates on sulfidic waste rock leached for 100 weeks with 5 wt.% lime kiln dust.

Till mamma och pappa

ABSTRACT

One of the central and most challenging environmental problems related to mining is the gen-eration of acid rock drainage (ARD). The drainage is characterized by low pH and elevated concentrations of sulfate, metals, and metalloids formed when sulfide-bearing minerals are sub-jected to oxygen and water. Current remediation solutions, including active and passive tech-niques, have been developed to reduce ARD's negative impact. However, these treatments re-quire continuous maintenance due to a persistent need for the addition of chemicals, requiring a high energy output, and a need for long-term monitoring until sulfide oxidation has ceased. Once it has been initiated, ARD formation could last for hundreds to thousands of years, making these approaches costly and unsustainable. A more strategic and environmentally sustainable approach would focus on preventing sulfide oxidation rather than treating its symptoms.

This thesis explores five different secondary raw materials (SRM) (denoted below) for their use as amendments to prevent pyrite oxidation during storage. A combination of several mineralog-ical and geochemical methods was used to assess the materials' ability to maintain circumneu-tral pH when leaching pyritic waste rock (> 60% pyrite) in small-scale test cells (10L) to pro-mote hydrous ferric oxides precipitation on the pyrite surfaces.

The oxidation of waste rock resulted in drainage characterized by low pH (<2) and extensive element mobilization of up to 80% of the original content during the first two years. The results highlight the importance of trace element characterization and the need for early preventive measures to hinder or reduce the risk of acid drainage formation that requires active and costly long-term treatment. Conversely, adding 1-5 wt.% SRM to the waste rock created drainages with circumneutral pH and substantially lower sulfate and metal concentrations. However, not all materials could maintain circumneutral pH for an extended time, such as blast furnace slag (air-cooled and granulated) and cement kiln dust. These materials either require larger volumes of water to dissolve or contain minerals that allowed the material to harden upon contact with water which can inhibit its neutralization capacity. Biomass fly ash showed a similar but less extensive, hardening effect resulting in a better ability to maintain circumneutral pH for more than two years despite its small addition (1-2.5wt.%). A similar ability was observed for lime kiln dust (5 wt.%). Conversely, to lime kiln dust, the ash contained high soluble elements of potential concern, and its usage should be questioned despite only a temporary increase of ele-ments through wash-out. However, the correlation between the amount of fly ash added and the timespan of circumneutral pH was not linear, resulting in the risk of prematurely declining pH if too little is added. Conversely, adding too much fly ash increases the risk of material harden-ing.

One major concern with this treatment method is that it can inflate secondary minerals for-mation, leading to latent acidity and element release through their dissolution in changing geo-chemical conditions, such as wet or dry coverage measures. However, the addition of small

amounts of SRM (1-4% of the waste rock's net neutralizing potential) to the waste rock dramat-ically improves the overall drainage quality without increasing the total amount of secondary minerals formed compared to no addition. In general, the type of secondary minerals formed on the waste rock without SRM treatment was considered less stable in an oxidizing environ-ment than those formed through SRM treatment, suggesting that not treating the waste rock is inferior to SRM treatment both before and after covering measures.

In conclusion, this thesis's results show that using small amounts of SRM can prevent oxidation during at least two years, likely due to HFO formation on the reactive surfaces. Consequently, it can substantially limit the need for treatment measures, both before and after remediation, decreasing the overall need and cost for chemicals, energy, and long-term monitoring, stressing the need for applying preventive measures during the storage time from mining to remediation. However, secondary minerals' long-term stability needs further evaluation and understanding before this method can be applied on a larger scale.

ACKNOWLEDGMENT

Although this thesis only holds my name, there are many other people without whom this work would not have seen the light of day.

First of all, I want to thank Prof. Lena Alakangas, my supervisor. I suppose you saw something in me that I could not see myself, and for that, I am forever grateful. Thank you for giving me this opportunity, for pushing me forwards, for pulling me back, and for always challenging me to never stop growing as a researcher and as a human being. Secondly, I want to thank my other supervisors Prof. Andrzej Cwirzen and Dr. Christian Maurice, for giving perspective to my work and helping at all levels from sample preparation to reviewing and everything in between.

Sometimes the most important work is performed over a cup of coffee. Therefore, I want to thank all my friends, previous and current colleagues at Luleå University of technology, espe-cially Anders, Britta, Björn, Helen, Joel, Johan, Leslie, Malin, Musah, Tobi, and Thomas, for sharing their ideas, always being open for discussions, and for making work one of my favorite places to be. I want to give a special thanks to Hanna, Lina, Roger, Sarah, Susi, Sussi, and Stacy for always being there, both professionally and personally. Thank you for sharing your time and painting in colors when all seemed black.

To Johan, life is not about the destination, it is about the journey, and although our journey has not always been easy, I am glad you are here to take it with me. Thank you for being the most important rock in my life. And to Edwin and Klara. May you never blindly accept other peoples' truths as your own, and may you never stop asking the question of "why". Regardless of previ-ous and future achievements, being your mom is and always will be, by far, my greatest accom-plishment.

Last but not least, my parents to whom this thesis is dedicated. Tack för allt ni gjort, fortfa-rande gör och för att er kärlek inte finner gränser. Utan er vore jag ingenting men tack vare er kan jag bli vad som helst.

This work was supported by the Strategic Innovation Programme for the Swedish Mining and Metal Producing Industry (SIP-STRIM) of Vinnova, Formas, and the Swedish Energy Agency, with co-financial support from Boliden Mineral AB, Centre of Advanced Mining and Metal-lurgy (CAMM2) at Luleå University of Technology and. J. Gust. Richert foundation.

LIST OF PAPERS

The thesis is based on the following papers:

I. Nyström, E., Thomas, H., Wanhainen, C., Alakangas, L. (2021). Occurrence and Release ofTrace Elements in Pyrite-Rich Waste Rock. Minerals, 11(5), 495. doi.org/10.3390/min11050495

II. Nyström, E., Kaasalainen, H., & Alakangas, L. (2019). Suitability study of secondary raw ma-terials for prevention of acid rock drainage generation from waste rock. Journal of CleanerProduction, 232, 575-586. doi:10.1016/j.jclepro.2019.05.130

III. Nyström, E., Kaasalainen, H., & Alakangas, L. (2019). Prevention of sulfide oxidation in wasterock by the addition of lime kiln dust. Environmental Science and Pollution Research, 26(25),25945-25957. doi:10.1007/s11356-019-05846-z

IV. Nyström, E., Cwirzen, A., Alakangas, L. (Manuscript) The use of biomass fly ash to preventsulfide oxidation in waste rock.

The author's contribution to the appended papers was:

I. Most of the laboratory work, interpretation of data, and writing

II. Laboratory work, interpretation of data, and most of the writing

III. Most of the laboratory work, interpretation of data, and writing

IV. Most of the laboratory work, interpretation of data, and writing

Publications that have been written within the degree but are not included in this thesis:

Peer-reviewed conference proceedings:

Nyström, E., Kaasalainen, H., Alakangas, L. (2017). Prevention of sulfide oxidation in waste rock using by-products and industrial remnants: A suitability study. Paper presented at the 13th International Mine Water Association Congress – "Mine Water & Circular Economy – A Green Congress", Lappeenranta, Finland, 25-30 June 2017. 1170-1178

Peer-reviewed conference abstracts:

Nyström, E., Alakangas, L. (2019). The occurrence of As, Hg, Sb, and Tl in pyritic waste rock and the ability to prevent their release. Abstract presented at the Goldschmidt conference, Barcelona, Spain, 18-23 August 2019.

Nyström, E., Alakangas, L. (2015). Prevention of sulfide oxidation in sulfide-rich waste rock. Abstract presented at the European Geosciences Union - General Assembly 2015, Vienna, Austria, 12-17 April 2015.

Licentiate thesis:

Nyström, E. (2018). Suitability of industrial residues for preventing acid rock drainage gener-ation from waste rock. Licentiate thesis, Luleå University of Technology, Sweden

PART I

INTRODUCTION 2

AIM AND SCOPE OF THE THESIS 2 OUTLINE OF THE THESIS 3

CURRENT APPROACHES TO PREVENT AND MITIGATE ACID ROCK DRAINAGE FORMATION 4

SPECIFIC CHARACTERISTICS OF WASTE ROCK 4 SULFIDE OXIDATION AND ACID ROCK DRAINAGE FORMATION 4 ACID BUFFERING REACTIONS 6 ATTENUATION PROCESSES 7 ACID ROCK DRAINAGE TREATMENT 8 METHODS FOR ACID ROCK DRAINAGE PREVENTION 9 PASSIVATION OF SULFIDIC SURFACES 10 HYDROUS FERRIC OXIDES 11

STUDY MATERIALS 14

SULFIDIC WASTE ROCK 14 ALKALINE SECONDARY RAW MATERIALS 14 BLAST FURNACE SLAG 14 CEMENT KILN DUST 14 LIME KILN DUST 15 BIOMASS FLY ASH 15

METHODS 16

GEOCHEMICAL CHARACTERIZATION OF THE WASTE ROCK 16 AUTOMATED MINERALOGY 16 WHOLE-ROCK ANALYSIS 16 GEOCHEMICAL CHARACTERIZATION OF SECONDARY RAW MATERIALS 17 XRD 17 THERMOGRAVIMETRIC ANALYSIS 17 SEM-EDS 17 BATCH TESTING 17 KINETIC TESTING 18 EXPERIMENTAL SETUP 18 TRACE ELEMENT DETECTION IN PYRITE 19 LA-ICP-MS 19 GEOCHEMICAL CHARACTERIZATION OF SECONDARY MINERALS FORMATION 20

SEQUENTIAL EXTRACTION 20 GEOCHEMICAL CALCULATION 20 RAMAN ANALYSIS 20

SUMMARY OF THE FINDINGS 22

MINERALOGY OF THE WASTE ROCK 22 TRACE ELEMENT OCCURRENCE IN PYRITE 24 GEOCHEMICAL CHARACTERIZATION OF SECONDARY RAW MATERIALS 25 SULFIDE OXIDATION AND SUBSEQUENT ACID ROCK DRAINAGE FORMATION 28 PREVENTING ACID ROCK DRAINAGE FORMATION 30 MIXTURES OF (GRANULATED) BLAST FURNACE SLAG AND CEMENT KILN DUST 30 GRANULATED BLAST FURNACE SLAG 30 LIME KILN DUST 31 BIOMASS FLY ASH 33 SECONDARY MINERALS FORMATION 36

DISCUSSION 38

CHALLENGES WITH USING SECONDARY RAW MATERIALS FOR PREVENTIVE PURPOSES 38 THE IMPORTANCE OF KNOWING THE MATERIAL 38 LESSONS LEARNED AND OPPORTUNITIES FOR IMPROVING ARD PREVENTION APPROACH 39

CONCLUSIONS 42

REFERENCES 44

PART II

PAPER I - OCCURRENCE AND RELEASE OF TRACE ELEMENTS IN PYRITE-RICH WASTE ROCK

PAPER II - SUITABILITY STUDY OF SECONDARY RAW MATERIALS FOR PREVENTION OF ACID ROCK DRAINAGE GENERATION FROM WASTE ROCK.

PAPER III - PREVENTION OF SULFIDE OXIDATION IN WASTE ROCK BY THE ADDITION OF LIME KILN DUST

PAPER IV - THE USE OF BIOMASS FLY ASH TO PREVENT SULFIDE OXIDATION IN WASTE ROCK

PART I

2

INTRODUCTION

Extraction and processing of mineral reserves produce considerable amounts of waste around the world. Technological innovation in mining and processing technology has increased extrac-tion efficiency. However, most of the material mined to reach the coveted ore is still discarded as waste, resulting in several Gtons of mine waste annually (Hudson-Edwards et al., 2011). The quantities and complexity of mine wastes are expected to increase significantly due to the ex-ponential increase in demand for innovative critical metals for green technology development. Until the day there is a financial incentive or technical ability to reuse, recycle, or re-mine sul-fide-rich waste rock, it must be stored on-site indefinitely. However, even adequately stored waste rock can impact ecological, hydrological, geotechnical, and environmental aspects such as ground- and surface water, and soils, to mention a few, and the deterioration often occurs gradually, with an example being acid rock drainage formation. The origination of ARD begins with the oxidation of sulfide-bearing mine waste and is considered one of the most problematic environmental problems related to mining (Kefeni et al., 21017) which may persist for hundreds to thousands of years once initiated (INAP, 2021; Höglund et al., 2003). The reactivity of sul-fidic mine waste limits its ability to be reused and recycled unless the sulfides are exploited. A need for alternative methods to reduce the oxidation and reactivity of these waste types is there-fore needed. Traditional remediation methods include active treatment of drainage water during operation followed by dry covering or subaqueous disposal during the decommissioning of the mine. However, this strategy is not suitable for reactive waste since the barrier is often applied too late at the end of mine life. Therefore, new strategies to limit ARD formation during oper-ation need to be developed.

Passivation of sulfide surfaces is a novel inhibition technique for controlling ARD formation by creating a coating on the sulfide surface capable of protecting against oxidation (Zhang and Evangelou, 1998). Several additives for enhancing surface coatings have been explored, both organic and inorganic in solid or liquid form, sprinkled, mixed, or layered on top (Park et al., 2019). Despite promising results, methods for passivation are still only in the development stage as studies are primarily dominated by small-scale laboratory tests focusing on single mineral systems. Larger-scale testing on more complex sulfide-rich materials such as waste rock is therefore needed to ensure sustainable mine waste and water management.

Many reagents used for passivation are associated with high costs and could be potentially harmful to the environment if not stable in the long term. A way of avoiding this could be the use of alkaline secondary raw materials (SRM) to substitute virgin materials. When combined with reactive mine waste, they can form hydrous ferric oxides (HFO) on the reactive surface to create a semi-permanent solution for preventing sulfides from oxidizing (Pérez-López et al., 2007).

3

This thesis is based on studies exploring the use of five different SRM (blast furnace slag, gran-ulated blast furnace slag, cement kiln dust, lime kiln dust, and biomass fly ash) as amendments for preventing sulfide oxidation by HFO formation on the reactive surfaces in sulfide-rich waste rock.

Aim and scope of the thesis

This thesis aims to reduce sulfide oxidation and the subsequent formation of ARD in sulfidic waste rock using secondary raw materials to enable precipitation of hydrous ferric oxides on the reactive surfaces. The focus was primarily to reduce oxidation during a time frame for the mining operations. Furthermore, this thesis aimes to reduce the consumption and transport of virgin natural resources and favor secondary raw materials for a circular approach but not ex-clude aspects such as valorization or green economy.

The specific objectives of this thesis were to:

• Geochemically characterize the waste rock and identify minerals hosting selected traceelements.

• Identify SRM suitable for passivation purposes.• Investigate the extent of sulfide oxidation decrease by addition of SRM and the geo-

chemical process behind• Estimate the sequestration of trace elements and their mobilization in changing chemical

conditions.

Outline of the thesis

This thesis is based on studies brought forward in four appended papers (Papers I-IV). Part I relate to the main findings described in the papers to the objectives defined above. Background of the waste rock characteristics, ARD generation, and methods for preventing ARD formation is provided in Chapter 2. The materials used, along with a summary of methods, are described in Chapter 3. The main findings from the studies are summarized in Chapter 4, discussed and considered in a broader context in Chapter 5, and concluded in Chapter 6. Part II of the thesis consists of the appended papers.

4

CURRENT APPROACHES TO PREVENT AND MITIGATE ACID ROCK DRAINAGE FORMATION

Specific characteristics of waste rock

Waste is defined as unwanted matter or material of any type, especially what is left after useful substances or parts have been removed (Cambridge University Press, 2021). For mine waste, this definition means all types of non-economic material produced within the mine workings. On a global scale, mine wastes represent one of the largest waste sources by volume (Hudson-Edwards et al., 2011). On a regional scale, in Sweden, 105 Mton mine waste was generated in 2018 (Swedish Geological Survey, 2019), comprising approximately 75% of the total amount of waste generated in the country (Naturvårdsverket, 2020).

Regardless of the mineral commodity, deposit type, or extraction method, mine waste primarily constitutes tailings and waste rock (Hudson-Edwards et al., 2011) at approximately equal amounts (Swedish Geological Survey, 2019). The former is a residue resulting from milling and processing the ore into a metal concentrate, and the latter consists of non-valuable rock, which needs to be removed to access the profitable ore. Based on its definition, waste rock is commonly a very heterogeneous material comprising various rocks, sedimentary, metamorphic, or igneous. Due to blasting, the rock contains a large spread of particle sizes, varying from clay to boulder size fragments compared to silty tailings (Lottermoser, 2010).

Storage practices for tailings and waste rock substantially differ from each other, causing changes in geochemical processes and transport mechanisms. Tailings slurries are commonly pumped into tailings ponds where particulates consolidate, causing limited ambient exposure. In contrast, waste rock is commonly stored underground or in stockpiles close to the mine workings if backfilled is not possible. The stockpiles are porous and hydraulically unsaturated. Therefore, the waste rock is relatively exposed to the atmosphere (Vriens et al., 2020). The fine-grained characteristics of the tailings results in a larger surface area that (depending on miner-alogy) can increase geochemical reaction rates, such as sulfide oxidation (Hollings et al., 2001). Large variation in waste rock particle size can result in higher permeability causing non-uni-form hydrodynamic behavior allowing for faster transport of oxygen and water, which can sub-stantially increase sulfide oxidation (Amos et al., 2015).

Waste rock from non-ferrous mining often contains a non-negligible fraction of sulfide minerals and can generate negative environmental impact through ARD formation, where heterogeneity, type, and amount of non-negligible sulfides and storage are a few of the underlying factors determining the size of the problem. Although some waste rocks may be highly reactive (Nys-tröm et al., 2021, 2019a, 2019b; Ragnvaldsson et al., 2014; Alakangas et al., 2012), even low-sulfidic waste rock can generate ARD.

5

Sulfide oxidation and acid rock drainage formation

Sulfide-bearing minerals occur in small amounts in the Earth's crust with local enrichment through mineralization. These minerals are formed and remain stable in environments in which oxygen is absent. When subjected to oxygen and water, sulfide minerals will dissociate and form an acidic, metal(loid) and sulfate-rich water denoted as acid rock drainage (ARD). Alt-hough ARD generation naturally occurs in many sulfide-containing rocks exposed to oxygen, the process is often slow. However, anthropogenic activities such as mining can accelerate its generation due to displacement and downsizing of material, subjecting an increasing amount of sulfides to the atmosphere. The most abundant sulfide mineral is pyrite (Vaughan, 2021), and its oxidation is described by Singer and Stumm (1970) through the following reactions:

FeS2 + 72

O2 (aq) + H2O ↔ Fe2+ + 2SO42− + 2H+ (1)

Fe2+ + 14

O2 + H+ ↔ Fe3+ + 12

H2O (2)

Fe3+ + 3H2O ↔ Fe(OH)3 + 3H+ (3)

The overall reaction is written as:

FeS2(s) + 154

O2(g) + 72

H2O(l) ↔ Fe(OH)3(s) + 2SO4(aq)2− + 4H+ (4)

At near neutral pH, pyrite's oxidation driven by atmospheric oxygen produces Fe(II), sulfate, and acidity (Reaction 1). Ferrous iron (Fe(II)) can, can then be oxidized into Fe(III) (Reaction 2) followed by its hydrolysis into various forms of Fe (oxy)hydroxides and/or (oxy)hydroxy-sulfates depending primarily on pH, redox potential, and sulfate content of the solution(Nordstrom, 1982). These phases can be grouped as hydrous ferric oxides (HFO), but are hereshown as ferrihydrite, Fe(OH)3 (Reaction 3). In reality, pyrite's oxidation is a complex bio-chemical process involving several redox reactions with a few electrons' moving at a time, cat-alyzed by microbial activity (Nordstrom and Alpers, 1999; Nordstrom, 1982). Herein the over-all reaction is described according to reaction 4. At low pH (<3.5), the hydrolysis of Fe(III)dramatically decreases, leaving Fe(III) as an additional oxidant capable of speeding up the ox-idation and forming important catalytic feedback (Figure 1, Reaction 5).

FeS2 + 14Fe3+ + 8H2O ↔ 15Fe2+ + 2SO42− + 16H+ (5)

The overall oxidation rate rapidly decreases in the absence of oxygen. Iron(III) oxidation of pyrite is much faster than that by oxygen, and chemical conditions favoring Fe3+ should there-fore be prevented. However, it is also faster than the oxidation of Fe(II) to Fe(III) (Nordstrom, 1982). Despite Fe(III) 's catalytic effect on sulfide oxidation, the rate of abiotic pyrite oxidation is described as slow, especially at low pH connected to the rate-limiting step of Fe(II) oxidiza-tion (Singer and Stumm, 1970). The presence of organisms can accelerate oxygenation (Singer and Stumm, 1970). Consequently, Fe(III) as the main oxidant has a substantial impact on ARD generation due to its ability to generate four times the acidity compared to oxygen (Reaction 5).

6

In general, pyrite oxidation releases Fe, SO42-, and acidity into solution. However, trace ele-

ments present in the pyrite such as As, Cu, Hg, Pb, Sb, Tl, and Zn (Abraitis et al. 2004) and other weathered minerals will determine the chemistry of the ARD.

Due to the amount of waste rock produced, its potential sulfide content, and exposure to ambient conditions, it may form large volumes of ARD if deposited without applying preventive measures. Moreover, new extraction methods and an increasing need for metals and minerals lead to the mining of more complex and low-grade ores, resulting in larger volumes of increas-ingly complex mine waste, further stressing the need for early preventive measures.

Acid buffering reactions

Carbonate minerals (e.g., calcite), Al- and Fe-(oxy)hydroxides (e.g., gibbsite and HFO), and silicate minerals (e.g., feldspars, chlorites, and micas) can counteract ARD formation through their ability to consume acids when dissolving. The combination of neutralizing minerals tends to create a buffering sequence describing a specific pH interval within which they are typically depleted. Carbonates are consumed around pH 6-7, whereas Al- and Fe- (oxy)hydroxides typi-cally buffer acidity around pH 4-5 and 3-4, respectively (Vriens et al., 2020; Lottermoser, 2010; Salomons, 1995). Silicates are known to buffer acidity over a wide pH range. However, their ability to buffer acidity at low pH is highly limited (Plumlee, 1999). Acid-buffering minerals can dissolve alongside progressing sulfide oxidation when they are naturally co-located in the waste rock matrix, creating neutral drainage with low metal(loid) concentrations (Vriens et al., 2019). However, that is rarely the case since buffering mineral dissolution is generally faster than sulfide oxidation (Lottermoser, 2010), which is one reason why ARD formation can take decades to develop.

Figure 1. Simplified illustration describing the oxidative dissolution of pyrite by oxygen and ferric iron through catalytic feedback. Modified from Vriens et al. (2020).

FeS2

Fe3+

O2

H2O

O2

Fe2+

H+SO42-

7

Attenuation processes

Attenuation processes such as adsorption and secondary minerals formation can largely influ-ence drainage quality through differences between whole-rock elemental composition and ob-served drainage loads. (Nordstrom, 2011).

Geochemical adsorption refers to the attachment of aqueous solutes, such as anions, cations, and complex ligands, to mineral surfaces. Many minerals have a negatively charged surface within normal pH ranges. Therefore, these surfaces can attract cations, rendering the mineral "neutralized". However, these attracted ions can bond to the mineral surface in different ways by relatively strong covalent bonding (inner-sphere), weaker electrostatic attraction (outer-sphere), or simply by residing close to the surface (diffusive) (Figure 2). It is important to note that whatever ions are adsorbed, they can be substituted. This process is commonly known as ion exchange.

Various approaches have been taken to categorize the types of minerals in mine waste. Herein, classification, according to Jamieson et al. (2015), is used by defining primary minerals as those existing pre-mining and secondary as those formed post-mining. Secondary minerals often in-clude water-soluble salts, (oxy)hydroxides, (oxy)hydroxysulfates, (oxy)hydroxycarbonates, phosphates or arsenates depending on primary waste rock mineralogy and chemistry, but also physical and hydrological conditions of the waste rock piles (size, particle size, hydraulic re-tention time, etc.) (Vriens et al., 2020).

Figure 2. A schematic representation of a mineral surface illustrating differences in adsorption types. Modified from Zhu and Anderson (2002).

O

O

Si O

O

Si O

O

Si O

O

Si O

OOCa

Ca

O

O O

O

O

O

Ca OOInner-sphere

Outer-sphere

Diffusive

Bulk-water

8

Secondary minerals are often seen as distinct rims or coatings on weathered particles. The het-erogeneity of waste rock often results in the mobilization of elements at one location followed by its subsequent precipitation elsewhere in, or adjacent, or at a distance from the waste rock pile. Although secondary precipitates such as HFO are common (Nordstrom et al., 2015), it still takes time for them to accumulate into detectable levels, and therefore, they are more commonly inferred by geochemical calculations and modeling rather than being analytically identified. Nevertheless, even a low abundance of secondary mineral phases, such as HFO, can be im-portant sinks through their ability to adsorb and co-precipitate with a wide range of metal(loid)s acting as an attenuation sink.

Acid rock drainage treatment

Acid rock drainage develops when neutralization capacity is absent, depleted due to ongoing sulfide oxidation, or when carbonate dissolution is inefficient due to limited contact between the neutralizing and acid-producing minerals.

Numerous strategies have been developed to manage ARD, where the most common treatment method is through the active addition of chemicals. Alkaline reagents, such as Ca(OH)2, CaO, or NaOH, raise the drainage's pH, which together with aeration allows for Fe(II) oxidation and precipitation of Me-(oxy)hydroxides removed through settling and filtration systems. Lime (CaO) and hydrated lime (Ca(OH)2) are considered the most efficient neutralizing materials, and thus, are the most widely used for ARD treatment (Johnson and Hallberg, 2005; Brown et al., 2002; Younger et al., 2002). However, most chemicals used are commercially manufactured

Figure 3. A simplified and non balanced schematic figure illustrating the formation of hydrous ferric oxides (HFO) coatings on pyrite. I: Oxygen reduction, II: Pyrite dissolution and sulfate oxidation, III: Ferrous oxidation, IV: Formation of HFO through the hydration of ferric iron. Modified from Park et al. (2019).

FeS2

HFO

H2O

Fe2+

SO42-

H+

Fe3+

O2

H2O

H+

H2O

H+

O2

H2O

e-

e-

III

III

IV

I

9

and can be rather expensive depending on the type (Acharya and Kharel, 2020; Kefeni et al., 2017). In recent years, research has been looking into exchanging some of these alkaline chem-icals for secondary raw materials since their use can promote valorization as cost-effective al-ternatives. Studied SRMs include, but not limited to, cement kiln dust (Sulaymon et al., 2015; Mackie and Walsh, 2012; Doye and Duchesne, 2003), lime kiln dust (Tolonen et al., 2014), coal fly ash (Jones and Cetin, 2017; Madzivire et al., 2014), and paper mill residues (Alakangas et al., 2013; Pérez-López et al., 2011). However, regardless of the amendment choice, active treatment is based on the precipitation of various secondary minerals resulting in the formation of a secondary waste in the form of a sludge that needs to be disposed of additional treatment and deposition (Acharya and Kharel, 2020).

Despite its success in limiting overall adverse effects, treatment methods are not long-term so-lutions since they require continuous maintenance with an incessant addon of chemicals, energy consumption, and long-term monitoring until oxidation has ceased. Thus, they do not apply to the "walk away principle" in Sweden, which refers to the need for a long-term stable solution capable of self-managing before the mine site can be abandoned.

Methods for acid rock drainage prevention

During the last decades, numerous methods for preventing sulfide oxidation and subsequent ARD formation have been developed (Park et al., 2019). Due to the nature of ARD formation, solutions to prevent it lies within methods to control Fe(II) oxidation. Strategies for such control resides in either kinetic or thermodynamic control.

Kinetic control requires the suppression of the catalytic agent responsible for the increased ox-idation rate of Fe(II), namely bacteria. Bactericides work by disrupting bacteria's contact with mineral surfaces or harming their protective cell membrane (Baker-Austin and Dopson, 2007), which allow them to function in an acidic environment by using, e.g., anionic surfactants (Evan-gelou, 1995; Zhang and Wang, 2017) or organic acids (Baker-Austin and Dopson, 2007). How-ever, bactericides are not permanent solutions as they are commonly water-soluble and easily transported away from the sulfide surfaces, making them most applicable when there is an im-minent need of suppressing the sulfide oxidation since repetitive treatments are needed (Klein-mann, 1999). Moreover, Sand et al. (2007) reported that bactericides do not kill the bacteria completely, it only reduces their activity and numbers. On a final note, since bactericides are harmful to bacteria, they can also be harmful to other living organisms (Hodges et al., 2006; Liwarska et al., 2005).

Thermodynamic control is based on the total elimination of oxygen and creating reducing con-ditions in the mine waste. One of the most commonly used techniques is applying oxygen bar-riers such as dry covers or subaqueous disposal. Oxygen has a lower solubility and diffusion in water than in the air, resulting in that mine waste underwater can reduce redox conditions. Backfilling of waste rock is commonly performed during the decommissioning of a mine where flooding of and sealing above the waste rock is a common approach in Sweden (Villain, 2014). On the other hand, dry covers are commonly constructed by till of varying quality, which can have large discrepancies in its construction, but the common denominator is creating a high

10

degree of saturation in the sealing layer above the mine waste to reduce water percolation and O2 influx. Recent studies have examined the potential to exchange virgin natural resources with SRM such as green liquor dregs (Sirén et al., 2016; Mäkitalo et al., 2015, 2016), fly ash (Demers et al., 2017; Lu et al., 2013; Hallberg et al., 2005) and cement kiln dust (Duchesne and Doye, 2005).

Passivation of sulfidic surfaces

Whereas physical oxygen barriers have been adopted for centuries in various types and man-ners, a more recent approach is inhibition, passivation, and microencapsulation, which are more or less synonymously used in the literature and involve creating a chemical barrier particle surface to prevent oxidation.

There are several types of passivation methods, each utilizing different amendments for creating coatings ranging from organic to inorganic or a combination of the two. Coatings for pas-sivation can consist of different type of ferric minerals such as ferric phosphates (Kollidas et al., 2019; Evangelou 2001, 1995; Fytas and Evangelou, 1998; Vandiviere and Evangelou, 1998; Nyanvor and Egiebor, 1995; Huang and Evangelou, 1992) or ferric hydroxide-silica (Kollidas et al., 2018, 2015; Fan et al., 2017; Kang et al., 2016; Evangelou 2001; Vandiviere and Evan-gelou, 1998; Zhang and Evangelou, 1998; Evangelou, 1996).

Unfortunately, most of these methods require pre-oxidation with H2O2 resulting in high costs and difficulty to use in practice (Ouyang et al. 2015). Also, amendments such as phosphate and organics can be potentially harmful to the environment when entering water bodies, leading to eutrophication (Khummalai and Boonamnuayvitaya, 2005).

Other methods such as permanganate (Ji et al., 2012; Misra et al., 2006; De Vries. 1996) and organic coatings have been proposed where the latter renders the reactive surfaces hydrophobic, which limits the water-mineral interactions and reduce the amount of area exposed to oxidation (Ačai et al., 2009; Kargbo et al., 2004; Elsetinow et al., 2003; Zhang et al., 2003; Jiang et al., 2000; Belzile et al., 1997; Lalvani et al., 1990). However, the stability of organic coating has been questioned since microorganisms can degrade these complex organic compounds, render-ing the coating inefficient.

Silane-based coating combines an inorganic and organic component to enhance durability and adhesion but remains flexible and crack resistant (Ouyang et al., 2015; Diao et al., 2013; You et al., 2013; Khummalai and Boonamnuayvitaya, 2005). Unfortunately, relatively large amounts of these amendments must be used to avoid cracking. Therefore, recent studies include mixing with nano- or microparticles (Gong et al., 2021; Li et al., 2021; Liu et al., 2017)

A technique called carrier microencapsulation was developed to target sulfides in complex min-eral systems specifically. The method combines a redox-sensitive organic compound with metal(loid) ions into a soluble metal(loid) and organic complex capable of selectively decom-posing on sulfide minerals and precipitate as metal(loid) oxyhydroxides (Li et al., 2019; Park et al., 2021, 2020, 2018a, 2018b; Tabelin et al., 2019; Yuniati et al., 2015a, 2015b; Jha et al., 2008, 2011, 2012; Satur et al., 2007). Due to its ability to selectively target sulfides, it has been

11

suggested that it could treat low sulfide mine waste without the excess use of reagents (Park et al., 2019).

However, despite promising results, methods for passivation often rely on the complete sub-mersion of the reactive mineral in the solute in question. These studies are still under the devel-opment stage. However, if applied on waste rock at a larger scale, a new strategy for waste rock deposition would likely need to be adopted, including milling and processing the material.

An alternative method is the co-disposal and blending of either acid-consuming or alkaline generating materials with reactive mine waste to enhance HFO formation on the reactive sur-faces. Various SRM have shown promising results in inhibiting the sulfide oxidation and cost-effective alternatives to virgin materials such as limestone or lime (Alakangas et al., 2013; Sa-hoo et al., 2013; Pérez-López et al., 2007, 2009).

Hydrous ferric oxides

Pyrite oxidation in a wide pH range can form various HFO minerals. If the oxidation occurs at near-neutral pH and in the presence of enough alkalinity, HFO formation is intensified and can develop coatings on the surface, synonymous with rims (Figure 3). If these conditions are up-held for a more extended period, these coatings will grow thicker, densify and eventually reduce the rate of pyrite oxidation. When pyrite oxidizes under these conditions, acidity is neutralized near the mineral surface. Simultaneously, the rapid oxidation of Fe(II) to Fe(III) followed by hydration allows for speedy precipitation into poorly crystalline HFO colloids. Huminicki and Rimstidt (2009) suggested that the process of HFO coating formation on pyrite surfaces consists of two stages (Figure 4).

At operating conditions (near-neutral pH with sufficient alkalinity), the HFO colloid is oppo-sitely charged to the pyrite surfaces, which allow for attraction and attachment of the HFO colloid on the pyrite surface forming a thin, porous, and permeable layer (Stage I, Figure 4). At this stage, the coating is still relatively ineffective in preventing further oxidation. However, with time, the coating grows thicker with the continuous release of Fe(II) by pyrite oxidation and subsequent precipitation of HFO colloids between previously attached HFO colloids, de-creasing not only the porosity but also permeability, lowering the oxidation rate (Stage II, Fig-ure 4).

The creation of HFO on reactive sulfide surfaces has been demonstrated by, e.g., Nicholson et al. (1990), who showed that the HFO layer accumulates in carbonate solution with near-neutral pH, which substantially lowered the oxidation rate. The effectiveness of the approach is sup-ported by Holmström et al. (1999), using humidity cell testing. An advantage to other methods, not relying on HFO formation for preventing purposes, is that, given the right conditions, HFO will form naturally, are self-healing, and become more effective with time, decreasing the need for alkalinity supply (Huminicki et al., 2009).

12

Figure 4. Schematic diagram showing the steps of hydrous ferric oxide (HFO) coating formation on pyrite. Stage I shows the initial deposition of a porous and permeable HFO coating by the attachment of HFO colloidal particles. Stage II shows the transition from reaction-limited to diffusion-limited rates as the coating becomes denser and thicker resulting from the HFO precipitation in space between HFO colloids. Modified from Huminicki and Rimstidt (2009).

O2

Fe2+ O2

Fe3+ HCO3-

HFO colloid

HFO

≡

HFO

Rea

ctio

n-lim

ited

oxid

atio

nC

ollo

id fo

rmat

ion

and

atta

chm

ent

STAG

E I

HFO

HFO

HFO

HFO

STAG

E II

Fe2+ O2

Fe3+ HCO3- colloid

overgrowth

HFO

HFO

HFO

HFO

Col

loid

cem

enta

tion

and

over

grow

thG

row

th o

f coa

ting

inw

ard

O2 diffusion-limited oxidation

14

STUDY MATERIALS

Sulfidic waste rock

The sulfidic waste rock originated from a Boliden Mineral AB-owned base- and precious metal (copper, gold, silver, and zink) open-pit mine in northern Sweden. The ore deposit is a volcanic-associated massive sulfide (VMS) deposit hosted in the Skellefte group with an approximate age of 1.89 Ga (Montelius, 2005). Production began in the year 2000, after which 9.9 Mtons of waste rock have been generated with the expectation that 10 Mtons of waste rock will be present at the end of the mine's life. In total, 9.3 Mtons is predicted to be potentially acid-producing. Alakangas et al. (2014) previously described the waste rock sampling, which included the screening of waste rock piles with a handheld X-ray Fluorescence (XRF) from Olympus Innov-x systems, USA, to select waste rock with high sulfur content. Alakangas et al. (2014) also characterized the waste rock for its major and minor element content showing an average sulfur content of ca. 30%, and Nyström et al. (2017) showed that the waste rock was "potentially acid-producing" with an average net neutralization potential (NNP) of -946 kg CaCO3/ton waste rock.

Alkaline secondary raw materials

SRMs range from wastes to commercially established (by)-products with the common denom-inator of being salvaged materials that can be substituted for virgin raw materials. Their use can promote valorization and be cost-effective alternatives.

Blast furnace slag

Blast furnace slag (BFS) originates from the manufacturing of crude iron, whereby iron ore is added to a blast furnace with coke (reduction agent) and limestone (slag former) to remove impurities. The residues, mainly coke ash and non-metallic components, are removed by chem-ically combining them into a liquid slag. The slag can be air-cooled (BFS), resulting in a pre-dominantly crystalline material, or water granulated (GBFS), resulting in amorphous sand-like material (0-4 mm). Both BFS and GBFS are known to have excellent geotechnical properties. For these reasons, BFS is often used in road construction. In contrast, ground GBFS is mainly used as a supplementary cementing material blended with cement for concrete production. The BFS has been crushed to sizes 0-4 mm before testing. MEROX AB supplied both the BFS (Luleå) and GBFS (Oxelösund).

Cement kiln dust

Cement kiln dust (CKD) originates from the manufacturing of cement, whereby CaCO3 and Si-, Al- and Fe-oxides are added to a rotary kiln and heated to a temperature of 1450 oC to form alite (main mineral in Portland cement) (Hökfors, 2014). During the heating, flue gasses are emitted, generating CKD from stripping the gasses from dust. CKD is primarily used as a com-ponent in cement. Cementa AB supplied the CKD used in this study.

15

Lime kiln dust

Lime kiln dust (LKD) originates from lime (CaO) manufacturing, whereby limestone is added to a rotary kiln and heated to temperatures up to 1200-1300 oC. During the heating, flue gasses are emitted, generating LKD from stripping the gasses from dust. Some quicklime manufactur-ers make briquettes out of LKD, whereas others deposit it in piles/silos on-site. LKD can also be mixed with varying amounts of crushed limestone to produce niche products. Nordkalk sup-plied the LKD, a mixture of partially calcined material and finely crushed limestone (too fine for the kiln). The LKD had been stored in piles outside.

Biomass fly ash

Biomass fly ash (BA) originates from wood pulp manufacturing, one of the principal compo-nents in paper. Wood is washed and debarked before being chopped into wood chips and di-gested in the Kraft production. The bark is combusted for energy, resulting in an ash material from stripping the flue gases from dust during the incineration. BillerudKorsnäs supplied the BA used in this study, which is usually landfilled. The BA was fresh and dry.

16

METHODS

The experimental design, sampling, analytical, and data interpretation methods are described briefly in this section. Detailed information can be found in the respective article denoted for each specific method.

Geochemical characterization of the waste rock

Automated mineralogy (Paper I)

Polished uncovered thin sections of crushed waste rock (5-30 mm sieve size) were prepared by Vancouver Petrographics Ltd., Vancouver, BC, Canada, with special care taken not to dissolve water-soluble phases in the sample. An optical mineral examination was performed on polished thin sections in reflected and transmitted light using a standard Nikon Eclipse E600POL petro-graphic microscope (Nikon Instruments INC., Tokyo, Japan).

Automated quantitative mineralogical characterization by QEMSCAN® 650 with two Bruker EDX detectors was performed at Boliden Mineral AB using their mine site database as a refer-ence. A minimum of 28 waste rock pieces (5-30 mm sieve size) were analyzed. QEMSCAN step size was set to 6 μm as a compromise between the number of analysis points yielded and the time for analysis (approximately seven hours/thin section).

Whole-rock analysis (Papers I-IV)

The total element content of the waste rock was determined using a combination of methods due to apprehension regarding the overall low Cu content in the waste rock. The results were compared with those of Alakangas et al. (2014), showing a similar spread in element concen-trations.

i. Screening of more than 70 elements performed by using Inductively Coupled PlasmaMass Spectroscopy (ICP-MS) by the SVEDAC-accredited ALS Scandinavia laboratoryin Luleå, Sweden. Total element concentrations were analyzed after lithium borate fu-sion and three acid digestion (nitric acid, hydrochloric acid, and hydrofluoric acid).

ii. ALS Brisbane, Australia, analyzed the three waste rock samples. Two waste rock sam-ples were analyzed with XRF after lithium borate fusion containing 20% sodium nitrateas an oxidizing agent. One sample was analyzed by Inductively Coupled Plasma AtomicEmission Spectroscopy (ICP-AES) after being fused with sodium peroxide and dis-solved in diluted hydrochloric acid.

iii. ALS Loughrea, Ireland, analyzed waste rock samples using ICP-MS or ICP-AES. Totalelement concentrations were analyzed after lithium borate fusion and two acid digestion(nitric acid and hydrochloric acid).

17

iv. Sulfidic sulfur was determined using 25% HCl leach followed by leco furnace melt.ALS Vancouver, Canada, analyzed the samples by inductively coupled plasma opticalemission spectroscopy (ICP-OES).

Geochemical characterization of secondary raw materials

XRD (Papers II, III, and IV)

Mineralogical analysis of the BA, CDK, and LKD was performed using X-ray powder diffrac-tion (XRD) and recorded with a PANalytical Empyrean diffractometer run in Bragg-Brentano geometry with CuKα radiation (λ=1.5406Å). The samples were scanned over the 2Ɵ range of 5-90° with a step size of 0.0130°. The scan time was set to 47 minutes.

Thermogravimetric analysis (Paper III)

Due to the relative pureness of LKD, minerals present in the material were quantified using thermogravimetric measurements by heating the material to 1000 °C in an inert argon gas using a NETZSCH STA 409 C/CD. Mineral content was calculated by coupling XRD results with thermogravimetric curves for each mineral (Földvári, 2011).

SEM-EDS (Paper IV)

For the BA, minerals were confirmed using a ZEISS Sigma 300 VP scanning electron micro-scope (SEM) with two Bruker XFlash 6160 energy-dispersive spectroscopy detectors (EDX). Due to the material's fine-grained character, the primary electron beam's acceleration voltage was set to 10-15 kV to reduce the interaction volume. The aperture was set to 30 µm to avoid image distortion.

High vacuum can alter minerals such as ettringite, commonly present in materials such as BA. Therefore, a sample of wet leached BA was investigated in low vacuum (25Pa) using Scanning Electron microscope, type Jeol JSM-IT100 combined with EDS analyzer, type Bruker (JEOL Nordic AB, Sollentuna, Sweden). The acceleration voltage was set to 18kW.

Batch testing (Paper II)

The SRM were subjected to batch testing modified from Swedish standard SS-EN 12457-2 (SIS, 2003) for prediction of neutralization potential and determination of the readily soluble elements. No downsizing of the materials was performed, and blank tests did not exceed 20% of the concentrations determined in the eluate of the tested wastes. According to the standard, conducting a batch test for 24 h is needed to dissolve all water-soluble fractions in a material. However, there is a risk that the solution may reach an equilibrium that prevents dissolution (SIS, 2003). Therefore, more complete dissolution was investigated by extending the batch test to 72 h (3 days, L/S 12) and 600 h (25 days, L/S 14), preferentially to conducting a two-stage batch test (L/S 2 and 10). After 24 h rotation on an end-over-end rotating device, the samples

18

were decanted, and 20% of the volume was replaced with MilliQ water. The procedure was repeated at 72 h (L/S 12) and 600 h (L/S 14).

Kinetic testing

Experimental setup (Papers I-IV)

Kinetic testing was conducted in small-scale test cells to reproduce the oxidation conditions of waste rock and promote secondary mineral formation for passivation of the sulfidic surfaces. The experimental design consisted of eight high-density polyethylene cells with surface areas of 513 cm2 and a total volume of 10 L (Figure 5). The cells were filled with crushed waste rock (5-30 or 30-60 mm) and irrigated weekly with 600 mL of MQ water (0.05 μS/cm), correspond-ing to average annual precipitation in the area a yearly L/S ratio of approximately 4. The bottom of each cell was lined with geotextile to avoid clogging of the tap. The waste rock in all cells was leached for up to eight weeks before adding SRM (Table 1). A geotextile liner was used between the larger waste rock (30-60 mm) and SRM to avoid downward movement of the SRM. Mixing of CKD with BFS and GBFS respectively was performed by hand where the CKD's hydrophilic character enabled a more homogenous mixture without any obvious layering of the materials. Cell A was a control cell filled with solely waste rock (7.55 kg) and acted as a refer-ence displaying the evolution of ARD production without any preventive measures for pas-sivation.

Table 1. Differences in leaching conditions in the eight small-scale test cells filled with high sulfidic waste rock and varying additions of lime kiln dust (LKD), blast furnace slag (BFS), granulated blast furnace slag (GBFS), and biomass fly ash (BA)

Cell Waste rock LKD BFS GBFS CKD BA

5-30 mm 30-60 mm wt.% wt.% wt.% wt.% wt.%

A 7.75 kg Reference

B 7.75 kg 5

C 7.75 kg 4 1

D 7.75 kg 4 1

E 11 kg 5

F 11 kg 1+1.5

G 11 kg 2.5

H 11 kg 5

19

Sampling of leachates

Leachates were collected from the tap located at the bottom front of the cell weekly the day after irrigation. The collected leachates were measured for pH, electrical conductivity (EC), and temperature connected with the sampling in closed containers to avoid exposure to air. The pH and EC were measured using a WTW Multi 3420 multimeter equipped with Sentix® 940 (pH) and TetraCon® 925 (EC) electrodes. Leachate samples were filtrated through a 0.22-μm nitro-cellulose membrane into high-density polyethylene bottles using vacuum filtration. Washing of filters and filter equipment was performed in 5% acetic acid and 5% nitric acid, respectively. Filtrated samples were stored cold (4 °C) and in darkness until analysis.

Analysis of dissolved elements and chloride

Before analysis, the samples were acidified with 1 mL nitric acid (suprapur) per 100 mL sample. Selected samples were analyzed for major and trace element composition using inductively coupled plasma atomic emission spectroscopy (ICP-AES) and inductively coupled plasma sec-tor field mass spectrometry (ICP-SFMS) at the SVEDAC accredited laboratory ALS Scandina-via in Luleå. The analysis was performed according to US EPA Method 200.7 (modified) and 200.8 (modified) or by quantitative screening analysis for over 70 elements. Chloride and flu-oride concentrations were determined using ion-selective electrodes.

Trace element detection in pyrite

LA-ICP-MS (Paper I)

Trace element occurrence and distribution in pyrite grains were performed on thin sections of the initial waste rock and grab samples after one and two years of leaching (week 52 and 103). An ESI NWR193-nm excimer laser microprobe coupled to a Thermo Scientific iCAP Q quad-rupole ICPMS at Luleå University of Technology, Sweden, was used. Analyses on pyrite were performed by laser-ablating spots of 35 μm in diameter, a repetition rate of 5 Hz, and laser beam

Figure 5. Experimental design of small-scale test cells filled with (A) sulfidic waste rock, (B-H) sulfidic waste rock with 1-5 wt.% secondary raw material.

20

energy maintained between 3 and 4 Jcm−2. Analysis time was restricted to 56 s, comprising 20 s background (laser shutter closed) and 100 laser bursts/36 s analysis (laser shutter open). Im-aging of pyrite was performed by ablating a set of parallel lines arranged in a grid over the sample where line space corresponded to beam size. Beam size was 10 μm2

to investigate zoning in pyrites of less than 500 μm2. The beam was rastered over the lines at half speed of the beam size (i.e., spot = 10 μm2, speed = 5 μm/s−1) at 5 Hz repetition rate. Acquisition times were as follows: major silicate elements (i.e., Mg, Al, Si, Ca) and S = 0.005 s; major elements (i.e., Fe, Ni) = 0.005 s; Ag and Au = 0.5 s; other elements = 0.01 s; with a total sweep time of approxi-mately 0.5 s. Data reduction was carried out using Iolite4 software (Paton et al., 2011). Iron was used as the internal standard (assuming stoichiometric Fe of pyrite). The primary calibra-tion standard used was NIST 612 (Pearce et al., 1997) with UQAC FeSb (Savard et al., 2018), MASS1, USGS (Wilson et al., 2002), and NIST 610 (Pearce et al., 1997) used as secondary standards. Standards were analyzed every one and a half hours to correct for instrument drift.

Geochemical characterization of secondary minerals formation

Sequential extraction (Papers I, III, and IV)

Sequential extraction by Dold (2003) was used to gain quantitative information on trace element association to secondary minerals. Grab samples of leached and non-leached waste rock were subjected to the seven-step sequential extraction at the SGS laboratory (Pittsburgh, USA). The method is adapted to the specific primary and secondary mineralogy associated with Cu-sulfide ores but has also been applied to other types of mine waste and was chosen due to its ability to differentiate various HFO phases. Whole-rock analysis of the samples was performed by NaOH fusion for comparison with Σ seven-step sequential extraction.

Geochemical calculation (Papers I, III, and IV)

Geochemical calculations, including aqueous species distribution, and mineral saturation state, were carried out using PHREEQC version 3.4.0 (Parkhurst and Appelo, 1999) using the wateq4f.dat thermodynamic database (Ball and Nordstrom, 1991). Ion imbalance for consid-ered samples was within ± 14%. The solubility constants of more ordered ferrihydrites, 2- and 6-line, were used as reported by Stefánsson (2007). For schwertmannite, the solubility constantfrom Majzlan et al. (2004), originating from Bigham et al. (1996), was considered.

Raman analysis (Paper III)

Mineral identification was performed with Raman analysis (532-nm laser excitation) using a Senterra Raman spectrometer connected to an Olympus BX microscope. For primary minerals, the laser power was set to 2 mW with an exposure time of 1 s with five scans and a pixel size of 1 μm. For secondary minerals, the laser power was set to 0.2 mW with an exposure time of 60 s with three scans and a pixel size of 1 μm. Raman analysis was coupled with surface optical imaging using a ZEISS SteREO V.8 stereomicroscope followed by a ZEISS Gemini Merlin scanning electron microscope (SEM).

22

SUMMARY OF THE FINDINGS

Mineralogy of the waste rock

Mineralogical investigation of the waste rock revealed that it was dominated by pyrite (66%) with smaller amounts of quartz (17%), muscovite/sericite (6%), chlorite (4%), and calcite (1%). Other sulfides found at trace amounts in the waste rock included arsenopyrite (FeAsS, 1.8x10-

4 %), bournonite (PbCuSbS3, 4.1x10-4%), chalcopyrite (CuFeS2, 12x10-4%), pyrrhotite (Fe1-xS, 2.4x10-4%), and sphalerite (ZnS, 2.6x10-4%), resulting in a low overall content of As-, Cu-, Pb-, and Zn-bearing minerals in the waste rock (Table 2). An overall high abundance of Fe and S was observed to correlate with pyrite content. Despite an overall low abundance of sulfides other than pyrite, elevated concentrations of As, Hg, Sb, and Tl were observed (Table 3), sug-gesting their preferential occurrence as trace elements in other minerals.

Table 2. Mineralogy of waste rock and secondary raw materials. The quantitative mineralogical composition of sulfide-rich waste rock was determined by quantitative automated mineralogy (QEMSCAN). Cement kiln dust (CKD), biomass fly ash (BA), and lime kiln dust (LKD) were obtained from X-ray powder diffraction (XRD). Quantitative mineralogy of lime kiln dust (LKD) was determined by combining XRD results with thermogravimetry. Lime content of the LKD was determined using the "sugar rapid" method (ASTM 2011). Anhydrite and sylvine content were estimated based on S and Cl concentrations, respectively. The mineralogy of the blast furnace slag (BFS) and granulated blast furnace slag (GBFS) was obtained from Merox (2015).

Waste rock % LKD % BFS GBFS CKD BA Pyrite 66 Calcite 85 Monticellite Glass Akermanite Albite Quartz 17 Mg-rich calcite 8 Akermanite Hercynite Alite Anorthite Muscovite 6 Anhydrite 2 Anhydrite Arcanite Chlorite 4 Gypsum 2 Anorthite Calcite Calcite 1 Lime 1 Arcanite Gehlenite Dravite 0.4 Slaked lime 2 Calcite Halite Kaolinite 0.3 Sylvine <0.2 Gehlenite Lime Albite 0.3 Gypsum Orthoclase Flourite 0.1 Larnite Perovskite Oligoclase 0.09 Lime Quartz Apatite 0.03 Portlandite Slaked lime Anatase 0.01 Quartz Sylvine Clinopyroxene 0.006 Sylvine Allanite 0.006 Anorthite 0.003 Sanidine 0.001 Chalcopyrite 0.001 Biotite 0.0009 Bournonite 0.0004 Sphalerite 0.0003 Arsenopyrite 0.0002 Pyrrhotite 0.0002

23

Table 3 Abundance of selected elements in sulfidic waste rock, lime kiln dust (LKD), blast furnace slag (BFS), granulated blast furnace slag (GBFS), cement kiln dust (CKD), and biomass fly ash (BA). Elements in the secondary raw materials are marked according to their enrichment relative to the waste rock. Elements presented in bold are enriched >3 times the average Earth’s crust (Krauskopf and Bird, 1995).

Waste LKD BFS GBFS CKD BA

Average Waste LKD BFS GBFS CKD BA

Average

% rock earth's crust mg/kg rock earth's crust

Al 2 0.3 7 7 2 4 8 Li 15 3 71 73 22 13 30

Ca 0.8 46 28 25 38 15 3 Lu 0.1 0.05 1 1 0.2 0.08 0.9

Cl 0.01 <0.1 <0.1 <0.1 5 0.4 0.05 Mn 189 178 3426 4961 454 8123 900

F 0.1 0.01 0.07 0.04 0.07 0.03 0.05 Mo 3 0.2 0.5 0.9 6 13 2

Fe 32 0.4 1 1 2 2 5 Nb 1 1 28 15 4 5 20

K 0.4 0.02 0.5 0.6 7 4 2.6 Nd 6 4 66 59 13 8 25

Mg 0.5 0.7 10 11 1 2 2.1 Ni 3 3 4 5 16 75 80

Na 0.3 0.02 0.5 0.5 0.5 2 2.4 Pb 22 3 0.2 0.8 861 51 16

P 0.005 0.003 0.004 0.002 0.02 1 0.1 Pd <0.1 0.1 <0.9 <0.9 <0.3 0.07 0.01

Stotal 30 0.04 1 2 4 1 0.05 Pr 2 0.9 17 15 3 834 -

Si 9 0.7 15 16 7 21 27 Rb 8 2 17 24 475 177 120

Ti 0.03 0.02 2 2 0.2 0.2 0.5 Re <0.01 0.001 0.0001 0.001 0.2 0.002 0.001

mg/kg Rh <0.01 0.01 0.001 0.01 0.02 0.004 -

As 217 0.8 0.3 0.7 9 3 2 Sb 38 0.09 0.1 0.2 5 2 0.15-1

Ag 1 0.02 0.1 0.1 12 0.5 0.07 Sc 5 0.8 30 32 4 4 5-22

Au <0.01 0.01 0.01 0.02 0.2 0.005 0.004 Se <1 0.5 5 6 19 0.9 0.09

B 5 18 83 80 72 202 10 Sm 1 0.8 12 11 3 1 7

Ba 65 22 495 524 179 1398 430 Sn 0.5 0.3 0.4 0.9 5 2 2.5

Be 0.2 0.1 6 7 0.7 1 3 Sr 29 168 496 425 249 583 350

Bi 0.2 0.03 0.001 0.01 7 3 0.2 Ta <0.01 0.05 0.6 0.5 0.3 0.6 2

Br <5 - - - - 19 3 Tb 0.2 0.1 2 2 0.4 0.2 -

Cd 0.3 0.1 0.01 0.01 46 7 0.2 Te 0.1 <0.05 0.1 0.1 1 0.07 0.00036-0.01

Ce 12 7 155 139 26 21 45 Th 1 0.8 65 66 4 3 10

Co 2 0.9 0.7 0.9 6 10 25 Tl 26 0.1 0.002 0.004 8 4 1

Cr 51 7 43 74 41 176 200 Tm 0.1 0.1 0.8 0.8 0.2 0.08 -

Cs 0.2 0.1 0.1 1 46 2 2 U 1 1 13 14 2 1 3

Cu 19 2.008 7 12 54 80 60 V 15 0.1 0.3 0.2 1 46 150

Dy 1 0.7 9 10 2 1 4.5 W 2 31 654 560 71 5 1

Er 0.9 0.4 6 6 1 0.6 - Y 8 5 56 59 14 7 30

Eu 0.7 0.2 2 2 0.5 0.2 1.2 Yb 0.9 0.4 5 6 1 0.5 3

Ga 5 1 0.6 0.3 5 11 17 Zn 68 25 16 13 568 1257 70

Gd 1 0.8 10 10 3 1 7 Zr 50 9 280 294 65 38 160

Ge 0.4 0.1 0.2 0.2 0.6 1 15 >0 No increased concentration compared to waste rock

Hf 2 0.2 7 7 2 1 - >1 Increased concentration compared to waste rock

Hg 17 0.1 0.1 0.1 0.1 0.2 0.08 >20 Enriched when 5wt.% is added to waste rock

Ho 0.3 0.1 2 2 0.4 0.2 - >25 Enriched when 4wt.% is added to waste rock

La 6 4 85 78 14 11 6.5-100 >100 Enriched when 1wt.% is added to waste rock

24

Trace element occurrence in pyrite

As pyrite is the most abundant mineral in the waste rock, knowledge about trace element com-position and deportment within the pyrite grains is essential. Impurities and mineral inclusions can lead to physical stress in the crystal structure, making the mineral more susceptible to oxi-dation (Hutchinson and Ellison, 1992), profoundly affecting the leaching behavior and its qual-ity. A combination of spot analyses and mapping of the pyrite using LA-ICP-MS was performed to display trace element allocation. Chart counting data of spot analyses were used to track changes in trace element content during the analysis as deeper portions of the grains are ana-lyzed by the laser. Observations using charting count data show that the pyrite has several types of zonations containing trace elements.

Figure 6. Quantified LA-ICP-MS element distribution maps of an area of pyrite grains in the initial sulfidic waste rock (A) showing the spatial distribution of As, Cu, Pb, Sb, Tl, and Zn. Non-pyrite minerals are shaded.

25

Studies on trace element occurrence in the pyrite revealed that trace elements such as As, Cu, Hg, Pb, Sb, Tl, and Zn were associated with the pyrite, either as impurities or inclusions (Figure 6). Arsenic is seemingly the most evenly distributed trace element in the pyrite, varying between 67-585 ppm (average 272 ppm). However, mapping the pyrite infers concentrations of around1000 ppm, where the highest concentrations were observed in the more crystalline parts of thepyrite. Arsenic mainly occurs as an impurity, suggesting the occurrence of arsenic-bearing py-rite. Calculations regarding the mineral association of As suggest that the majority is associatedwith pyrite.

Compared to As, Cu, Pb, and Zn are more sporadically distributed in the pyrite, and spot anal-yses show spatial variations of 4-45 ppm (average 16 ppm), 0-50 ppm (average 14 ppm), and 1-882 ppm (average 105 ppm) respectively. Impurities of Cu, Pb, and Zn within the pyritematch low concentrations, whereas high concentrations are likely due to their presence as in-clusions associated with the pyrite's more porous parts. Although a substantial portion of theseelements are associated with pyrite, approximately 32% Cu can be explained by its associationwith chalcopyrite and bournonite. The corresponding amount for Zn was approximately 3%associated with sphalerite (Table 2 and 3).

Similarly, Hg, Sb, and Tl are also found in the more porous parts. However, predominantly located in and around fractures. Whereas Cu and Zn were mainly found as inclusions in the pyrite, Hg, Sb, and Tl were mainly found as impurities within the pyrite. Concentrations of Hg, Sb, and Tl varied at 0-449 (average 61 ppm), 2-102 (average 50 ppm), and 0-106 (average 14 ppm), respectively. Elevated concentrations of As, Hg, Sb, and Tl were observed compared to the average Earth's crust (Krauskopf and Bird, 1995) (Table 3). Elevated concentrations agree with trace element impurities in the pyrite showing the importance of detailed mineralogical studies.

Geochemical characterization of secondary raw materials

The different secondary raw materials used herein differ noticeably from each other regarding mineralogy (Table 2). The two slags BFS and GBFS, had similar chemical compositions (Table 3) as they originate from the same liquid slag. However, differences in cooling techniques resultin different degrees of crystallization. The air-cooled BFS mainly consists of monticellite(Ca(Mg, Fe)SiO4) and akermanite (Ca2Mg(Si2O7)), and the water granulated GBFS constitutesa glass phase (amorphous) and hercynite (Fe(II)Al2O4). Batch testing shows that both slags cangenerate alkaline leachates. The two slags BFS and GBFS, generated pH of 11.5 and 9.9. Ex-tending the batch test revealed that the pH of the GBFS (11.7) increased with time and wateramount, whereas no increase was seen for the BFS (Figure 7). This is likely an effect of GBFSdissolution at alkaline pH (Hooton, 2000), whereas BFS is more likely to dissolve with decreas-ing pH (Engström et al., 2013). This difference implies that the chemical conditions can have amajor impact on the dissolution of BFS and GBFS. A limited dissolution of easily water-solubleminerals was seen from both slags during batch testing (Table 4). The results imply that the twoslags, despite their limited dissolution, can effectively prevent ARD formation but could benefitfrom mixing with more reactive materials. Although silicates can contribute to the overall buff-ering capacity of a material, other minerals such as lime (CaO) and portlandite (Ca(OH)2) are

26

more reactive but soon exhausted, resulting in an increased buffering capacity in the short-term perspective. Conversely, silicates are perhaps too slow to give that initial increase in pH re-quired. For long-term neutralization of reactive waste rock, a mixture of reactive, intermediate, and more slowly buffering minerals would be preferred.

Comparable to the slags, both BA and CKD contained silicates but similar to the LKD, they also contained calcite (CaCO3), lime (CaO), and portlandite (Ca(OH)2). Both BA and CKD contained various salts such as anhydrite (CaSO4), arcanite (K2SO4), and sylvine (Na,KCl), whereas the LKD only contained sylvine (Na, KCl) (Table 2). The slightly higher pH observed for BA and CKD (12.9, respectively) compared to LKD (12) is likely related to variations in lime and portlandite content, whereas extensive salt dissolution can explain the substantially higher EC (27.9 and 31.2 mS/cm, respectively) (Duchesne and Reardon, 1998) (Table 4). The presence of a wide range of neutralizing minerals, both short and long-term dissolution, sug-gests that both BA and CKD can be more effective in preventing ARD formation than the slags. However, if the lime content is high, its hydration can result in the hardening of the material (Bulusu et al., 2007).

Moreover, a comparison to recommended levels for materials use above sealing layer in landfill construction (SEPA, 2010) showed that the use of SRM such as CKD and BA should be mini-mized. The soluble fraction of CKD and BA during batch testing showed that they both ex-ceeded concentrations for Cr and Pb (Table 4). Moreover, the CKD also exceeded recom-mended values for Cl and S and the BA for Zn. The enrichment and leachability of Pb in the

Figure 7. Correlation between the electrical conductivity (EC) and pH of granulated blast furnace slag (GBFS), granulated blast furnace slag (BFS), lime kiln dust (LKD), cement kiln dust (CKD) and biomass fly ash (BA) during an extended batch test (L/S 10, 12, 14). Results from kinetic leaching of sulfide-rich waste rock during two years is used as for reference (WR). Arrows indicate variations with time.

0 5 10 15

0

5

10

15

20

25

30

35

40

45

0

5

10

15

20

25

30

35

40

45

0 5 10 15

Axis

Title

EC [m

S/cm

]

pH

27

CKD and Cr and Zn in the BA (Table 3) suggest the presence of elements of potential concern that could limit the use of the materials.

Table 4. Concentrations of selected elements in leachate from batch testing (L/S 10) the secondary raw materials (SIS, 2003). Element concentrations are classified according to whether they exceed recommended levels for the use of waste materials in landfill construction above a sealing layer (SEPA, 2010). The classification for the soluble fraction includes elements As, Cl, Cr, Cu, Hg, Ni, Pb, S and Zn.

LKD BFS GBFS CKD BA Concn. (mg/L)

Concn. (mg/L)

Concn. (mg/L)

Concn. (mg/L)

Concn. (mg/L)

Al 4 4 0.1 0.04 0.04 Ca 338 110 9 1562 299 Cl 0.5 0.1 0.1 8737 728 Fe <0.002 <0.002 <0.002 0.004 0.01 K 5 16 2 5738 3354 Mg 0.02 0.1 3 0.1 <.010 Mn 0.0005 0.002 0.001 0.0002 0.001 Na 0.3 4 0.002 377 465 P 0.01 0.01 0.004 0.1 <0.005 S 2 271 5 892 26 Si 1 3 7 1 0.4 Ti <0.02 <0.002 <0.002 0.002 <0.002

µg/L µg/L µg/L µg/L µg/L As 0.08 <0.05 0.1 0.1 0.07 Ba 34 355 2 585 2250 Cd <0.025 <0.025 <0.025 0.5 0.3 Co <0.05 <0.05 <0.05 0.1 <0.05 Cr 6 1 <0.5 145 477 Cu <0.5 <0.5 <0.2 2 1.5 F <200 229 233 2600 <200 Hg <0.05 <0.05 <0.05 0.8 0.1 Ni <0.5 <0.5 <0.5 <0.5 <0.5 Pb 0.1 <0.05 <0.05 986 102 Sb 0.08 0.05 <0.05 0.5 0.1 Sr 438 275 24 7535 5904 U 0.01 <0.01 0.01 0.1 0.01 V 2 156 25 1 <0.5 Y <0.02 <0.02 <0.02 <0.05 0.03 Zn <2 <2 <2 10 897 Zr <0.05 <0.05 <0.05 <0.05 0.1

Exceed recommended values for landfill above a sealing layer (SEPA, 2010)

28

Sulfide oxidation and subsequent acid rock drainage formation

The leachate from the waste rock was initially characterized by a pH around 3 and an EC of approximately 25 mS/cm simultaneously as concentrations of Fe, Mn, and S spiked but slowly decreased alongside the EC (Figure 8). In contrast to examinations of the waste rock using optical microscopy, sequential extraction on the waste rock revealed that the waste rock was partially weathered with approximately 9% Fe and 7% S associated with what is herein assumed to be secondary phases. The majority of the secondary phases in the initial waste rock associated with steps I and II in sequential extraction, suggesting the presence of, e.g., hydrous soluble salt and adsorbed and exchangeable ions and limited occurrence of HFO, Al hydroxides, and Al hydroxysulfates.

During the initiation of kinetic testing, these readily soluble minerals were released into the leachate resulting in the spiking EC. In tandem with declining EC, the pH rose to levels of 4.6.with comparably low concentrations of Fe (4.4 mg/L), whereas S remained stable around 500-600 mg/L enables the precipitation of HFO such as schwertmannite, 2-, and 6-line ferrihy-drite but also gypsum which was supported by geochemical modeling (Figure 9).

Figure 8. Correlation between the electrical conductivity (EC) and pH from kinetic leaching of solely waste rock (A) and waste rock covered with 5 wt.% lime kiln dust (B), 4wt.% blast furnace slag and 1 wt.% cement kiln dust (C), 4 wt.% granulated blast furnace slag and 1 wt.% cement kiln dust (D), 5 wt.% granulated blast furnace slag (E), 1+1.5 wt.% bio-mass fly ash (F), 2.5 wt.% biomass fly ash (G), and 5 wt.% biomass fly ash (H) during up to two years of leaching. Arrows indicate changes with time

0 2 4 6 8 10 12 14

0

5

10

15

20

25

30

35

40

45

0

5

10

15

20

25

30

35

40

45

0 2 4 6 8 10 12 14

EC [m

S/cm

]

pH

A B C D E F G H

29

Despite the high pyrite content and low abundance of buffering minerals, it took approxi-mately 29 weeks of leaching before the waste rock any signs of accelerated sulfide oxidation and declining pH (<3.5). Rapid increasing concentrations peaked around week 62-71 with concentrations of Fe (18 g/L) and S (17 g/L) and As (21 mg/L), Hg (13 µg/L), Sb (967 µg/L), and Tl (317µg/L). Elements such as Cu (20 mg/L) and Zn (23 mg/L) peaked earlier, around week 31, whereas Pb (856 µg/L) displayed the highest concentrations much later, after 86 weeks. During the first 52 weeks of leaching, Fe and S were released, corresponding to ap-proximately 3% of their initial content.

At 103 weeks of leaching, the mobilization of Fe and S had increased by a magnitude of four to around 11%, and elements such as As (16%), Pb (7%), and Sb (4%) followed similar release patterns as Fe and S suggesting their occurrence in pyrite. In general, chalcopyrite is more weathering resistant than pyrite and arsenopyrite at both neutral and low pH and when Fe(III) is the primary oxidant (Plumlee, 1999; Rimstidt et al. 1994). However, when sulfides are in contact with each other, such as inclusions embedded in pyrite, electrons move between the sulfides, creating a galvanic cell where the mineral with higher rest potential acts as a cathode and the one with lower as an anode. As a result, the mineral with lower rest potential is prefer-entially weathered (Abraitis et al., 2004; Kwong et al., 2003; Nordstrom and Alpers, 1999; Evangelou, 1995; Evangelou and Zhang, 1995). In systems combining pyrite with chalcopyrite, the latter is more prone to oxidize (Chopard et al., 2017; Mehta and Murr, 1982; Majima, 1969), resulting in the extensive mobilization of Cu (79%). Similar reasoning could explain the exten-sive weathering of Zn (70%). In other words, it is likely that the combination of pyrite and inclusions within results in the high mobilization of Cu and Zn. Also, Cu's and Zn's preferential occurrence in and around fractures and the pyrite's more porous parts is likely to further con-tribute to the mobilization.

Figure 9. Geochemical calculations using PHREEQC version 3.4.0 (Parkhurst and Appelo, 1999) of saturation index of gyp-sum, jarosite, goethite (wateq4f.dat thermodynamic database (Ball and Nordstrom, 1991)), schwertmannite (Majzlan et al., 2004) originating from (Bigham et al., 1996) and 6-line ferrihydrite (Stefánsson, 2007) in the leaching of the sulfidic waste rock.

30

Provided the overall release of S and limited formation of secondary minerals, an extrapolation of the average oxidation rate suggests depletion of the pyrite within approximately 18 years. However, the high mobilization of metals such as Cu and Zn after only two years of leaching suggests that they, despite the low abundance, could be recovered by accelerating the sulfide oxidation.

Preventing acid rock drainage formation

Mixtures of (granulated) blast furnace slag and cement kiln dust

CKD (1 wt.%) was mixed with BFS (4 wt.%) or GBFS (4 wt.%) and added to the waste rock (5-30 mm) under the assumption that the quicklime content in the CKD would enable hydration of the slag and enable stabilization/solidification of the waste rock similar to as described by Tariq and Yanful (2013). However, minerals with cementing properties present in the CKD led to a smaller proportion of CKD compared to (G)BFS than reported by Chaunsali and Peetham-paran (2013).

Both mixtures barely managed to increase pH to near neutral (6.5 and 6.1, respectively), and neither could sustain it for a more extended time (Figure 8). The mixtures showed signs of hardening in addition to the waste rock, likely resulting from the CKD since the cementation rate of slag is generally slower (Tariq and Yanful, 2013). Moreover, Merox (2015) reported that cementation of BFS only occurs when the amount of fine particles is high, which was not the case, although the BFS had been crushed to a size of 0-4 mm before the study. Furthermore, the GBFS was not ground, which limits mineral hydration. The inability to maintain near-neu-tral pH is assumed to result from the hardening of the mixture. However, it should be noted that adding either of the mixtures to the waste rock had a positive effect on metal(loid) mobilization such as As and Zn. The decrease suggests that the use of CKD can have a positive effect on leachate chemistry. However, as the pH was not upheld, this may only occur in the short-term perspective. The cells were terminated 14 weeks post additions when the leachate pH had dropped below 4, respectively.

Granulated blast furnace slag

The addition of GBFS (5 wt.%) to the waste rock (30-60 mm) showed a similar trend as the GBFS and CKD mixture with a leachate pH barely increasing to near-neutral (6.2) and inability to maintain it for an extended period (Figure 8). A secondary mineral precipitate similar to carbonate was observed at the GBFS's surface a few weeks of leaching. No analysis was per-formed for confirmation. However, Huijgen and Comans (2005) reported that ground GBFS under ambient conditions could sequestrate CO2 through CaCO3 formation at a rate determined by Ca release. Altogether, the addition of GBFS to the waste rock decreased metal(loid) mobi-lization. However, the inability of GBFS dissolution resulting from low water content and sec-ondary mineral formation likely resulted in the inability to maintain near-neutral pH. The test cell was terminated ten weeks post the addition when the leachate had dropped below 4.

31

Lime kiln dust

The addition of LKD to the waste rock increased the pH to slightly above near-neutral without substantially increasing the EC (Figure 8). The pH was maintained for more than three years (only two years are presented for consistency) of leaching, and by creating a near-neutral pH environment, metals were immobilized through precipitation as or associated with secondary minerals. As seen in Figure 10, the addition of LKD to the waste rock was associated with significantly lower mobilization of elements by several orders of magnitude. For comparison, Cu mobilization when treating the waste rock with LKD was <0.01%, whereas it without treat-ment was 80%, illustrating the substantial improvement in water quality seen through the addi-tion of LKD to the waste rock.

The first four steps in sequential extraction were assumed to mainly consist of secondary min-erals and showed that the addition of LKD to the waste rock changed the behavior of the ele-ments studied. The waste rock in cell A was dominated by water-soluble phases such as gypsum and melanterite, whereas the geochemical environment in cell B promoted precipitation of and co-precipitation with secondary minerals such as HFO (Figure 11).

Gypsum is one of the most common and abundant sulfate minerals in mining environments (Jambor et al., 2013; Lottermoser, 2010), resulting from the oxidation and weathering products of pyrite and calcite (Plumlee et al., 2009). During extensive sulfide oxidation, Ca is often the rate-limiting element in gypsum formation since sulfide oxidation is faster than calcite dissolu-tion (Salmon and Malmström, 2002) which was seen in cell A. However, due to gypsum's high solubility, it is challenging to lower SO4 concentrations in the leachate since the mechanism that removes SO4 from the solution also causes the elevated concentrations (Lottermoser, 2010).

Geochemical calculation on gypsum saturation was used to indicate the sulfide oxidation rate in the waste rock. In cell B, the molar ratio of Ca/S was initially dominated by S. However, after 44 weeks of leaching, the ratio had increased to one simultaneously as gypsum saturation started to decrease. At the end of leaching, gypsum was suggested to be undersaturated, whereas the molar ratio of Ca/S remained around one (Figure 12), suggesting that S concentrations could originate from gypsum dissolution rather than sulfide oxidation. In general, the results suggest a decreasing sulfide oxidation rate due to the formation of potentially long-term stable second-ary minerals such as goethite and ferrihydrite (Figure 13).

The addition of 5 wt.% LKD to the waste rock corresponds to less than 4% of the total neutral-ization potential required to reach a positive net neutralization potential in the waste rock, which substantially limits the amount of material required to neutralize the ARD. The results suggest that the addition of LKD successfully limits sulfide oxidation and subsequent release of metals and metalloids into the leachate.

32

Figure 10. Estimated accumulated content of selected elements released into the leachate during the leaching of sulfidic waste rock (A), sulfidic waste rock and 5 wt.% lime kiln dust (B), sulfidic waste rock and 1+1.5 wt.% biomass fly ash (F), sulfidic waste rock and 2.5 wt.% biomass fly ash (G). The element release is calculated based on element concentrations multiplied with leachate volume withdrawn. Filled markings are actual measurements in the leachate. In contrast, out-lined markings are estimated as the average between the two most adjacent actual measurements.

33

Biomass fly ash

The addition of BA to waste rock in cells F (1 wt.%), G (2.5 wt.%), and H (5 wt.%) resulted in alkaline leachate with pH levels around 12 and elevated EC of 6-8 mS/cm (Figure 8) resulting from the dissolution of minerals such as lime and portlandite. The addition of BA resulted in spiking concentrations of elements such as As, Cl, Cr, Mo, S, and Sb, likely due to their affinity of being condensed on the BA's particle surface (Kukier et al., 2003; Jones, 1995), resulting in their mobilization. However, the increased concentrations were temporary and soon declined to stable levels (Figure 10).

Figure 11. Extracted element content of the total element content in the waste rock (5-30 mm) before leaching (A0) and from cell A and B (waste rock and 5 wt.% lime kiln dust) on separate occations after 52 and 103 weeks or leaching. Comparison with larger waste rock size (30-60 mm) before leaching (Initial) and from cell F (waste rock and 1 wt.% bio-mass fly ash) and G (waste rock and 2.5 wt.% biomass fly ash).Step I-IV in sequential extraction after Dold (2003). Water-soluble fraction (I), exchangeable fraction (II), Fe(III)oxyhydroxides (III), and Fe(III)oxides (IV). Concentrations below de-tection limit are not presented.

34

Quickly after adding BA to cells G (2.5 wt.%) and H (5 wt.%), the BA started showing signs of hardening resulting from the BA's ability to spontaneously react with moisture, water, and CO2 in the air to form hydroxides, carbonates, and sulfate minerals such as gypsum and ettring-ite. The hardening results in decreased solubility of the BA with limited ability to buffer the pH (Karltun et al., 2008), which can be seen as decreased leaching rate of Ca and lower alkalinity and pH (Etíegni and Campbell, 1990). In the case of cell H (5 wt.%), the extensive hardening led to the inability to maintain a near-neutral pH, and the test cell was terminated ten weeks post addition of the BA. However, the hardening is not synonymous with depleted neutraliza-tion potential; it is rather an effect of the inability of buffering minerals to dissolve. The BA in test cell G (2.5 wt.%) showed similar but less extensive hardening and a brittleness test of the BA after 14 weeks of leaching led to spiking concentrations of As, Ba, Ca, Cl, Cr, K, Mg, Mo, Na, P, S, Sb, Sr and V correlating with elements spiking when the BA was first added suggest-ing that the hardening of the BA limited the mobilization of these elements.

After 61 weeks of leaching the waste rock in cell F (1 wt.%), the pH had dropped below 4.6 due to sulfide oxidation. An increase of elements such as Al, Cu, Mn, and Pb in secondary phases was seen after leaching the waste rock for 69 weeks. Owing to the abundance of these elements in the BA and the amount of BA added to the waste rock, their increase in secondary phases is more likely to originate from sulfide oxidation, where the favorable pH allowed for their immobilization in secondary phases. The only element whose association with secondary phases can be explained by the BA addition is Zn in cell G.

After 69 weeks of leaching the waste rock with BA in cell F (1 wt.%), an additional 1.5 wt.% was added, resulting in a similar flush of elements as previously described. In general, a higher mobilization was seen for cell G (2.5 wt.%) (Figure 10). However, these are approximations, and the disturbance of the BA in cell 2 likely contributed to the increasing element mobilization. Overall, this concludes that it does not matter substantially whether one or multiple BA additions are made concerning easily mobilized elements from the BA.

Figure 12. Molar ratio of Ca/S and saturation index of gypsum during the leaching of sulfidic waste rock covered with 5 wt.% lime kiln dust. Solid line illustrate the 1:1 relationship between Ca and SO4. Dashed lines mark values close to zero (±0.5) which indicate equilibrium between the solution and mineral phase.

35

There is no linear correlation between the amount of BA added and the time with which the BA can uphold a near-neutral pH. Hardening seems to have a substantial impact partially due to decreased dissolution rate seen through extensive hardening in cell H and the comparable higher dissolution rate seen in cell F through less extensive hardening.

Despite the lack of reference cell (30-60 mm) for direct comparison, it can be concluded that the BA has an overall positive effect on the mobilization of elements from the waste rock. As seen in Figure 10, the addition of BA resulted in a mobilization of elements corresponding to <0.5%. Assuming the BA has approximately 1/3 of the buffering capacity of BA, an addition of 2.5 wt.% BA to the waste rock would correspond to <1% of the total neutralization potential required to reach a positive net neutralization potential in the waste rock, which substantially limits the amount of material required to neutralize the ARD. However, it is essential to add enough BA to maintain pH for a more extended period.

Figure 13. Stereomicroscope images (A-C) and secondary electron images (D-F) of waste rock samples before leaching (initial) and after 52 and 103 weeks of leaching the waste rock with 5wt.% lime kiln dust (LKD). G. Raman spectrum of one representative point in the initial waste rock showing the presence of pyrite and gypsum/anhydrite. H. Representa-tive Raman spectrum of the waste rock with 5 wt.% lime kiln dust after 52 and 103 weeks of leaching showing the pres-ence of goethite and ferrihydrite. Reference spectras were taken from Monnier et al. (2011).

36

Secondary minerals formation

One apprehension with adding neutralizing minerals is that an excess of secondary minerals could form. If secondary minerals are built up, elements are stored and may be released in changing geochemical conditions by applying a dry cover or when backfilled into an open pit followed by flooding, causing drainage from pit lake or heap in need of treatment for an ex-tended time before released into the recipient.

While secondary minerals formation may look extensive through intense, sometimes vibrant, coloring, they often consist of very thin rims, making their identification difficult, not to men-tion their quantification. Although sequential extraction cannot identify the individual mineral, it can be a valuable tool to estimate the extent of secondary mineral formation and how elements associate with mineral characteristics.

Figure 11 shows both similarities and differences in leaching the waste rock with and without additions of SRM. Although no direct comparison for the extent of secondary minerals for-mation can be made between the different waste rock fractions due to no information about the onset of sulfide oxidation, some comparisons can still be made.

For instance, it could be seen that whereas the leaching of the reference waste rock (cell A) accumulated secondary minerals associated with a wide range of steps (I-IV), the addition of both cells with BA (G and F) and LKD (B) accumulated elements mainly associated to steps III-IV suggesting an accumulation of HFO with the association of elements such as As, Cr, Cu,Pb, and Zn. It is also likely that Cr, Cu, Pb, and Zn were adsorbed to these HFO minerals incells G and F, whereas Cr, Pb, and Zn were more likely in cell B.

It was also noticed that elements such as Cu, Pb, and Zn were higher in cell G than in F, likely resulting from the increased mobilization of these elements before the second addition of BA to the waste rock in this cell. As seen in Figure 11, a generally lower amount of secondary minerals was formed by adding LKD to the waste rock (cell B) compared to no addition (cell A). Whereas secondary minerals in cell B were generally associated with steps III and IV, sug-gesting their association to or co-precipitation with HFO, secondary phases in cell A were pri-marily associated with steps I and II in the presence of water-soluble minerals or adsorbed/ex-changeable ions.

38

DISCUSSION

Challenges with using secondary raw materials for preventive purposes

When attempting to prevent the formation of ARD from sulfidic waste rock using SRM for HFO precipitation on reactive surfaces, some aspects are vital to the success:

1. The sulfide oxidation rate and subsequent release of Fe.2. The dissolution rate of the SRM and its overall buffering capacity to maintain near-

neutral pH in a long-term perspective.3. Matching 1 and 2 physically on a time-scale perspective to enable HFO precipitation.

In other words, the alkaline material must dissolve close to and simultaneously as sulfides are oxidizing. The supply of alkalinity must therefore be as fast as or faster than the generation of acids by sulfide oxidation. The weathering mechanisms for sulfides and SRM are different (ox-ygen compared to water and/or acid), meaning that if the SRM is added before the onset of sulfide oxidation, its neutralization capacity may be exhausted. It is necessary to determine when the SRM must be added to the waste rock after it has been deposited to treat the waste rock in the most efficient and economic manner.

Considering the logistics of SRM being generated in Sweden, problems can arise from adopting this approach on a larger scale. As greater volumes of SRM are needed, they may be required to accumulate over time or be combined from competing producers. One of the fundamentals with secondary raw materials is that it is secondarily produced, resulting in small to large vari-ations in the material's characteristics. These variations can occur with time for the same pro-ducer as a result of differences in feeding material, storing of the material, or changes in the production process. Variations in SRM characteristics can also occur between producers due to differing manufacturing practices (Alakangas et al., 2014).

A problem related to the requirement of larger volumes of SRM is that if the material needs to be stored on-site, it must be stored suitably, or else reactive SRM such as BA and LKD risk being altered. Lime kiln dust typically contains a substantial amount of lime (Bulusu et al., 2007; Miller and Callaghan, 2004). However, the LKD used within this thesis showed approx-imately 1/10th of the lime content reported by the producer. Contrary to these findings, calcite was highly abundant in the material, suggesting an alteration of the material, by the hydration and re-carbonation of lime, during the storage of LKD in piles under ambient conditions. This problem of alternation may have actually had a positive effect as it has caused the material to accumulate into porous aggregates. This aggregation suggests that the carbonate-rich LKD can be superior to coarser crushed calcite through its increased specific surface area, which would increase its ability to dissolve.

Biomass fly ash is an attractive SRM as it is highly accessible and generated in numerous loca-tions around Sweden. The depletion of reactive minerals such as lime and calcite can result in a substantial loss of short-term neutralizing capacity. Provided that not all reactive minerals are

39

depleted during storing of the BA, then the storing can positively affect the hardening and wash out of easily soluble salts and other readily soluble minerals to create a "cleaner" material in-creasing its suitability for preventive purposes. However, other SRM may behave differently.

The importance of knowing the material

Characterization for predicting a materials' behavior and, more importantly, how these materi-als behave together is essential for the outcome of preventing ARD formation and requires research and a strategic approach to maximize the output. The choice of method for character-ization of an SRM largely influences the type of information which is generated.

National guidelines regarding material characterization for landfilling include two-stage batch testing and upward percolation testing (NFS 2010:4). However, this approach cannot give a comprehensive view of the material's harmfulness. Batch-testing of the BA studied in this thesis did not show elevated P concentration which was later shown when adding the SRM to the waste rock. Instead, it was knowledge about the production process and detailed mineralogical studies, such as those performed by Kukier et al. (2003), that revealed the preferential conden-sation of P on the BA's particle surface.

The waste rock similarly needs to be studied since the truth behind a behavior often lies within the details. Small-scale mineral features are generally unknown in practice since their determi-nation requires more advanced and sometimes costly techniques (Parbhakar-Fox and Lotter-moser, 2015). However, these factors could explain the error seen between predictive testing and field results (Maest and Nordstrom, 2017). A solution could be the use of automated min-eralogy, which allows for relatively rapid analysis of thousands of particles by grain size, mineralogy, morphology, and degree of liberation from other minerals (Brough et al., 2013). Although, the question remains whether or not it is detailed enough.

By doing research and being strategic in the choices for material characterization, the output of information to the amount of effort spent can be maximized. However, despite doing detailed mineralogical studies and suitable geochemical investigations, standard analytical methods can largely diminish the output. An example being Tl which is considered more toxic to humans than both Cd and Hg (Peter and Viraraghavan, 2005). However, Tl is rarely studied and often undetected or deselected when choosing which parameters to analyze despite its association to sulfide minerals like pyrite (Abraitis et al. 2004). By adopting the strategy of screening, elements like Tl do not go undetected. Thus, detailed geochemical and mineralogical characteri-zation is clearly essential for improving the prediction of and long-term performance of pre-vention measures at mine sites.

Lessons learned and opportunities for improving ARD prevention approach

The main objective of this thesis was to reduce sulfide oxidation and the subsequent formation of ARD in sulfidic waste rock. Although the long-term stability of this techniques could be debated, it has some advantages. A significant finding was that a highly limited amount of SRM could effectively decrease the sulfide oxidation rate. The overall benefits are that a smaller

40

amount of SRM was used for preventive purposes than required to treat ARD generated without preventive measures.

Encapsulation by HFO can take decades to be considered a "final solution" (Park et al., 2019). However, this type of technique has some clear advantages without being complete by pushing back the time of onset accelerated sulfide oxidation and can substantially limit the overall en-vironmental impact. Even though it has been shown that the physical contact between the acid-producing and neutralizing minerals is crucial for secondary mineral formation for passivation of the sulfide surface, for example, by mixing the materials (Pérez-López et al. 2007) applying the neutralizing material on top, or in layers in between, the acid-producing material can be advantageous due to costs and operational issues.

However, one needs to adopt a holistic approach before commencing this type of preventive strategy, such as starting at the exploration stage, building on the information from characteri-zation and prediction with mine planning design and waste management strategies, just to men-tion a few. Concerning national legislation, would a successfully passivated mine waste be al-lowed, or would there still be requirements for additional remediation methods such as back-filling followed by flooding at mine closure, which jeopardizes the effect of the preventive method? Such questions must be answered before committing to preventive methods relying on HFO formation.

The peer-reviewed literature related to ARD prevention techniques, especially passivation ap-proaches, has shown promising results. However, the studies are still primarily dominated by laboratory studies focusing on pure pyrite samples. Moreover, they often rely on the complete submersion of the waste in the solution in question for successful passivation, which, due to current methods of mine waste handling, make these methods questionable. Whether they could be successfully applied on large-scale tailings with more complex miner-alogy under complex environmental conditions is still unknown. Moreover, for their application to waste rock, an alternative approach to the current traditional deposition strategy of placing the material in tall stockpiles may need to be adopted. Regardless, there is an impending need for preventative methods, stressing the need for these types of studies.

42

CONCLUSIONS

Geochemical characterization of the waste rock identified As, Cu, Hg, Pb, Sb, and Tl as trace elements associated with the pyrite, either as impurities or inclusions, and that their association with the pyrite resulted in their weathering where Cu (79%) and Zn (72%) was most extensive. These results suggest that these elements can be recovered from the leachate rather than stored in the sludge.

When comparing to a single addition of BA, the addition of BA in sequences does not seem to result in a higher mobilization of elements. The apparent advantage of the sequential addition is that it can be balanced with the sulfide oxidation rate without risking using an excess of material. However, since there is no linear correlation between the amount of BA added and the time with which it can maintain near-neutral pH, one must ensure that enough material is added in each sequence.

The timely addition of small amounts of SRM such as BA or LKD, corresponding to <1-4% of the waste rock's net neutralizing potential, can push back the onset of accelerated sulfide oxi-dation and significantly lessen the overall negative environmental impact that no addition of SRM would result in.

The fear of an excess amount of minerals forming will dissolve once the material has been covered and the environment becomes reduced is still relevant. However, the amount of sec-ondary minerals accumulated when treating the waste rock with BA or LKD is low compared to the amount formed without these measures. The type of minerals formed without the addition of SRM are likely to dissolve regardless of a changing geochemical environment since they are dissolved by water or adsorbed to other minerals.

44

REFERENCES

A

Abraitis, P. K., Pattrick, R. A. D., & Vaughan, D. J. (2004). Variations in the compositional, textural and electrical properties of natural pyrite: A review. International Journal of Mineral Processing, 74(1-4), 41-59. doi:10.1016/j.minpro.2003.09.002

Ačai, P., Sorrenti, E., Gorner, T., Polakovič, M., Kongolo, M., & de Donato, P. (2009). Pyrite pas-sivation by humic acid investigated by inverse liquid chromatography. Colloids and Surfaces A: Physicochemical and Engineering Aspects, 337(1-3), 39-46. doi:10.1016/j.colsurfa.2008.11.052

Acharya, B. S., & Kharel, G. (2020). Acid mine drainage from coal mining in the united states – an overview. Journal of Hydrology, 588 doi:10.1016/j.jhydrol.2020.125061

Alakangas, L., Andersson, E., & Mueller, S. (2013). Neutralization/prevention of acid rock drainage using mixtures of alkaline by-products and sulfidic mine wastes. Environmental Science and Pollu-tion Research, 20(11), 7907-7916. doi:10.1007/s11356-013-1838-z

Alakangas, L., Maurice, C., Macsik, J., Nyström, E., Sandström, N., Andersson-Wikström, A., & Häll-ström, L. (2014). Kartläggning av restprodukter för efterbehandling och inhibering av gruvavfall: Funktion tillgång och logistik. Luleå: Luleå tekniska universitet. Retrieved from RIS Format UTF-8 Retrieved from http://proxy.lib.ltu.se/login?url=http://search.ebscohost.com/login.aspx?di-rect=true&db=edsswe&AN=edsswe.oai.DiVA.org.ltu.24560&lang=sv&site=eds-live&scope=site

Amos, R. T., Blowes, D. W., Bailey, B. L., Sego, D. C., Smith, L., & Ritchie, A. I. M. (2015). Waste-rock hydrogeology and geochemistry. Applied Geochemistry, 57, 140-156. doi:10.1016/j.apgeo-chem.2014.06.020

ASTM. (2011). Standard test methods for chemical analysis of limestone, quicklime, and hydrated lime. ASTM Standard C25–11, ASTM International, West Conshohocken, Pennsylvania, United States

B

Baker-Austin, C., & Dopson, M. (2007). Life in acid: PH homeostasis in acidophiles. Trends in Micro-biology, 15(4), 165-171. doi:10.1016/j.tim.2007.02.005

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 wa-ters. ( No. 189). Open-file report: US Geological Survey.

Belzile, N., Maki, S., Chen, Y., & Goldsack, D. (1997). Inhibition of pyrite oxidation by surface treat-ment. Science of the Total Environment, 196(2), 177-186. doi:10.1016/S0048-9697(96)05410-1

Bigham, J. M., Schwertmann, U., Traina, S. J., Winland, R. L., & Wolf, M. (1996). Schwertmannite and the chemical modeling of iron in acid sulfate waters. Geochimica Et Cosmochimica Acta, 60(12), 2111-2121. doi:10.1016/0016-7037(96)00091-9

Brough, C. P., Warrender, R., Bowell, R. J., Barnes, A., & Parbhakar-Fox, A. (2013). The process min-eralogy of mine wastes. Minerals Engineering, 52, 125-135. doi:https://doi.org/10.1016/j.mineng.2013.05.003

Brown, M., Barley, B., & Wood, H. (2002). Minewater treatment: Technology, application and policy. London: IWA Publishing.

Bulusu, S., Aydilek, A. H., & Rustagi, N. (2007). CCB-based encapsulation of pyrite for remediation of acid mine drainage doi:https://doi.org/10.1016/j.jhazmat.2007.01.035

45

C

Cambridge University Press. (2021). Cambridge dictionary. Retrieved from https://dictionary.cam-bridge.org/

Chaunsali, P., & Peethamparan, S. (2013). Novel cementitious binder incorporating cement kiln dust: Strength and durability. ACI Materials Journal, 110(3), 297-304.

Chopard, A., Plante, B., Benzaazoua, M., Bouzahzah, H., & Marion, P. (2017). Geochemical investiga-tion of the galvanic effects during oxidation of pyrite and base-metals sulfides. Chemosphere, 166, 281-291. doi:10.1016/j.chemosphere.2016.09.129

D

De Vries, N. H. C. (1996). Process for treating iron-containing sulfide rocks and ores. US Patent: 5,587,001.

Demers, I., Mbonimpa, M., Benzaazoua, M., Bouda, M., Awoh, S., Lortie, S., & Gagnon, M. (2017). Use of acid mine drainage treatment sludge by combination with a natural soil as an oxygen barrier cover for mine waste reclamation: Laboratory column tests and intermediate scale field tests. Min-erals Engineering, 107, 43-52. doi:10.1016/j.mineng.2016.11.017

Diao, Z., Shi, T., Wang, S., Huang, X., Zhang, T., Tang, Y., & Qiu, R. (2013). Silane-based coatings on the pyrite for remediation of acid mine drainage. Water Research, 47(13), 4391-4402. doi:10.1016/j.watres.2013.05.006

Dold, B. (2003). Speciation of the most soluble phases in a sequential extraction procedure adapted for geochemical studies of copper sulfide mine waste. Journal of Geochemical Exploration, 80(1), 55-68. doi:10.1016/S0375-6742(03)00182-1

Dold, B. (2017). Acid rock drainage prediction: A critical review doi:10.1016/j.gexplo.2016.09.014

Doye, I., & Duchesne, J. (2003). Neutralisation of acid mine drainage with alkaline industrial residues: Laboratory investigation using batch-leaching tests. Applied Geochemistry, 18(8), 1197-1213. doi:10.1016/S0883-2927(02)00246-9

Doye, I., & Duchesne, J. (2005). Column leaching test to evaluate the use of alkaline industrial wastes to neutralize acid mine tailings. Journal of Environmental Engineering, 131(8), 1221-1229. doi:10.1061/(ASCE)0733-9372(2005)131:8(1221)

Duchesne, J., & Doye, I. (2005). Effectiveness of covers and liners made of red mud bauxite and/or cement kiln dust for limiting acid mine drainage. Journal of Environmental Engineering, 131(8), 1230-1235. doi:10.1061/(ASCE)0733-9372(2005)131:8(1230)

Duchesne, J., & Reardon, E. J. (1998). Determining controls on element concentrations in cement kiln dust leachate. Waste Management, 18(5), 339-350. doi:10.1016/S0956-053X(98)00078-6

E

Elberling, B., Schippers, A., & Sand, W. (2000). Bacterial and chemical oxidation of pyritic mine tail-ings at low temperatures. Journal of Contaminant Hydrology, 41(3-4), 225-238. doi:10.1016/S0169-7722(99)00085-6

Elsetinow, A. R., Borda, M. J., Schoonen, M. A. A., & Strongin, D. R. (2003). Suppression of pyrite oxidation in acidic aqueous environments using lipids having two hydrophobic tails. Advances in Environmental Research, 7(4), 969-974. doi:10.1016/S1093-0191(02)00101-6

Engström, F., Adolfsson, D., Samuelsson, C., Sandström, Å., & Björkman, B. (2013). A study of the solubility of pure slag minerals. Minerals Engineering, 41, 46-52. doi:10.1016/j.mineng.2012.10.004

46

Etiégni, L., & Campbell, A. G. (1991). Physical and chemical characteristics of wood ash doi:https://doi-org.proxy.lib.ltu.se/10.1016/0960-8524(91)90207-Z

Evangelou, V. P. (1995). Potential microencapsulation of pyrite by artificial inducement of ferric phos-phate coatings. Journal of Environmental Quality, 24(3), 535-542.

Evangelou, V. P. (1996). Oxidation proof silica surface coating iron sulfides. US Patent: 5,494,703.

Evangelou, V. P. (2001). Pyrite microencapsulation technologies: Principles and potential field appli-cation. Ecological Engineering, 17(2-3), 165-178. doi:10.1016/S0925-8574(00)00156-7

Evangelou, V.P., & Zhang, Y.L. (1995). A review: pyrite oxidation mechanisms and acid mine drainage prevention. Crit Rev Environ Sci Technol, 25, 141–199

F

Fan, R., Short, M. D., Zeng, S., Qian, G., Li, J., Schumann, R. C., & Gerson, A. R. (2017). The formation of silicate-stabilized passivating layers on pyrite for reduced acid rock drainage. Environmental Sci-ence and Technology, 51(19), 11317-11325. doi:10.1021/acs.est.7b03232

Fytas, K., & Evangelou, B. (1998). Phosphate coating on pyrite to prevent acid mine drainage. Int J Surf Min Reclam Environ, 12, 101–104

Földvári M (2011). Handbook of the thermogravimetric system of minerals and its use in geological practice. Occasional Papers of the Geological Institute of Hungary. 213. Budapest ISBN 978-963-671-288-4

G

Gong, B., Li, D., Niu, Z., Liu, Y., & Dang, Z. (2021). Inhibition of pyrite oxidation using PropS-SH/se-piolite composite coatings for the source control of acid mine drainage. Environmental Science and Pollution Research, 28, 11090-11105. doi:10.1007/s11356-020-11310-0

H

Hallberg, R. O., Granhagen, J. R., & Liljemark, A. (2005). A fly ash/biosludge dry cover for the miti-gation of AMD at the falun mine. Chemie Der Erde, 65(SUPPL. 1), 43-63. doi:10.1016/j.chemer.2005.06.008

Hodges, G., Roberts, D.W., Marshall, S.J., & Dearden, J.C. (2006). The aquatic toxicity of anionic surfactants to Dophnia magna—a comparative QSAR study of linear alkylbenzene sulphonates and ester sulphonates. Chemosphere 63, 1443–1450

Hollings, P., Hendry, M. J., Nicholson, R. V., & Kirkland, R. A. (2001). Quantification of oxygen con-sumption and sulphate release rates for waste rock piles using kinetic cells: Cluff lake uranium mine, northern saskatchewan, Canada. Applied Geochemistry, 16(9), 1215-1230. doi:10.1016/S0883-2927(01)00005-1

Holmström, H., Ljungberg, J., & Öhlander, B. (1999). Role of carbonates in mitigation of metal release from mining waste. Evidence from humidity cells tests. Environmental Geology, 37(4), 267-280. doi:10.1007/s002540050384

Hooton, R. D. (2000). Canadian use of ground granulated blast-furnace slag as a supplementary cement-ing material for enhanced performance of concrete. Canadian Journal of Civil Engineering, 27(4), 754-760.

Huang, X., & Evangelou, V. P. (1992). Abatement of Acid Mine Drainage by Encapsulation of Acid Producing Geologic Materials,

47

Hudson-Edwards, K. A., Jamieson, H. E., & Lottermoser, B. G. (2011). Mine wastes: Past, present, future. Elements, 7(6), 375-380. doi:10.2113/gselements.7.6.375

Huijgen, W. J. J., & Comans, R. N. J. (2005). Mineral CO2 sequestration by steel slag carbonation. En-vironmental Science and Technology, 39(24), 9676-9682. doi:10.1021/es050795f

Huminicki, D. M. C., & Rimstidt, J. D. (2009). Iron oxyhydroxide coating of pyrite for acid mine drain-age control. Applied Geochemistry, 24(9), 1626-1634.

Hutchison, I., & Ellison, R. (1992). Mine waste management; Lewis Publishers: Boca Raton.

Höglund, L-O., Herbert, R., Lövgren, L., Öhlander, B., Neretniks, I., Moreno, L., Malmström, M., Elan-der, P., Lindvall, M., & Lindström, B. (2003). Mimi-Performande Assessment, Main Report, Mimi-Report 2003:3, Moreno, Italy. ISSN 1403-9478, ISBN 91-89350-27-8

Hökfors, B. (2014). Phase chemistry in process models for cement clinker and lime production (Doc-toral thesis). Retrieved from http://urn.kb.se/resolve?urn=urn:nbn:se:umu:diva-86004

I

INAP (International Network for Acid Prevention) Global Acid Rock Drainage (GARD) Guide. http://www.gardguide.com/images/5/5f/TheGlobalAcidRockDrainageGuide.pdf (accessed 18 Janu-ary 2021).

J

Jambor, J.L., Blowes, D.W., & Ritchie, A.I.M. (2013). Environmental aspects of mine wastes.Miner-alogical Association of Canada, Nepean, Ontario

Jamieson, H. E., Walker, S. R., & Parsons, M. B. (2015). Mineralogical characterization of mine waste. Applied Geochemistry, 57, 85-105. doi:10.1016/j.apgeochem.2014.12.014

Jha, R. K. T., Satur, J., Hiroyoshi, N., Ito, M., & Tsunekawa, M. (2008). Carrier-microencapsulation using si-catechol complex for suppressing pyrite floatability. Minerals Engineering, 21(12-14), 889-893. doi:10.1016/j.mineng.2008.02.011

Jha, R.K.T., Satur, J., Hiroyoshi, N., Ito, M., & Tsunekawa, M., 2012. Suppression of pyrite oxidation by Carrier microencapsulation using Si and catechol. Min. Proc. Ext. Met. Rev. 33 (2), 89-98.

Jha, R. K. T., Satur, J., Hiroyoshi, N., Ito, M., & Tsunekawa, M. (2011). Suppression of floatability of pyrite in coal processing by carrier microencapsulation. Fuel Processing Technology, 92(5), 1032-1036. doi:10.1016/j.fuproc.2010.12.028

Ji, M., Gee, E., Yun, H., Lee, W., Park, Y., Khan, M. A., & Choi, J. (2012). Inhibition of sulfide mineral oxidation by surface coating agents: Batch and field studies. Journal of Hazardous Materials, 229-230, 298-306. doi:10.1016/j.jhazmat.2012.06.003

Jiang, C. L., Wang, X. H., & Parekh, B. K. (2000). Effect of sodium oleate on inhibiting pyrite oxida-tion. International Journal of Mineral Processing, 58(1-4), 305-318. doi:10.1016/S0301-7516(99)00045-9

Johnson, D. B., & Hallberg, K. B. (2005). Acid mine drainage remediation options: A review. Science of the Total Environment, 338(1-2 SPEC. ISS.), 3-14. doi:10.1016/j.scitotenv.2004.09.002

Jones, D. R. (1995). The leaching of major and trace elements from coal ash. Netherlands: Kluwer Ac-ademic Publishers. Retrieved from http://inis.iaea.org/search/search.aspx?orig_q=RN:27031404

Jones, S. N., & Cetin, B. (2017). Evaluation of waste materials for acid mine drainage remedia-tion doi:https://doi-org.proxy.lib.ltu.se/10.1016/j.fuel.2016.10.018

K

48

Kang, C., Jeon, B., Park, S., Kang, J., Kim, K., Kim, D., & Kim, S. (2016). Inhibition of pyrite oxidation by surface coating: A long-term field study. Environmental Geochemistry and Health, 38(5), 1137-1146. doi:10.1007/s10653-015-9778-9

Kargbo, D. M., Atallah, G., & Chatterjee, S. (2004). Inhibition of pyrite oxidation by a phospholipid in the presence of silicate. Environmental Science and Technology, 38(12), 3432-3441. doi:10.1021/es0352552

Karltun. E., Saarsalmi, A., Ingerslev. M., Mandre, M., Andersson, S., Gaitnieks, T., Ozolinčius, R., & Varnagiryte-Kabasinskiene, I. (2008). Wood Ash Recycling – Possibilities And Risks. In: Röser D., Asikainen A., Raulund-Rasmussen K., Stupak I. (eds) Sustainable Use of Forest Biomass for En-ergy. Managing Forest Ecosystems, Springer, Dordrecht, 12.,78-108. doi:10.1007/978-1-4020-5054-1_4

Kefeni, K. K., Msagati, T. A. M., & Mamba, B. B. (2017). Acid mine drainage: Prevention, treatment options, and resource recovery: A review. Journal of Cleaner Production, 151, 475-493. doi:10.1016/j.jclepro.2017.03.082

Khummalai, N., & Boonamnuayvitaya, V. (2005). Suppression of arsenopyrite surface oxidation by sol-gel coatings. Journal of Bioscience and Bioengineering, 99(3), 277-284. doi:10.1263/jbb.99.277

Kleinmann, R.L.P. (1999). Bactericidal control of acidic drainage. In: Brady, K.B.C., Smith, M.W., & Schuek, J., (eds) Coal mine drainage prediction and pollution prevention in Pennsylvania. The Penn-sylvania Department of Environmental Protection, Chapter 15, 15.1 to 15.6

Kollias, K., Mylona, E., Adam, K., Chrysochoou, M., Papassiopi, N., & Xenidis, A. (2019). Character-ization of phosphate coating formed on pyrite surface to prevent oxidation. Applied Geochemis-try, 110, 104435. doi:10.1016/j.apgeochem.2019.104435

Kollidas, K., Mylona, E., Papassiopi, N., &Xenidis, A. (2018). Development of silica protective layer on pyrite surface: a column study. Environ Sci Pollut Res, 25, 26780-26792. doi:10.1007/s11356-017-0083-2

Kollias, K., Mylona, E., Papassiopi, N., & Xenidis, A. (2015). Conditions favoring the formation of iron phosphate coatings on the pyrite surface. Desalination and Water Treatment, 56(5), 1274-1281. doi:10.1080/19443994.2014.958537

Krauskopf, K. B., & Bird, D. K. (1995). Introduction to geochemistry (3.ed.). New York: McGraw-Hill.

Kukier, U., Ishak, C. F., Sumner, M. E., & Miller, W.P. (2003). Composition and element solubility of magnetic and non-magnetic fly ash fractions. Environmental Pollution, 123(2), 255-266. doi:10.1016/S0269-7491(02)00376-7

Kwong, Y.T.J., Swerhone, G.W., & Lawrence, J.R. (2003). Galvanic sulphide oxidation as a metal-leaching mechanism and its environmental implications. Geochem. Explor. Environ. Anal. 3, 337–343.

49

L

Lalvani, S. B., DeNeve, B. A., & Weston, A. (1990). Passivation of pyrite due to surface treat-ment. Fuel, 69(12), 1567-1569. doi:10.1016/0016-2361(90)90209-9

Li, D., Gong, B., Liu, Y., & Dang, Z. (2021). Self-healing coatings based on PropS-SH and pH-respon-sive HNT-BTA nanoparticles for inhibition of pyrite oxidation to control acid mine drainage. Chem-ical Engineering Journal, 415, 128993. doi:10.1016/j.cej.2021.128993

Li, X., Hiroyoshi, N., Tabelin, C. B., Naruwa, K., Harada, C., & Ito, M. (2019). Suppressive effects of ferric-catecholate complexes on pyrite oxidation. Chemosphere, 214, 70-78. doi:10.1016/j.chemo-sphere.2018.09.086

Liu, Y., Hu, X., & Xu, Y. (2017). PropS-SH/SiO2 nanocomposite coatings for pyrite oxidation inhibi-tion to control acid mine drainage at the source. Journal of Hazardous Materials, 338, 313-322. doi:10.1016/j.jhazmat.2017.05.043

Liwarska, B.E., Miksch, K., Malachowska, J.A., & Kalka, J. (2005). Acute toxicity and genotoxicity of five selected anionic and nonionic surfactants. Chemosphere 58, 1249–1253

Lottermoser, B. G. (2010). Mine wastes (third edition): Characterization, treatment and environmental impacts. Mine wastes (third edition): Characterization, treatment and environmental impacts (pp. 1-400) doi:10.1007/978-3-642-12419-8

Lu, J., Alakangas, L., & Wanhainen, C. (2014). Metal mobilization under alkaline conditions in ash-covered tailings. Journal of Environmental Management, 139, 38-49.

M

Mackie, A. L., & Walsh, M. E. (2012). Bench-scale study of active mine water treatment using cement kiln dust (CKD) as a neutralization agent. Water Research, 46(2), 327-334. doi:10.1016/j.wa-tres.2011.10.030

Madzivire, G., Maleka, P. P., Vadapalli, V. R. K., Gitari, W. M., Lindsay, R., & Petrik, L. F. (2014). Fate of the naturally occurring radioactive materials during treatment of acid mine drainage with coal fly ash and aluminium hydroxide doi:https://doi-org.proxy.lib.ltu.se/10.1016/j.jenvman.2013.11.041

Maest, A. S., & Nordstrom, D. K. (2017). A geochemical examination of humidity cell tests. Applied Geochemistry, 81, 109-131. doi:10.1016/j.apgeochem.2017.03.016

Majima, H. (1969). How oxidation affects selective flotation of complex sulphide ores. Canadian Met-allurgical Quarterly, 8(3), 269-273. doi:10.1179/cmq.1969.8.3.269

Majzlan, J., Navrotsky, A., & Schwertmann, U. (2004). Thermodynamics of iron oxides: Part III. en-thalpies of formation and stability of ferrihydrite (~Fe(OH)3), schwertmannite (~FeO(OH)3/4(SO4)1/8), and e-Fe2O3. Geochimica Et Cosmochimica Acta, 68(5), 1049-1059. doi:10.1016/S0016-7037(03)00371-5

Mehta, A. P., & Murr, L. E. (1982). Kinetic study of sulfide leaching by galvanic interaction between chalcopyrite, pyrite, and sphalerite in the presence of T. ferrooxidans (30°C) and a thermophilic microorganism (55°C). Biotechnology and Bioengineering, 24(4), 919-940. doi:10.1002/bit.260240413

MEROX. (2015). Handbok hyttsten typ L väg- och anläggningsarbeten. Retrieved from http://www.merox.se/uploads/images/1020/Handbok_Hyttsten_typ_L__A4_Ver_4_Slut-lig.pdf

Miller, M. M., & Callaghan, R. M. (2004). Lime kiln dust as a potential raw material in portland cement manufacturing. (No. 2004-1336).U.S. GEOLOGICAL SURVEY. Retrieved from https://pubs.usgs.gov/of/2004/1336/2004-1336.pdf

50

Misra, M., Chen, S., & Fuerstenau, M.C. (2006). Passivation of acid mine tailings. Paper presented at the IMPC 2006 - Proceedings of 23rd International Mineral Processing Congress, 2388-2393.

Monnier, J., Bellot-Gurlet, L., Baron, D., Neff, D., Guillot, I., & Dillmann, P. (2011). A methodology for Raman structural quantification imaging and its application to iron indoor atmospheric corrosion products. J Raman Spectrosc. 42(4),773–781

Montelius, C. (2005). The genetic relationship between rhyolitic volcanism and Zn-Cu-Au deposits in the Maurliden volcanic centre, skellefte district, Sweden: Volcanic facies, lithogeochemistry and geochronology (Doctoral thesis). Retrieved from http://ltu.diva-portal.org/smash/rec-ord.jsf?pid=diva2%3A999019&dswid=-3423

Mäkitalo, M., Lu, J., Maurice, C., & Öhlander, B. (2016). Prediction of the long-term performance of green liquor dregs as a sealing layer to prevent the formation of acid mine drainage. Journal of En-vironmental Chemical Engineering, 4(2), 2121-2127. doi:10.1016/j.jece.2015.10.005

Mäkitalo, M., Mácsik, J., Maurice, C., & Öhlander, B. (2015). Improving properties of sealing layers made of till by adding green liquor dregs to reduce oxidation of sulfidic mine waste. Geotechnical and Geological Engineering, 33(4), 1047-1054. doi:10.1007/s10706-015-9886-4

N

Naturvårdsverket. (2020). Avfall i sverige 2018 - uppkomst och behandling. ( No. 6932). Retrieved from https://www.naturvardsverket.se/Documents/publ-filer/6900/978-91-620-6932-2.pdf?pid=26946

Nicholson, R. V., Gillham, R. W., & Reardon, E. J. (1988). Pyrite oxidation in carbonate-buffered so-lution: 1. experimental kinetics. Geochimica Et Cosmochimica Acta, 52(5), 1077-1085. doi:10.1016/0016-7037(88)90262-1

Nordstrom, D.K. (2011). Mine waters: Acidic to circumneutral. Elements, 7(6), 393-398. doi:10.2113/gselements.7.6.393

Nordstrom, D. K. (1982). Aqueous pyrite oxidation and the consequent formation of secondary iron minerals. In Kittrick,J.A.,Fanning,D.S.,Hossner,L.R (Ed.), Acid sulfate weathering, Soil Science Society of America. doi:10.2136/sssaspecpub10.c3

Nordstrom, D. K. (2011). Hydrogeochemical processes governing the origin, transport and fate of major and trace elements from mine wastes and mineralized rock to surface waters. Applied Geochemis-try, 26(11), 1777-1791. doi:10.1016/j.apgeochem.2011.06.002

Nordstrom, D. K., & Alpers, C. N. (1999). Negative pH, efflorescent mineralogy, and consequences for environmental restoration at the iron mountain superfund site, California. Proceedings of the Na-tional Academy of Sciences of the United States of America, 96(7), 3455-3462. doi:10.1073/pnas.96.7.3455

Nordstrom, D.K., Blowes, D.W., & Ptacek, C.J. (2015). Hydrogeochemistry and microbiology of mine drainage: An update. Applied Geochemistry, 57, 3-16. doi: 10.1016/j.apgeochem.2015.02.008

Nyavor, K., & Egiebor, N. O. (1995). Control of pyrite oxidation by phosphate coating. Science of the Total Environment, 162(2-3), 225-237. doi:10.1016/0048-9697(95)04467-F

Nyström, E., Kaasalainen, H., & Alakangas, L. (2017). Prevention of sulfide oxidation in waste rock using by-products and industrial remnants: A suitability study. Paper presented at the 13th Interna-tional Mine Water Association Congress – "Mine Water & Circular Economy – A Green Con-gress", Lappeenranta, Finland, 25-30 June 2017. , 2 1170-1178.

Nyström, E., Kaasalainen, H., & Alakangas, L. (2019). Prevention of sulfide oxidation in waste rock by the addition of lime kiln dust. Environmental Science and Pollution Research, 26(25), 25945-25957. doi:10.1007/s11356-019-05846-z

51

Nyström, E., Kaasalainen, H., & Alakangas, L. (2019). Suitability study of secondary raw materials for prevention of acid rock drainage generation from waste rock doi:10.1016/j.jclepro.2019.05.130

Nyström, E., Thomas, H., Wanhainen, C., & Alakangas, L. (2021). Occurrence and Release of Trace Elements in Pyrite-Rich Waste Rock. Minerals, 11(5), 495. doi:10.3390/min11050495

O

Ouyang, Y., Liu, Y., Zhu, R., Ge, F., Xu, T., Luo, Z., & Liang, L. (2015). Pyrite oxidation inhibition by organosilane coatings for acid mine drainage control. Minerals Engineering, 72, 57-64. doi:10.1016/j.mineng.2014.12.020

P

Parbhakar-Fox, A., & Lottermoser, B. G. (2015). A critical review of acid rock drainage prediction methods and practices. Minerals Engineering, 82, 107-124. doi:10.1016/j.mineng.2015.03.015

Park, I., Higuchi, K., Tabelin, C. B., Jeon, S., Ito, M., & Hiroyoshi, N. (2021). Suppression of arseno-pyrite oxidation by microencapsulation using ferric-catecholate complexes and phosphate. Chemo-sphere, 269, 129413. doi:10.1016/j.chemosphere.2020.129413

Park, I., Tabelin, C.B., Seno, K., Jeon, S., Inano, H., Ito, M., & Hiroyoshi, N. (2020) Carrier-microen-capsulation of arsenopyrite using Al-catecholate complex: nature of oxidation products, effects on anodic and cathodic reactions, and coating stability under simulated weathering conditions. Heliyon, 6(1). doi:10.1016/j.heliyon.2020.e03189

Park, I., Tabelin, C. B., Jeon, S., Li, X., Seno, K., Ito, M., & Hiroyoshi, N. (2019). A review of recent strategies for acid mine drainage prevention and mine tailings recycling. Chemosphere, 219, 588-606. doi:10.1016/j.chemosphere.2018.11.053

Park, I., Tabelin, C. B., Magaribuchi, K., Seno, K., Ito, M., & Hiroyoshi, N. (2018). Suppression of the release of arsenic from arsenopyrite by carrier-microencapsulation using ti-catechol complex. Jour-nal of Hazardous Materials, 344, 322-332. doi:10.1016/j.jhazmat.2017.10.025

Park, I., Tabelin, C. B., Seno, K., Jeon, S., Ito, M., & Hiroyoshi, N. (2018). Simultaneous suppression of acid mine drainage formation and arsenic release by carrier-microencapsulation using aluminum-catecholate complexes. Chemosphere, 205, 414-425. doi:10.1016/j.chemosphere.2018.04.088

Parkhurst, D. L., & Appelo, C. A. J. (1999). User's guide to PHREEQC – A computer program for speciation, batch-reaction, one-dimensional transport, and inverse geochemical calculations [com-puter software]. Denver Colorado: US Geological Survey Water-Resour. Invest. Rep.

Paton, C., Hellstrom, J., Paul, B., Woodhead, J., & Hergt, J. (2011). Iolite: Freeware for the visualisation and processing of mass spectrometric data. Journal of Analytical Atomic Spectrometry, 26(12), 2508-2518. doi:10.1039/c1ja10172b

Pearce, N. J. G., Perkins, W. T., Westgate, J. A., Gorton, M. P., Jackson, S. E., Neal, C. R., & Chenery, S. P. (1997). A compilation of new and published major and trace element data for NIST SRM 610 and NIST SRM 612 glass reference materials. Geostandards Newsletter, 21(1), 115-144. doi:10.1111/j.1751-908X.1997.tb00538.x

Pérez-López, R., Cama, J., Miguel Nieto, J., Ayora, C., & Saaltink, M. W. (2009). Attenuation of pyrite oxidation with a fly ash pre-barrier: Reactive transport modelling of column experiments. Applied Geochemistry, 24(9), 1712-1723.

Pérez-López, R., Cama, J., Nieto, J. M., & Ayora, C. (2007). The iron-coating role on the oxidation kinetics of a pyritic sludge doped with fly ash. Geochimica Et Cosmochimica Acta, 71(8), 1921-1934. doi:10.1016/j.gca.2007.01.019

52

Pérez-López, R., Quispe, D., Castillo, J., & Nieto, J. M. (2011). Acid neutralization by dissolution of alkaline paper mill wastes and implications for treatment of sulfide-mine drainage. American Min-eralogist, 96(5-6), 781-791. doi:10.2138/am.2011.3685

Peter, A. L. J., & Viraraghavan, T. (2005). Thallium: A review of public health and environmental concerns. Environment International, 31(4), 493-501. doi:10.1016/j.envint.2004.09.003

Plumlee, G.S. (1999). The environmental geology of mineral deposits. In The Environmental Geochem-istry of Mineral Deposits, Part A: Processes, Techniques, and Health Issues; Plumlee, G.S., Logsdon, M.J., Eds.; Reviews in Economic Geology; Society of Economic Geologists: Littleton, CO, USA, 6, 71–116.

Plumlee, G.S., Ludington, S., Vincent, K.R., Verplanck, P.L., Caine, J.S., & Livo, K.E. (2009). Questa baseline and pre-mining ground-water quality investigation, 7. A pictorial record of chemical weath-ering, erosional processes, and potential debris-flow hazards in scar areas developed on hydrother-mally altered rocks. U.S. Geol. Surv. Open-File Report 2006e1205

R

Ragnvaldsson, D., Bergknut, M., Lewis, J., Drotz, S., Lundkvist, A., Abrahamsson, K., & Fernerud, S. (2014). A novel method for reducing acid mine drainage using green liquor dregs. Environmental Chemistry Letters, 12(3), 443-447. doi:10.1007/s10311-014-0469-z

Rimstidt, J.D., Chermak, J.A., & Gagen, P.M. (1994). Rates of reaction of galena, sphalerite, chalcopy-rite and arsenopyrite with Fe(III) in acidic solutions. In: Alpers CN, Blowes DW (eds) Environmen-tal geochemistry of sulfide oxidation. American Chemical Society Symposium Series 550, Wash-ington, DC, 2–13

S

Sahoo, P. K., Tripathy, S., Panigrahi, M. K., & Md Equeenuddin, S. (2013). Inhibition of acid mine drainage from a pyrite-rich mining waste using industrial by-products: Role of neo-formed phases. Water, Air, and Soil Pollution, 224(11) doi:10.1007/s11270-013-1757-0

Salmon, S.U., & Malmström, M.E., (2002). MiMi-steady state, geochemical box model of tailings im-poundment. Application to impoundment 1, Kristineberg, Sweden, and prediction of effect of reme-diation MiMI 2002:2

Salomons, W. (1995). Environmental impact of metals derived from mining activities: Processes, pre-dictions, prevention. Journal of Geochemical Exploration, 52(1-2), 5-23. doi:10.1016/0375-6742(94)00039-E

Sand, W., Jozsa, P.G., Kovacs, Z.M., Săsăran, N., Schippers, A. (2007). Long-term evaluation of acid rock drainage mitigation measures in large lysimeters. J Geochem Explor, 92, 205–211

Satur, J., Hiroyoshi, N., Tsunekawa, M., Ito, M., & Okamoto, H. (2007). Carrier-microencapsulation for preventing pyrite oxidation. International Journal of Mineral Processing, 83(3-4), 116-124.

Savard, D., Bouchard-Boivin, B., Barnes, S.J., & Garbe-Schönberg, D. (2018). UQAC-FeS: A new series of base metal sulfide quality control reference material for LA-ICP-MS analysis. Proceedings of the 10th International Conference on the Analysis of Geological and Environmental Materials, Sydney, Australia, 8–13 July 2018.

SEPA (Swedish Environmental Protection Agency). (2010). Återvinning av avfall i anläggningsar-beten. (No. 2010:1). Retrieved from http://www.naturvardsverket.se/Documents/publikationer/978-91-620-0164-3.pdf

Singer, P. C., & Stumm, W. (1970). Acidic mine drainage: The rate-determining step. Science, 167(3921), 1121-1123.

Sirén, S., Maurice, C., & Alakangas, L. (2016). Green liquor dregs in mine waste remediation, from laboratory investigations to field application. Paper presented at the 12th International Mine Water

53

Association Congress – “Mining Meets Water – Conflicts and Solutions”, Leipzig, Germany, 11-15 July 2016. 706-713.

SIS, (2003). Characterization of Waste e Leaching - Compliance Test for Leaching of Granular Waste Materials and Sludges, Part 2: One Stage Batch Test at Liquid Solid Ratio of 10l/kg for Materials with Particle Size below 4 Mm (Without or with Size Reduction). Swedish Standard SS-EN 12457-2. Swedish Standard Institute, Stockholm, Sweden.

Stefánsson, A. (2007). Iron(III) hydrolysis and solubility at 25°C. Environmental Science and Technol-ogy, 41(17), 6117-6123. doi:10.1021/es070174h

Sulaymon, A. H., Faisal, A. A. H., & Khaliefa, Q. M. (2015). Cement kiln dust (CKD)-filter sand per-meable reactive barrier for the removal of cu(II) and zn(II) from simulated acidic groundwater. Jour-nal of Hazardous Materials, 297, 160-172. doi:10.1016/j.jhazmat.2015.04.061

Swedish Geological Survey. (2019). Statistics of the swedish mining industry 2018. ( No. 2019:2). Re-trieved from http://resource.sgu.se/produkter/pp/pp2019-2-rapport.pdf

T

Tabelin, C.B., Park, I., Li, X., Seng, S., Villacorte-Tabelin, M., Igarashi, T., Ito, M., Hiroyoshi, N. (2019). Development of advanced pyrite passivation strategies towards sustainable management of acid mine drainage. 2019 International Conference on Advances in Civil and Ecological Engineering Research 1–4 July 2019, Kaohsiung, Taiwan. doi: 10.1088/1755-1315/351/1/012010

Tariq, A., & Yanful, E. K. (2013). A review of binders used in cemented paste tailings for underground and surface disposal practices. Journal of Environmental Management, 131, 138-149.

Tolonen, E., Sarpola, A., Hu, T., Rämö, J., & Lassi, U. (2014). Acid mine drainage treatment using by-products from quicklime manufacturing as neutralization chemicals. Chemosphere, 117(1), 419-424. doi:10.1016/j.chemosphere.2014.07.090

V

Vandiviere, M. M., & Evangelou, V. P. (1998). Comparative testing between conventional and micro-encapsulation approaches in controlling pyrite oxidation. Journal of Geochemical Explora-tion, 64(1-3 -3 pt 1), 161-176. doi:10.1016/S0375-6742(98)00030-2

Vaughan, D. J. (2021). Sulfides☆. In D. Alderton, & S. A. Elias (Eds.), Encyclopedia of geology (sec-ond edition) (pp. 395-412). Oxford: Academic Press. doi:https://doi.org/10.1016/B978-0-12-409548-9.02903-1

Vriens, B., Peterson, H., Laurenzi, L., Smith, L., Aranda, C., Mayer, K. U., & Beckie, R. D. (2019). Long-term monitoring of waste-rock weathering at the antamina mine, peru. Chemosphere, 215, 858-869. doi:10.1016/j.chemosphere.2018.10.105

Vriens, B., Plante, B., Seigneur, N., & Jamieson, H. (2020). Mine waste rock: Insights for sustainable hydrogeochemical management. Minerals, 10(9), 1-38. doi:10.3390/min10090728

W

Wilson, S. A., Ridley, W. I., & Koenig, A. E. (2002). Development of sulfide calibration standards for the laser ablation inductively-coupled plasma mass spectrometry technique. Journal of Analytical Atomic Spectrometry, 17(4), 406-409. doi:10.1039/b108787h

Y

54

You, G., Yu, C., Lu, Y., & Dang, Z. (2013). Evaluation of the protective effect of polysiloxane coating on pyrite with electrochemical techniques. Electrochimica Acta, 93, 65-71. doi:10.1016/j.electacta.2013.01.098

Younger, P. L., Banwart, S. A., & Hedin, R. S. (2002). Mine water: Hydrology, pollution, remediation. Dordrecht: Kluwer Academic.

Yuniati, M. D., Hirajima, T., Miki, H., & Sasaki, K. (2015). Silicate covering layer on pyrite surface in the presence of siliconcatechol complex for acid mine drainage prevention. Materials Transac-tions, 56(10), 1733-1741. doi:10.2320/matertrans.M-M2015821

Yuniati, M. D., Kitagawa, K., Hirajima, T., Miki, H., Okibe, N., & Sasaki, K. (2015). Suppression of pyrite oxidation in acid mine drainage by carrier microencapsulation using liquid product of hydro-thermal treatment of low-rank coal, and electrochemical behavior of resultant encapsulating coat-ings. Hydrometallurgy, 158, 83-93. doi:10.1016/j.hydromet.2015.09.028

Z

Zhang, X., Borda, M.J., Schoonen, M.A.A., & Strongin, D.R.. (2003). Adsorption of phospholipids on pyrite and their effect on surface oxidation. Langmuir 19, 8787e8792.

Zhang, M., & Wang, H. (2017). Utilization of bactericide technology for pollution control of acidic coal mine waste. Adv.Eng.Res., 129, 667-670.

Zhang, Y. L., & Evangelou, V. P. (1998). Formation of ferric hydroxide-silica coatings on pyrite and its oxidation behavior. Soil Science, 163(1), 53-62.

Zhu, C., & Anderson, G.M. (2002). Environmental Applications in Geochemical Modeling. Cambridge University Press

DOCTORA L T H E S I S

Elsa N

yström Preventing A

cid Rock D

rainage Formation from

Sulfidic Waste R

ock Using Secondary R

aw M

aterials

Department of Civil, Environmental and Natural Resources EngineeringDivision of Geosciences and Environmental Engineering

ISSN 1402-1544ISBN 978-91-7790-825-8 (print)ISBN 978-91-7790-826-5 (pdf)

Luleå University of Technology 2021

Preventing Acid Rock Drainage Formation from Sulfidic Waste Rock

Using Secondary Raw Materials

Elsa Nyström

Applied Geochemistry

Tryck: Lenanders Grafiska, 136575

136575 LTU_Nystrom.indd Alla sidor136575 LTU_Nystrom.indd Alla sidor 2021-05-10 11:042021-05-10 11:04