Passive Treatments for Mine Drainage: A Guide for Early Researchers (SpringerBriefs in Applied Sciences and Technology) 3031320484, 9783031320484

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SpringerBriefs in Applied Sciences and Technology Cassandra Chidiac · Aaron Bleasdale-Pollowy · Andrew Holmes · Frank Gu

Passive Treatments for Mine Drainage A Guide for Early Researchers

SpringerBriefs in Applied Sciences and Technology

SpringerBriefs present concise summaries of cutting-edge research and practical applications across a wide spectrum of fields. Featuring compact volumes of 50 to 125 pages, the series covers a range of content from professional to academic. Typical publications can be: . A timely report of state-of-the art methods . An introduction to or a manual for the application of mathematical or computer techniques . A bridge between new research results, as published in journal articles . A snapshot of a hot or emerging topic . An in-depth case study . A presentation of core concepts that students must understand in order to make independent contributions SpringerBriefs are characterized by fast, global electronic dissemination, standard publishing contracts, standardized manuscript preparation and formatting guidelines, and expedited production schedules. On the one hand, SpringerBriefs in Applied Sciences and Technology are devoted to the publication of fundamentals and applications within the different classical engineering disciplines as well as in interdisciplinary fields that recently emerged between these areas. On the other hand, as the boundary separating fundamental research and applied technology is more and more dissolving, this series is particularly open to trans-disciplinary topics between fundamental science and engineering. Indexed by EI-Compendex, SCOPUS and Springerlink.

Cassandra Chidiac · Aaron Bleasdale-Pollowy · Andrew Holmes · Frank Gu

Passive Treatments for Mine Drainage A Guide for Early Researchers

Cassandra Chidiac Department of Chemical Engineering and Applied Chemistry University of Toronto Toronto, ON, Canada

Aaron Bleasdale-Pollowy Department of Chemical Engineering and Applied Chemistry University of Toronto Toronto, ON, Canada

Andrew Holmes Geosyntec Toronto, ON, Canada

Frank Gu Department of Chemical Engineering and Applied Chemistry University of Toronto Toronto, ON, Canada

ISSN 2191-530X ISSN 2191-5318 (electronic) SpringerBriefs in Applied Sciences and Technology ISBN 978-3-031-32048-4 ISBN 978-3-031-32049-1 (eBook) https://doi.org/10.1007/978-3-031-32049-1 © The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 This work is subject to copyright. All rights are solely and exclusively licensed by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, expressed or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Switzerland AG The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

Graphical Abstract

1 Prior to mine drainage

Preventative Strategies

Construction

Capping

Blending

Passivation

2

External Treatments Permeable Reactive Barriers Cd /Zn

Sulfide mineral oxidation

Leachate or pumped drainage

2

Co /Ni

Cu /Pb

Constructed Wetlands

At-Source Treatments

Ni

In-Pit

Saturated Rock Fill

Se/ Cr

Cu /Pb

Cd /Zn

Gravel Bed Reactors

No leachate Se

Adsorptive

Alkaline Co /Ni

Co /Ni

Cu /Pb

Cd /Zn

Biological Cu /Pb

Cd /Zn

Se/ Cr Se/ Cr

v

Contents

1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1 7

2 Acid Mine Drainage Prevention . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1 Construction Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2 Blending or Co-disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3 Physical Barriers or Capping . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4 Passivation/Microencapsulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.1 Organic Coating . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.2 Inorganic Coatings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.3 Organosilane Coatings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4.4 Carrier Microencapsulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

9 9 13 16 18 19 21 23 27 29

3 In-Situ Remedies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1 In-Pit Treatments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.1 Alkaline In-Pit Treatments . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.2 Immobilization In-Pit Treatments . . . . . . . . . . . . . . . . . . . . . . . 3.1.3 Biological In-Pit Treatments . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1.4 Mixed In-Pit Treatments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2 Saturated Rock Fills . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.1 Saturated Rock Fills (SRFs) Technology Description . . . . . . 3.2.2 Saturated Rock Fill Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

35 35 36 41 51 56 60 60 62 68

4 Ex-Situ Remedies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1 Anoxic Limestone Drains (ALD) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2 Permeable Reactive Barrier (PRB) . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.1 PRB Design: Associated Materials and Removal Mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.2 Reactive PRBs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.3 Alkaline PRBs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

73 73 75 76 77 83

vii

viii

Contents

4.2.4 4.2.5 4.2.6 4.2.7

Dispersed Alkaline Substrate (DAS) PRB . . . . . . . . . . . . . . . Adsorptive PRBs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological PRBs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Diffusion-Active (DAPRB)/Sulfidogenic Exchange System (SDES) PRBs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3 Constructed Wetlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.1 Subsurface Flow Constructed Wetlands . . . . . . . . . . . . . . . . . 4.3.2 Floating Constructed Wetlands . . . . . . . . . . . . . . . . . . . . . . . . . 4.4 Gravel Bed Reactors (GBRs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

85 88 91 95 98 102 106 107 111

5 Recommendations and Challenges . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1 Prevention of Mine Drainage Production . . . . . . . . . . . . . . . . . . . . . . . 5.2 On-Site Treatment of Mine Drainage . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.1 Saturated Rock Fill (SRF) and Gravel Bed Reactor (GBR) Recommendations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.2 In-Pit Treatment Recommendations . . . . . . . . . . . . . . . . . . . . . 5.2.3 Constructed Wetland Recommendations . . . . . . . . . . . . . . . . . 5.3 Capturing and Treating Mine Drainage Seepage . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

119 119 120

6 Outlook . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1 Source Control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2 Permeable Reactive Barriers (PRBs) . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3 Constructed Wetlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4 Saturated Rock Fills/Gravel Bed Reactor . . . . . . . . . . . . . . . . . . . . . . . 6.5 In-Pit Treatments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

133 133 134 135 136 136 137

127 128 129 130 131

List of Variables

Acronym APTMS 8-HQ AMD ALD ATP BOF BTA BCF B APS CME Cat CKD COD PSPT CW CND DAPRB DAS EPA FTIR GAC GBR HNT HA HRT LDH LCA LGO MTMS

Meaning 3-aminopropyltrimethoxysilane 8-hydroxyquinoline Acid mine drainage Anoxic limestone drains Attapulgite clay Basic oxygen surface Benzotriazole Bioconcentration factor Biodegradability C-aminopropyltrimethoxysilane Carrier microencapsulation Catechol Cement kiln dust Chemical oxygen demand Complex formed with PropSH and sepiolite Constructed wetlands Contaminated neutral drainage Diffusion-active PRB Dispersed alkaline substrate Environmental protection agency Fourier transform infrared spectroscopy Granular-activated carbon Gravel bed reactor Halloysite nanotubes Humic acid Hydraulic retention times Layered double hydroxide Life cycle assessment Low-grade nickel ore Methyltrimethoxysilane ix

x

m-ZVI n-ZVI NMD NPS OLD ORP PRB RCA Eh SRF SEM SeRB SAF SRB SDES TEOS TS TF VTMS VS XPS ZVI PropS-SH

List of Variables

Micro-ZVI Nano-ZVI Near neutral mine drainage N-propyltrimethoxysilane Oxic limestone drains Oxidative reductive potential Permeable reactive barriers Recycled concrete aggregate Redox potential Saturated rock fills Scanning electron microscopy Selenium reducing bacteria Submerged arc furnace Sulfate-reducing bacteria Sulfidogenic exchange system Tetraethyl orthosilicate Total solids Translocation factor Vinyltrimethoxysilane Volatile solids X-ray photoelectron spectroscopy Zero valent iron U-mercaptopropyltrimethoxysilane

Chapter 1

Introduction

Abstract This chapter introduces the concept of acid mine drainage and its derivatives (e.g., contaminated mine drainage) regarding their chemical characteristics and production conditions in mining and engineering sites. It further discusses the environmental and economical implications of mine drainage production and introduces the concept of active and passive treatments for its remediation. Keywords AMD · CND · Passive · Pyrite · Mining · Drainage

Mine drainage is a global problem that impacts 117 countries worldwide, with the affected mining areas totaling nearly 32,000 km2 (Fig. 1.1) [1]. Acid mine drainage (AMD), consisting of acidic waters rich in sulfides, iron, and metal(loid)s, typically havoc the targeted areas once oxidants (e.g., air, water) infiltrate sulfide minerals during mining operations (e.g., ground mining, strip mining, open-pit mining) or in engineering projects (e.g., road construction, foundation excavation, airport developments) [1, 2]. AMD is not only catastrophic to surrounding environments, but it imposes economic threats to the affected countries. For instance, in America alone, remediation costs were estimated at $12.1 to 26.5 billion USD (2022) for the identified abandoned mining sites [3]. Furthermore, for projects that commenced priorly to 2001, where it is not evident of the companies involved, the financial burden will instead be imposed on taxpayers [3]. Thus, it is pertinent to identify remediation strategies that mitigate environmental and economical impacts. Pyrite (FeS2 ) is a common, highly reactive sulfide mineral in natural ore deposits. In aerobic environments, it oxidizes in the following reaction [4, 5] 2+ 2FeS2 (s) + 2H2 O + 7O2 → 4H+ + 4SO2− 4 + 2Fe

(1.1)

which is catalyzed by acidophilic iron- and sulfur-oxidizing bacteria (e.g., Thiobacillus, Acidithiobacillus, and Leptospirillium) that flourish in AMD’s low pH environment and use sulfide minerals as electron donors for their growth [6]. Furthermore, the Fe2+ ions generated in (1.1) subsequently oxidize to Fe3+ in the following reaction catalyzed by chemolithotrophs, Metallogenium (pH = 3.5 – 4.5), © The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 C. Chidiac et al., Passive Treatments for Mine Drainage, SpringerBriefs in Applied Sciences and Technology, https://doi.org/10.1007/978-3-031-32049-1_1

1

2

1 Introduction

Fig. 1.1 A global estimation of affected mine areas, reproduced with permission from [1]

or Thiobacillus ferooxidans (pH < 3.5) [4], 4Fe2+ + O2 + 4H+ → 4Fe3+ + 2H2 O.

(1.2)

Fe3+ generation introduces a more potent oxidizing agent capable of generating oxidation rates that are 10–100 times greater than that with oxygen [6]. Accordingly, it causes the following cascade of acid and sulfate production [2], + 2FeS2 (s) + 14Fe3+ + 8H2 O → 15Fe2+ + 2SO2− 4 + 16H .

(1.3)

However, if pH levels exceed 4, Fe3+ precipitates in the solution and forms hydrated iron oxide in the following manner, Fe3+ + 3H2 O ↔ Fe(OH)3 (s) + 3H+ .

(1.4)

Thus, the overall reaction for pyrite oxidation at pH > 4 can be described as, FeS2 (s) +

7 15 + O2 + H2 O → 2SO2− 4 + Fe(OH)3 (s) + 4H 4 2

(1.5)

establishing that 4–16 mols of H+ is produced per mole of pyrite oxidized. Other sulfide minerals that are susceptible to similar oxidative pathways include acid-producing minerals: pyrrhotite (Fe(1−x) S); chalcopyrite (CuFeS2 ); arsenopyrite (FeAsS); as well as non-acid producing minerals sphalerite (ZnS) and galena (PbS). For instance, pyrrhotite (Fe(1−x) S) oxidizes in the following respective pathways

1 Introduction

3

when oxygen is the primary oxidant (1.6) or under acidic conditions when S2− surfaces are exposed (1.7) [2, 6, 7]  x + O2 + xH2 O → (1 − x)Fe2+ + 2SO2− Fe1−x S + 2 − 4 + 2xH 2 Fe1−x S(s) + 2H+ → (1 − x)Fe2+ + H2 S.

(1.6) (1.7)

Similarly, chalcopyrite oxidizes in the respective pathways when oxygen (1.8) and Fe3+ (1.9) are the primary oxidants under acidic conditions, CuFeS2 + 4O2 → Cu2+ + Fe2+ + 2SO2− 4

(1.8)

CuFeS2 + 4Fe3+ → Cu2+ + 5Fe2+ + 2S0 .

(1.9)

Arsenopyrite (FeAsS) follows similar oxidation rates to pyrite when oxidized by Fe3+ in the following reaction, − + 4FeAsS + 13O2 + 6H2 O → 4Fe2+ + 4SO2− 4 + 4H2 AsO4 + 4H .

(1.10)

While non-acid producing minerals, sphalerite (ZnS) and galena (PbS), exhibit the following oxidation routes when oxygen is the primary oxidant: ZnS + 2O2 → Zn2+ + SO2− 4

(1.11)

PbS + 2O2 → Pb2+ + SO2− 4 .

(1.12)

Understanding the microbial processes of these reactions is critical, as genera like Thiobacillus, can increase conversion rates by factors of hundreds to millions through their metabolic routes [6]. Bacteria can catalyze these oxidative processes in direct and indirect pathways. Bacteria directly oxidize sulfide minerals when they attach to the minerals’ surfaces and use their iron and sulfur constituents as electron donors for their growth. The metabolic uptake of these electron donors oxidizes the minerals and leaches their metal(loid)s and Fe3+ into surrounding waters [2, 8]. On the other hand, indirect oxidation occurs from bacteria oxidizing Fe2+ to Fe3+ and elemental sulfur to SO4 2− in solution, thereby creating oxidants that can oxidize surrounding sulfide minerals. Bacteria can also indirectly oxidize sulfide minerals by forming biofilms on their surfaces since Fe2+ can be oxidized to Fe3+ in the biofilm layer [9]. Furthermore, biofilms form microenvironments on sulfide minerals that maintain local circumneutral pH levels. Circumneutral pH levels create suitable growth conditions to proliferate microbial communities with greater density and diversity that can catalyze these oxidative reactions [10]. At the same time, oxygen depletes as

4

1 Introduction

these microbial communities oxidize sulfide minerals. Oxygen depletion creates local concentration gradients at the surfaces of sulfide minerals. Oxygen concentration gradients create a driving force for oxygen to diffuse towards the exterior of sulfide minerals to promote further oxidation. Subsequently, local thermal gradients generate since sulfide mineral oxidation is exothermic. Local thermal gradients create convective heat flows that pull more air toward the interior of sulfide minerals by advective mass transport. Local thermal gradients can also water vaporization that alters local pressures and induce barometric pumping. Thus, sulfide mineral oxidation is affected by oxygen and Fe3+ concentration, temperature, pressure, pH, redox potentials, and indigenous microbial communities [2]. Overall, AMD has low pH values (2.2–3.5), oxidizing redox potentials (Eh < 0), high concentration of sulfates (> 100 mg/L), and dissolved iron and metal(loid)s (Fe, Al > Zn, Cu > Cd > Pb) [11, 12]. However, contaminated neutral drainage (CND) or near neutral mine drainage (NMD) can form from neighboring carbonatebased minerals, aluminum hydroxides, or ferric oxyhydroxides minerals that induce progressive pore-water neutralization in mine wastes and underlying aquifers [2]. The quickest and most effective neutralization occurs from surrounding carbonatemineral dissolution (e.g., calcite (CaCO3 ), dolomite (CaMg(CO3 )2 ), ankerite (Ca(Fe, Mg)(CO3 )2 ), and siderite (FeCO3 )). If carbonates are present in excess, circumneutral pH levels can be attained by the following reaction that consumes H+ and produces bicarbonate production, CaCO3 + H+ ↔ Ca2+ + HCO− 3

(1.13)

or equivalently by the following H+ consumption reaction at low pH levels, CaCO3 + 2H+ ↔ Ca2+ + CO2 + H2 O.

(1.14)

CND and NMD are typically less toxic than AMD since Fe3+ will precipitate and remove surrounding metal(loid)s via co-precipitation and sorption [11]. Nonetheless, Zn, Cd, Co, and Ni in NMD and CND cause concerns since these metals neither coprecipitate nor adsorb onto precipitates at pH levels less than 7. Figures 1.2a and b illustrate reported compositions and pH levels for real mine drainage, where ion and pH ranges were estimated by extracting raw data values from articles in this MiniBook Series. Figure 1.2’s red lines represent mining effluent limits imposed by the Environmental Protection Agency (EPA), gathered using their Enforcement and Compliance History Online Effluent Charts service [13]. The large spread observed in composition stems from sulfide minerals containing different metal(loid)s and acid-producing characteristics. Moreover, AMD releases sulfuric acid and dissolved metal(loid)s that collectively destroy surrounding ecosystems [2, 6]. Specifically, metal contamination in surrounding environments creates long-term issues due to their stability in aquatic ecosystems and tendency to bioaccumulate in surrounding organisms. The extent of their toxicity and bioaccumulation is proportional to the given environmental conditions (e.g., pH, organic carbon), where their relevant mobilities determine their

1 Introduction

5

Fig. 1.2 a Typical ion composition of mine drainage found in terms of heavy metals, salts, and b pH levels. Red lines represent EPA regulations of various mine sites for each respective element [13]

bioavailability (Ni > Co > Cd > Zn > Cu) [14]. Metals persist and disrupt surrounding living organisms’ metabolic functions in ecosystems by accumulating in vital organs and inhibiting adsorption, and displacing minerals. Metal accumulations also cause oxidative stress in plants that can disrupt their physiology and morphology [6]. Co and Ni are especially problematic since they readily complex with dissolved organic

6

1 Introduction

matter and humic acids in surrounding soils. Consequently, Co and Ni are more readily uptaken by plants [14]. Furthermore, mine drainage contains high amounts of sulfates since they are chemically stable, soluble, and difficult to remove by geo-physio-chemical processes [5]. Sulfates are problematic since they consume essential metals necessary for eco-hydrological functions and can form hydrogen sulfide, a corrosive and toxic gas, by biological reduction pathways [15]. The acidity in AMD further eliminates the ability of plants to obtain nutrients for growth and affects the normal physiological function of affected aquatic species. Yet, once pH levels exceed 4, Fe(OH)3 precipitates and forms orange deposits along ocean floors and river beds [2]. These deposits block sunlight and cement nutrients required for the growth and health of surrounding wildlife [2]. These phenomena collectively create catastrophic effects in nearby environments that can eradicate entire ecosystems. Active and passive treatments use physiochemical, electrochemical, or biological mechanisms to remediate mine drainage. Active treatments require equipment, maintenance, continual chemical dosing, or power. These treatments offer precise control, high effectiveness, and limited space constraints [16]. However, they incur high capital and operating costs and are susceptible to toxic sludge formation. Common active treatments include aeration, membrane filtration (e.g., reverse osmosis, nanofiltration), selective precipitation, electrosorption, electrocoagulation, bioreactors, and sludge blankets [17, 18]. Due to the changing characteristics of an active mine site (e.g., flow rate, water chemistry), active treatments are optimal due to their precise controls and quicker treatment times [18]. Alternatively, passive treatments use biological and geochemical/geophysical mechanisms that negate the need for power, labor, energy, or equipment [18]. Thus, passive treatments minimize operating and capital costs and do not require power access in remote mine sites. These systems are less environmentally detrimental and reduce costs but necessitate larger land area and longer treatment times. Generally, passive treatments are well suited for closed mines due to their stable chemistry and flow rates compared to active sites [16]. While closed mines are generally more of an environmental hazard since active mine sites can maintain low water tables using continuous pumping [6]. Passive treatment systems can be preventive, in-situ, or ex-situ solutions. Preventative solutions (i.e., source control) occur before or during mining operations, including construction techniques, physical barriers, blending, and passivation. Insitu treatments include in-pit treatments, and saturated rock fills, while ex-situ solutions include permeable reactive barriers, constructed wetlands, and gravel bed reactors. Herein, each chapter critically analyzes each passive treatment technology’s mechanisms, current research trends, and relevant case studies. Additionally, a recommendation chapter discusses each passive treatment’s strengths based on water chemistry and factors such as land availability, flow rate, and leachate development. Finally, an outlook chapter describes each technology’s trends and promising research directions.

References

7

References 1. T. Liang, T.T. Werner, X. Heping et al., A global-scale spatial assessment and geodatabase of mine areas. Glob. Planet Change 204, 103578 (2021). https://doi.org/10.1016/j.gloplacha. 2021.103578 2. D.W. Blowes, C.J. Ptacek, J.L. Jambor, C.G. Weisener, The geochemistry of acid mine drainage. Treatise Geochem. 9–9, 149–204 (2003). https://doi.org/10.1016/B0-08-043751-6/09137-4 3. THE MINING BURDEN: States Would Shoulder Significant Costs of Cleaning Up Abandoned Mines if They Take Over American Lands (2015) 4. US Environmental Protection Agency, Acid Mine Drainage Prediction. US Environmental Protective Agency (1994), p. 52 5. J.A. Galhardi, D.M. Bonotto, Hydrogeochemical features of surface water and groundwater contaminated with acid mine drainage (AMD) in coal mining areas: a case study in southern Brazil. Environ. Sci. Pollut. Res. 23, 18911–18927 (2016). https://doi.org/10.1007/s11356016-7077-3 6. G.S. Simate, S. Ndlovu, Acid mine drainage: challenges and opportunities. J. Environ. Chem. Eng. 2, 1785–1803 (2014). https://doi.org/10.1016/j.jece.2014.07.021 7. B. Dold, Basic Concepts in Environmental Geochemistry of Sulfidic Mine-Waste Management (IntechOpen, 2010) 8. P.C. Singer, W. Stumm, Acidic mine drainage: the rate-determining step. Science 167, 1121– 1123. (1979) https://doi.org/10.1126/SCIENCE.167.3921.1121 9. G.S. Simate, S. Ndlovu, M. Gericke, Bacterial leaching of nickel laterites using chemolithotrophic microorganisms: process optimisation using response surface methodology and central composite rotatable design. Hydrometallurgy 98, 241–246 (2009). https://doi.org/ 10.1016/J.HYDROMET.2009.05.007 10. W.W. Barker, S.A. Welch, S. Chu, J.F. Banfield, Experimental observations of the effects of bacteria on aluminosilicate weathering. Am. Miner. 83, 1551–1563 (1998). https://doi.org/10. 2138/AM-1998-11-1243 11. D.K. Nordstrom, R.J. Bowell, K.M. Campbell, C.N. Alpers, Challenges in recovering resources from acid mine drainage 12. L.G. Toler, D.L. Peck, Some chemical characteristics of mine drainage in illinois 13. ECHO Tool Guide | ECHO | US EPA (2021). https://echo.epa.gov/resources/general-info/toolguide. Accessed 15 Feb 2022 14. K. von Gunten, B. Bishop, L. Zhang et al., Biogeochemical behavior of metals along two permeable reactive barriers in a mining-affected wetland. J. Geophys. Res. Biogeosci. 124, 3536–3554 (2019). https://doi.org/10.1029/2019JG005438 15. H. Wang, Q. Zhang, Research advances in identifying sulfate contamination sources of water environment by using stable isotopes. Int. J. Environ. Res. Public Health 16 (2019). https://doi. org/10.3390/ijerph16111914 16. D. Trumm, Selection of active and passive treatment systems for AMD—flow charts for New Zealand conditions. NZ J. Geol. Geophys. 53, 195–210 (2010). https://doi.org/10.1080/002 88306.2010.500715 17. M.J. Alegbe, O.S. Ayanda, P. Ndungu et al., Physicochemical characteristics of acid mine drainage, simultaneous remediation and use as feedstock for value added products. J. Environ. Chem. Eng. 7, 103097 (2019). https://doi.org/10.1016/j.jece.2019.103097 18. B. Rezaie, A. Anderson, Sustainable resolutions for environmental threat of the acid mine drainage. Sci. Total Environ. 717, 137211 (2020). https://doi.org/10.1016/j.scitotenv.2020. 137211

Chapter 2

Acid Mine Drainage Prevention

Abstract Source control suppresses mine drainage production in-situ by neutralization and/or creating physical barriers to mitigate the exposure of bacteria and oxidants (e.g., air, water) on sulfide minerals. Source control tactics are subdivided into construction techniques; blending/co-disposal; capping/physical barriers; and passivation (e.g., organic coatings, inorganic coatings, silane coatings, carrier microencapsulation, etc.). Herein each technique’s suppression mechanism is discussed along with their advantages, disadvantages, and their recent research advances and literary gaps. Keywords AMD · Blending · Co-disposal · Capping · Passivation · Microencapsulation · Silane · Organic · Neutralization · Catechol

Acid mine drainage (AMD) prevention (i.e., source control) is the suppression of sulfide mineral oxidation during mining operations. The available techniques include construction techniques, physical barriers, blending, and passivation techniques. Each of these techniques along with their respective strengths and disadvantages will be discussed in the subsequent sections and are further summarized in Table 2.1.

2.1 Construction Techniques Waste rock piles are a mixture of non-economical materials (e.g., overburden rock) produced and stored as by-products in mining operations. If comprised of sulfide minerals, waste rock piles oxidize when exposed to the atmosphere. Waste rock oxidation occurs by three modes of oxygen transport: diffusion, advection, and barometric pumping [3]. For instance, once air intercepts sulfide minerals, they oxidize and consume oxygen. Oxygen consumption creates a concentration gradient that propagates air diffusion to the surface of sulfide minerals for further oxidation. Sulfide

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 C. Chidiac et al., Passive Treatments for Mine Drainage, SpringerBriefs in Applied Sciences and Technology, https://doi.org/10.1007/978-3-031-32049-1_2

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Organic coating (e.g., linoleic acid, humic acids, organic waste carbon, biosolids, biochar)

Passivation

[35, 37–39]

(continued)

Alkaline material (e.g., alkaline clay, fly ash, • Alkaline material blended with sulfidic minerals; [8, 9, 11–16] phosphate alkaline waste, red gypsum, sugar foam, requires large dosages • Low aspect ratios and completely mixed blends non-acid producing sulfidic rock) outperform • Subpar to capping; cannot limit oxidant exposures • Maintained circumneutral pH in a field study (> 5 y) when coupled with capping

Blending/co-disposal

• Prevents water permeation in sulfidic mineral through hydrophobic surface coatings • Biological growth can be stimulated to deplete O2 and Fe3+ • Susceptible to biodegradation and instabilities

Water Covers • Water covers submerge sulfide minerals in shallow [10, 17, 18, 23, 26–28, 33, 34] water, often combined with dry covers Dry Covers (e.g., green liquor dregs, bentonite and • Covers placed over sulfide minerals to reduce derivatives, soil, vegetation, sludge, fly ash, oxygen and water permeation cement) • Sustains cycles of heating and cooling without cracking • Requires saturation conditions, non-suitable for arid regions

Physical barriers/capping

[2, 4–7]

• Reduce sloping (~ 5°) • Implement run-off channels with multiple benches • Compact horizontal layers with a max height of ~6m • Eliminate face dumping • Compaction under saturation conditions reduces void spacing

N/A

Construction techniques

References

Mechanisms/highlights

Materials

Type

Table 2.1 Summary of source control tactics, their respective materials, and mechanisms/highlights

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Type

Table 2.1 (continued)

• Organosilane coatings form via Fe–O–Si bonds on [47, 64–66, 68, 71] sulfide mineral rock, and a crosslinked network over top via Si–O–Si bonds • Increases surface hydrophobicity • Produced using sol–gel methods, requiring high temperatures (> 80 °C) and solvents • Organosilane coatings susceptible to microcracks unless under high loading • SiO2 nanoparticles improve passivation efficiency • Organic carrier increases metal(loid) solubilities to sulfide mineral surface where it is oxidized and forms a precipitate layer as oxides/hydroxides • Demonstrates high selectivity toward sulfide minerals • Susceptible to dissolution under acidic conditions • Should be supplemented with alkaline material

Organosilane-coatings (e.g., c-aminopropyltrimethoxysilane (APS) vinyltri-methoxysilane (VTMS), methyltrimethoxysilane, organosilane, PropS-SH)

Carrier-microencapsulation (e.g., Fe3+ -, Al3+ -, Si4+ -, Ti4+ -catecholates)

[73, 74, 76, 77, 79, 80]

[47, 50, 52–54, 58, 59, 62, 63, 81]

• Forms iron-(oxy)hydroxide, iron-phosphate, and iron-silica surface layers • Mitigates oxygen and water permeation but often requires H2 O2 and buffers • Silicates create a smooth, amorphous iron oxyhydroxide surface that is crack resistant • Performance diminishes under acidic conditions

Inorganic coatings (e.g., silicate, Na2 SiO3, CaSiO3 , KH2 PO4 , lime kiln dust, lime, Al(OH)3 )

References

Mechanisms/highlights

Materials

2.1 Construction Techniques 11

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mineral oxidation is a highly exothermic reaction, where highly reactive minerals such as pyrite can produce approximately 1409 kJ of heat per mole of pyrite [3]. Thus, internal thermal gradients generate as sulfide minerals undergo oxidation from diffusion. These thermal gradients create advective air transport within waste rock piles that can pull air into greater depths in waste rock piles for further oxidation. Additional oxidation at greater depths of the waste rock piles creates internal oxygen concentration gradients that drive oxygen diffusion further. Local thermal gradients also cause water vaporization and therefore changes in barometric pressures on the surface of waste rock piles [3]. Changes in the barometric pressure influence the pneumatic head, which induces gas venting. Thus, temperature, gas composition, and air pressure are the interdependent variables that govern sulfide mineral oxidation. Push- and end-dumping are construction methods implemented to store waste rock piles on site. Push dumping involves an ejector system that pushes the load (e.g., waste rock) toward the vehicle’s rear and ejects it. In contrast, end dumping uses a truck or trailer that unloads waste rock by raising a bed or container to let the load fall out of the back end. End dumping offers little control over particle distributions and other intrinsic characteristics; it can also create rubble zones and interspersed layers of coarse and fine particles that produce preferential air flows internally [5]. Thus, adopting a push-dumping construction method with bench implementation is preferred since it increases control and predictiveness in waste rock design. Waste rock pile designs must be stable under long-term storage. Thus, waste rock pile designs must consider the geometry and packing configuration of the waste rock piles since these factors determine their overall stability [1]. Stable geometries in waste rock pile designs involve subdividing the pile’s height into shorter, compacted layers with reduced global slope angles (< 18 degrees) and with two or more benches [1, 2]. These geometric changes can reduce the overall stress state of the pile as shown in Fig. 2.1. Mitigating water infiltration and air flows in waste rock piles can improve their long-term stability since water infiltration creates shallow slides or deep-seated failures, and air flows oxidize present sulfide minerals [2]. Thus, the construction of compacted layers should be under saturated conditions to enable better particle-particle interlocking [6]. Particle-particle interlocking reduces voids in the compacted layer, which reduces preferential pathways for air and water and structural instabilities. Similarly, coarse and high slopes are not suitable in waste rock design. Implementing coarse materials along the incline and bottom of the rock pile can promote gas venting and preferential air flows, respectively. High slopes also promote gas venting effects that can cause oxygen levels to rise by 11 to 21% [4]. Instead, short lifts (< 4–6 m) should be used to create the compacted layers since this creates vertical resistance to convective air transport [3, 4]. Furthermore, water infiltration can be mitigated by negating fine particles along the perimeters of waste rock piles. Specifically, fine particles promote water infiltration since they exhibit lower saturated hydraulic conductivities and retain water at a greater capacity [5]. At the same time, sloping the compacted layers toward the exterior mitigates the likelihood of water percolation from rainfall events since it diverts water flow laterally [2]. Water diversion is achieved partially from a capillary barrier effect and by maximizing run-off. Although, potential erosion can occur from

2.2 Blending or Co-disposal

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Crushed waste rock (reactive) Increased stability

Compacted waste rock (non-reactive)

Fig. 2.1 Schematic of construction techniques that can be implemented to increase waste rock pile stability. Adapted with permissions from [1]

heavy rainfall events. Lastly, physical barriers significantly reduce oxygen and water permeation in waste rock piles. The physical barriers may include non-reactive rock, soil, vegetation, or a combination. The subsequent section will highlight promising materials researched to date in further detail. Several case studies have demonstrated the efficacy of construction methods in reducing sulfide mineral oxidation effects in waste rock piles. For instance, a waste rock pile in the Rum June mine site in Australia contained reactive pyrite that oxidized and raised temperatures to >50 degrees Celsius [7]. The waste rock pile rehabilitation came from reshaping its slopes along the top and sides to 5 degrees and 1:3, respectively. At the same time, run-off channels and clay and vegetation physical barriers were implemented to prevent future water and air infiltration. These tactics successfully reduced the waste rock pile’s internal temperature to its original levels. Similarly, Martin et al. rehabilitated rock piles in Lac Tio open-pit mine in Quebec, Canada using construction methods [2]. Push-dumping construction techniques were implemented to create a 5° surface downslope, and a flow control layer comprised of compacted sand and crushed anthracite waste rock was added along the surface. The implemented construction techniques reduced air infiltration within the slope by 90% and reduced water percolation. Nonetheless, construction methods only mitigate sulfide mineral oxidation since gas convection is inevitable with permeable and reactive waste rock [5]. Therefore, additional source control strategies should be implemented with construction techniques.

2.2 Blending or Co-disposal Blending or co-disposal involves mixing alkaline amendments with crushed sulfide minerals to remove their acid-producing capabilities and neutralize any developed leachate. A balance between the acid-producing minerals and the alkaline materials’ dosages is required to remove waste rock’s acid-producing characteristics. Thus, preliminary testing on dosage ratios is necessary. The alkaline amendments should be saturated with low air and water permeability to fill voids and macro pores in

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waste rock. Reducing void volumes reduces oxidant influxes and saturation hydraulic conductivities by 100 times [2, 3]. At the same time, total blending with low aspect ratios reduces acidification rates by enabling better waste coverage and reducing exposed area (Fig. 2.2a) [8]. Blending should implement mixed rock sizes for better interparticle locking and, therefore, better packing. Small rock sizes exhibit high surface reactivities that accelerate AMD production and should not be implemented in blending techniques (Fig. 2.2b) [9]. Blending and co-disposal require higher amendment dosages than other source control tactics (e.g., capping) since it neutralizes the entire mass of waste rock rather than limiting oxygen diffusion at the surface. For instance, limestone has the capacity to suppress AMD production at Freeport Indonesia and Grasberg mine but requires a 25% ratio of layering and covering [10]. From the sheer volume required, the use of waste materials to create a circular economy is pertinent. Blending amendments, therefore, should be alkaline with neutralization capabilities, available in high quantities, and be a repurposed waste material to create a circular economy. Additionally, implementing waste materials eliminates the production of alkaline amendments (e.g., lime; limestone) associated with greenhouse gas emissions. Fly ash has gained recent traction as a blending amendment since it is a coal waste, exhibits self-hardening properties, and has a multitude of alkaline moieties. However, there are inconsistencies in fly ashes’ success in column experiments due to different sourcing that change their chemical composition. For instance, fly ashes

a

pH

+ desulfurized tailings aspect ratio

Time (days)

b

Fig. 2.2 a The evolution of pH in column tests with waste rock alone and when mixed with desulfurized tailings under different aspect ratios (narrow and wide), adapted with permission from [8] and b the evolution of pH under different blended waste rock sizes with lignite fly ash and with a fly ash cover, adapted with permission from [9]

2.2 Blending or Co-disposal

15

have incurred large pH drops, increased metal(loid) mobilities, and leached contaminants (e.g., Al) when sourced with low carbonate content (e.g., 20 kg-CaCO3 /ton) [11]. In comparison, other studies have shown that fly ash comprised of an abundance of alkali moieties (e.g., CaO, MgO, Na2 O, K2 O) can maintain pH levels above 10 and reduce leachate metal concentrations to < 10 mg/L (Ni, Mn, Fe, Cu, Zn, Cr) [12]. These discrepancies highlight the importance of quantifying carbonate and other alkali moieties in fly ash before its use and conducting leachate studies since fly ash can contain heavy metal(loid)s themselves and destabilize. Alternative waste materials that have shown promise for co-disposal/blending include alkaline phosphate wastes, red gypsum, benign/non-acid producing sulfidic rock, phosphate mine waste, sugar foam, lime kiln dust, and bark ash [8, 9, 13–15]. These amendments contain several alkali moieties (e.g., carbonate, phosphate, lime) that can neutralize mine drainage and immobilize leachate contaminants (e.g., Zn, Cu) via precipitation. However, As and Mn leaching can incur from local pH increases [11, 13]. Thus, preliminary leaching experiments of amendments are required. Implementing additional adsorbents is required if metal(loid)s mobilities increase with blending (e.g., red gypsum for As). Additionally, alkaline clay is a basic, non-hazardous waste from alumina refinery processes that have demonstrated long-term neutralization abilities with successful contaminant immobilization (Fe, Mn, and SO4 ) [16, 17]. Its success led to a field study with additional compacted soil and sand layer (i.e., capping mechanism). The amended layer retained water which limited oxygen diffusion and supported vegetation growth. Kinetic and field studies demonstrated that > 3% alkaline clay generated sufficient alkalinity that could maintain neutral pH levels and immobilize sulfates and metals (Fe, Mn, Cu, Zn, Ni, Pb, Cd, and Co) [16, 17]. Overall, this study shows the significance of using waste materials with high neutralization capabilities and which can reduce waste rock porosity. Nonetheless, blending amendments will exhaust over time unless they mitigate air diffusion and prevent further oxidation. Accordingly, blending is more effective with capping mechanisms (Sect. 2.3). Overall, current research repurposing alkaline waste material to co-dispose waste rock. Co-disposal’s success depends on the amendments’ carbonatecomposition/neutralization capacity and their hydraulic properties to mitigate oxidant exposure. Future studies should assess fly ash and other common repurposed waste amendment’s neutralization abilities as a function of its composition/carbonate makeup. Predictive models that can correlate amendments’ composition and their packing/particulate size to performance would save extensive testing and allow researchers to focus on promising materials. Lastly, the idea of adhesive agents, such as epoxy, for enhanced particle–particle locking has yet to be explored and may aid in co-blending’s efficacy.

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2.3 Physical Barriers or Capping Physical barriers or capping uses wet or dry covers to reduce oxidant (e.g., air, Fe3+ ) permeation into sulfide minerals. Wet covers flood sulfide minerals in shallow water to lower oxygen diffusion rates [18]. Simultaneously, indigenous microbes may consume the remaining dissolved oxygen in the water cover to establish anoxic conditions [19, 20]. Therefore, wet covers are suited for abandoned and underground mines with sufficient rainfall and minimal wind mixing. Although wet covers have shown short-term success, they cannot eliminate oxidation in the long-term and is therefore not recommended as a stand-alone tactic [21, 22]. Dry covers can seal air entry pathways in wet covers by creating secondary diffusion barriers and eliminating wind mixing. If dry covers are carbon-rich, they can also provide substrates to promote aerobic microbial metabolism. They are suitable for regions where hydrology is well understood and practical to implement (i.e., areas with heavy rainfall) since saturation conditions are required to prevent cracking [23]. Water retaining materials are preferred as dry covers to reduce cracking and to reduce oxygen diffusion rates. Often water retention layers include top soil since it retains water, mitigates leachate migration and supports plant growth capable of immobilizing metals via phytoaccumulation and/or phytoextraction [23, 24]. Typically, dry covers implement alkaline amendments that neutralize mine drainage as it forms in situ. Simultaneous neutralization prevents contaminant dissolution and creates secondary diffusion boundaries via contaminant precipitation. Nonetheless, a subsection of metal(loid)s (e.g., As, Mn) solubilities can increase upon local pH increases and may require supplementary amendments. Bentonite is an excellent candidate for dry covers since its an alkaline, absorbent swelling clay with an expansive nature and buffering capabilities [25]. Thicker bentonite layers (0.5 m) have shown good retention abilities for heavy metals (e.g., Cu, Zn) but overall were only predicted to achieve an 87% efficiency over thirty years due to their loss of absorbance over time [26]. Bentonite’s implementation with bentonite pastes tailings has shown promise [27]. Accordingly, 8% of natural bentonite within the blend reduced hydraulic conductivities to 10–9 cm/s and could maintain a low hydraulic conductivity under six cycles of freeze-thawing events. It is, therefore, apparent that bentonite works best in blends due to its loss of absorbance over time, while the use of tailings further mitigates capping costs and creates a circular economy. Similarly, green liquor dregs produced from pulp and paper mills serve as an excellent alkaline barrier due to their high abundance of alkaline moieties (e.g., carbonates, hydroxides), low permeability from their silty to clayey structure, water retention properties, and large-scale yearly production [28, 29]. Green liquor dregs’ efficacy is at par with the conventional amendment, lime, in buffering pH levels and precipitating and removing metals (Cd, Co, Cr, Mn, Ni, and Zn) in leachates [30]. Pilot tests conducted with green liquor dregs’ have demonstrated their capacity to reduce leachate volumes by 40% and concentrations (As, Cd, Pb, Hg) by 67–87% [28]. Thus, if pulp and paper mills reside nearby mining sites, green liquor dregs are excellent candidates as capping material. Furthermore, desulfurized tailings are

2.3 Physical Barriers or Capping

17

another commonly implemented dry cover material since their sulfur constituents consumes oxygen and retain water [31, 32]. These phenomena enable green them to maintain circumneutral pH levels, reduce oxygen fluxes up to 96%, and reduce leachates’ metal (Zn, Cu) concentrations substantially (> 87%). Nonetheless, AMD can form from desulfurized tailings if sulfur levels reach a threshold concentration. Organic amendments are often used for dry covers since they can proliferate indigenous microbes and support biofilm formation. Biofilms create a secondary transport barrier for oxidants and create a tactic for oxygen and Fe3+ consumption by promoting microbial metabolic routes. Nonetheless, organic dry covers can cause the dissolution and releasement of Fe-precipitates into environments since certain microbes use Fe in their metabolism. Additionally, organic dry covers need replenishing or supplementation with recalcitrant organic fractions due to their susceptibility to biodegradation. Preliminary testing under representative conditions can predict their lifetime to avoid their premature degradation. AMD sludge has been a research interest as an organic dry cover since it creates a circular economy in mine drainage treatment. Demers et al. studied AMD sludge as a stand-alone oxygen barrier when blended with silty soil [33]. The AMD sludge blend maintained circumneutral pH and all metals (e.g., Cu, Fe, Zn) below detection limits for nearly 2.4 years. In contrast, non-coated, exposed waste rock experienced metal concentrations as high as 5 mg/L in its leachate. Therefore, sludge is a costeffective solution, as one can use local waste sludge. However, metal(oid)s and other contaminants may leach from the repurposed sludge. Promising amendments demonstrate synergism when combined for mine drainage treatment (e.g., portland cement, fly ash, and sludge from lime treatments) [32, 34]. For instance, fly ash contains fine particulates that can permeate into sulfide minerals’ voids, neutralize developed leachates’ acidity, and induce secondary mineral precipitation. Simultaneously, sludge and cement form thick, non-porous surface caps on sulfide minerals to reduce oxidant infiltration and metal leaching (Fig. 2.3a, b). Sludge alone can maintain circumneutral pH levels and induce sulfide- and carbonateprecipitation. However, when combined with Portland cement, the paper mill’s sludge works synergistically with the cement by suppressing its dissolution and enabling longer durations of alkalinity generation. Selecting fly ash thickness layers requires care since thick layers (e.g., 50 cm) can induce pH levels that are too high for environments (pH ≥ 11) and can exhibit instabilities that release contaminants to the environment (e.g., Si). Fly ash characterization and preliminary testing are recommended to prevent recontamination of surrounding environments [9]. Repurposed local sludge waste is recommended with locally sourced alkaline waste materials to create thick surface caps on waste rock with neutralization capabilities. Coupling these materials with fly ash would be beneficial due to its ability to permeate in sulfide minerals’ voids. However, fly ash sourcing requires care to prevent instabilities. It would be interesting to take a closer look at the biological role in this source control. Sludge provides an abundance of indigenous microbes proliferating at the sulfide minerals’ surface. It would be interesting to see if microbial activities affect this design’s success and if there’s a potential for their optimization. There is a lack of characterization of the promising amendment’s hydrophobicity

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a

b

Fig. 2.3 a Portland cement mechanisms as a physical barrier, reproduced from [33, 34] and b the placement of acid mine drainage treatment sludge and soil on top of acid-producing waste rock as a capping mechanism, reproduced with permission from [33, 34]

and microscopic structure and the implementation of chemical modifications. Polymers may provide non-wetting surfaces with ease of functionality, thereby enabling fine-tuning in their microscopic properties (e.g., porosity, hydrophobicity, packing). Although polymers are more expensive than waste material, their use could offset future costs observed for conventional alkaline materials.

2.4 Passivation/Microencapsulation Passivation/microencapsulation techniques use inert organic or inorganic coatings to protect sulfide minerals from oxidizing agents such as water, oxygen, and bacteria (i.e., iron- and sulfur-oxidizing bacteria). Passivation is an attractive strategy since it targets the site of reaction (i.e., minerals’ surface), thereby reducing material requirements. Techniques include organic coatings, inorganic coatings, silane coatings, and carrier-microencapsulation, where a discussion of each tactic and their respective research findings/trends is herein.

2.4 Passivation/Microencapsulation

19

2.4.1 Organic Coating Organic coatings are natural, green passivators that limit water permeation into waste rock by their hydrophobicity [35]. Additionally, organic coatings provide carbon sources to support heterotrophic microbial growth. The growth of heterotrophs diminishes oxidizing agents (e.g., Fe3+ , O2 ) in sulfide minerals since microbes utilize them as their terminal electron acceptor for their growth and metabolism. Since organic coatings actively promote microbial growth, biofilms are also produced and can act as additional diffusion barriers. Consequently, organic coatings are susceptible to biodegradation. Thus, biodegradation kinetics need to be well-understood prior to their implementation. Labile organics should be avoided as organic coatings since materials, such as linoleic acid, can degrade within 48 h when coated on pyrite [35]. Organics with higher recalcitrance, such as tannic acid, have shown stability. Tannic acid is a nontoxic, natural polyphenol extracted from plants and fruits [36]. Its phenolic hydroxyl moieties quickly coordinate with pyrite’s Fe3+ ions for deposition, scavenge free radicals and metals, and act as a natural antibacterial agent. Once deposited, tannic acid forms a dense layer around pyrite. The formed layer increases pyrite’s surface charge resistance and reduces its dissolution substantially (> 75%) [36]. Nonetheless, the efficacy of tannic acid was tested under a short term test without biotic influences and therefore warrants further investigation. Likewise, biosolids, a recalcitrant organic amendment, have shown promise as an organic passivation layer when supplemented with calcite (Fig. 2.4) [37, 38]. Specifically, biosolids stimulated heterotrophic microbial growth, which depleted oxygen levels and created a secondary diffusion barrier through biofilm formation. It further improved microbial abundance and diversity, creating a microbial community capable of competing with iron-oxidizing bacteria and proliferating iron- and sulfurreducers [37]. Calcite water encouraged passivation by neutralizing pH levels and favoring heterotrophic growth and precipitation. Suppression of pyrite oxidation was successful since there was only 2 wt% of S detected after 84 weeks [37]. Biochar can suppress arsenopyrite and chalcopyrite dissolution [39, 40]. Biochar reduced the dissolution of chalcopyrite from A. ferooxidans exposure by adsorbing bacteria and forming a passivation layer comprised of jarosite and S2− /S0 [39]. However, it only reduced dissolution rates by 17.7% when optimized (300 °C pyrolysis temperature; 3 g/L). The difference in performance between biochar and biosolids is from biochar’s non-biodegradability that cannot flourish iron- and sulfur-reducers. Biochar as a biofilm carrier with a more biodegradable organic source (e.g., biosolids) is of interest to create an organic coating for pyrite and chalcopyrite. On the other hand, biochar suppressed arsenic release (95.4%) from arsenopyrite by reducing charge transfer resistance at the double layer and film resistance at the passivation layer [40]. The reduction in charge transfer resistance within the passivation layer enabled the adsorption and complexation of Fe and As on biochar to create a biochar-Fe-As(V) complex. Biochar is an attractive organic coating amendment since it is a green method that is effective under ambient conditions and does

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Fig. 2.4 Passivation layer developed via stimulating heterotrophic microorganism growth using biosolid organics. Reproduced with permission from [37]

not require solvents for its deposition. However, biochar was assessed under abiotic conditions and short exposure times, and thus future studies should test biochar under representative conditions. Humic acids (HA) are supramolecular structures ubiquitous in natural organic matter containing biomolecular fragments that bind together through hydrophobic interactions, cation bridging, and H-bonding. HAs are excellent candidates as organic passivators due to their hydrophobic nature and abundance of functional groups (e.g., phenolic, carboxyl, amino, etc.) [41]. Therefore, HAs were tested as organic passivators for galena, arsenopyrite, pyrite, and chalcopyrite [42–46]. Under aseptic conditions, HA reduced leachate metal concentrations to suitable levels with all tested sulfide minerals by adsorbing onto their surface and forming complexations with present metals (e.g., Pb, As) [42–46]. However, HA enhances cell adsorption and growth on pyrite and arsenopyrite with microbes present. Thus, HA accelerates their biodissolution and Fe3+ releasement [44, 46]. Nonetheless, since As(V) is less soluble than its reduced form, its oxidation upon Fe3+ exposure has been successful in its in situ immobilization [46]. On the other hand, the number of bacteria (A. ferrooxidans) decreased using HA as a passivator for chalcopyrite since organic acids can inhibit microbial growth via cytoplasmic acidification and reduce available substrates [45]. HA, therefore, can suppress or promote cell growth on sulfide minerals. The impact of HAs is likely influenced by the indigenous microbes present that have different sensitivities to environmental parameters. Thus, HA can suppress sulfide mineral dissolution but must be pre-assessed under representative conditions. Therefore, although organics are susceptible to biodegradation, they hold their merits and should be considered with other source control tactics (i.e., alkaline amendments). Future work should extend the testing period for biosolids and calcite

2.4 Passivation/Microencapsulation

21

while capturing the transient behaviors of the microbial community. Specifically, looking at the performance from a phylogenetic view may provide information necessary for optimization (i.e., optimal growth conditions). Additionally, biochar as biofilm carriers, when used with labile and recalcitrant organics, such as HA could improve passivation efficiency.

2.4.2 Inorganic Coatings Inorganic coatings form passivation layers on sulfide minerals by inducing surface precipitation. Surface precipitation is facilitated by adding either (1) coating agents with oxidants and buffers or (2) by adding alkaline materials with or without silicate to create and deposit iron-(oxy)hydroxide or sulfide precipitates. Several material considerations are required to mitigate inorganic coating’s environmental impact. For instance, strategies that use H2 O2 are expensive and have difficulties with its handling and storage. Using H2 O2 for deposition makes the microencapsulation method infeasible in large-scale applications [47]. Second, phosphate-based microencapsulates can cause eutrophication and, thus, require long-term stability [47]. Third, diethylenetriamine or triethylenediamine is toxic to microbes and should be avoided [47]. Lastly, inorganic coating materials experience little selectivity with sulfide minerals and require care toward material costs and optimal liquid-to-solid ratios to ensure its implementation remains economical. Iron oxyhydroxide coatings form when pH levels are buffered at near-neutral levels as pyrite oxidizes [48]. Iron oxyhydroxide coatings grow under two phases. First, iron oxyhydroxide colloids form and attach to pyrite surfaces, where only marginal oxidant transport reduction occurs. In the second phase, iron oxyhydroxide continues to precipitate and occupy the interstices around the colloid particles, creating secondary diffusion boundaries that reduce oxidant diffusion coefficients’ magnitudes by factors > 5. By the end of stage 2, pyrite oxidation rates are reduced by a function of the square root of time. Iron oxyhydroxide coatings can form continuous, coherent, and stable passivation layers if circumneutral pH levels (pH > 6) are maintained and have further shown to have metal(loid) (e.g., As) adsorption capabilities [49, 50]. However, alkaline-induced inorganic coatings have only been successful for moderately reactive, potentially acid-forming waste rock. They are unsuitable for highly reactive waste rock since acidification rates exceed the rate at which the passivation layer forms [50, 51]. Tested alkaline amendments for iron oxy(hydroxide) precipitation include limesaturated water, lime kiln dust, calcium-saturated-water, calcite saturated water, blended calcite, and biosolids extract water [50, 51]. All amendments could maintain circumneutral levels for the passivation layer to form apart from calcite-saturated water, which exhibited a lag time that induced low pH levels. Blended calcite offers a cheap solution with less maintenance, while biosolids extract and lime kiln dust are low-cost solutions that meet the same performance as the standard amendment, lime. 5% lime kiln dust achieved secondary mineral formation, as shown in Fig. 2.5a, that

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maintained pH levels at 6 and decreased metal concentrations (Cu, Zn, As, Pb, Fe, Mn, and S) by > 99.9% for over 109 weeks [52]. Thus, several waste amendments have sufficient alkalinity to support the formation of iron oxyhydroxide passivation layers. The amendment choice should be according to their availability and proximity to the mine site. Silicate-stabilized iron oxyhydroxide can change the morphology of iron oxyhydroxide and oxide phases from a porous, crystalline structure into a smooth, coherent amorphous layer [48, 53, 54]. Silicates achieve this morphological change as they adsorb and structurally get incorporated into the layer as FeSi(OH)3 and FeOSi(OH)2 − . Its structural incorporation in the layer inhibits phase transitions from amorphous iron oxyhydroxide surface to goethite and creates a more thermodynamically stable layer [48, 55]. Silicate-stabilized iron oxyhydroxide coatings are conventionally synthesized by a pre-oxidation step using H2 O2 and the addition of silicate and a buffer (e.g., acetate) to maintain pH levels between 6 and 7 [56, 57]. Na2 SiO3 has shown to be superior to CaSiO3 as a silicate source, achieving 97.7% suppression of pyrite oxidation in a 449-day testing period under optimal conditions and was further found robust to cold climates (−10 °C) [58, 59]. Optimal conditions were achieved with 0.1–1 mM Si, creating stable coatings with Fe hydroxides and adsorbed silicate species. SiO2 precipitates formed and reduced the passivation layer’s efficacy once concentrated exceeded optimal levels [57, 60]. Time and liquid-to-solid ratio are insignificant parameters in improving the passivation layer’s

a

b Fig. 2.5 a Ferrihydrate and goethite passivation layer formation upon the addition of lime kiln dust (LKD) at 52 and 103 weeks, adapted with permission from [52]; b The progression of Al(OH)3 polymerization on pyrite’s surface at 0-, 7- and 282-days, adapted with permission from [63]

2.4 Passivation/Microencapsulation

23

performance. Furthermore, silicates are more effective with galena than sphalerite as bi-sulfide minerals form galvanic cells [54]. However, they are unsuitable for arsenopyrite as they increase As mobilization [60]. Moreover, a major drawback of silicate-based coatings is their sensitivity toward pH, as the passivation layer quickly destabilizes with slight pH drops [54]. Additionally, their dependence on H2 O2 and buffers for synthesis further limits their use at a large scale since this incurs operational costs and creates difficulties with storage and handling. A recent study demonstrated the use of CaSiO3 without the use of H2 O2 or buffers [61]. The synthesis of the passivation layer initiates under natural oxidation to produce sulfuric acid and iron oxyhydroxide. Sulfuric acid reacted with CaSiO4 to form CaSO4 and Si(OH)4, while iron oxyhydroxide forms a ferric hydroxide coating with polymerized silica. CaSO4 further cemented the ferric hydroxide coating with additional polymerized silica, thereby creating a passivation layer capable of suppressing pyrite oxidation by 85% based on Fe leaching. Nonetheless, the new fabricated coating still experienced instabilities at pH < 4, yielding pH instabilities as a current drawback for the inorganic passivation layer. Like silicate-based passivation layers, phosphate layers form as pyrite oxidizes, creating a smooth passivation layer comprised of ferrous phosphate with Fe2+ and Fe3+ hydroxides and oxy-hydroxy sulfate phases [62]. KH2 PO4 is the common PO4 source used in studies with 0.01 M as its optimal concentration. Although phosphatebased passivation layers demonstrated a robust performance under cold climates (−10 °C), they require the same H2 O2 and buffers for deposition and exhibit pH instabilities. Due to their additional potential of eutrophication if destabilized, phosphate-based passivation layers are not recommended as a passivation technique. Aluminum is another promising inorganic amendment that oxidizes on the surface of pyrite and forms Al(OH)3 under circumneutral pH (Fig. 2.5b) [63]. Pyrite oxidation was reduced by approximately 98% over a 282-day testing period using an Al(OH)3 layer- however, the aluminum layers evolved by its continual dissolution and reprecipitation. A long-term deposition of 120 days enabled the 98% suppression. Furthermore, as pyrite oxidation persisted, released aluminum was found to be incorporated into goethite. As additional goethite formed, Al(OH)3 was expelled to the surface, creating a template for further Al(OH)3 polymerization. However, the aluminum precipitates formed were sensitive to pH and were colloidal, making them susceptible to dissolution. Further studies should look at mechanisms for converting precipitates into amorphous morphologies and blending aluminum with another alkaline amendment to ensure sufficient suppression within the first 50 days.

2.4.3 Organosilane Coatings Organosilane coatings are hydrophobic coatings with organic and inorganic components. The organics in organosilane coatings increase their flexibility, improve their crack resistance, and ensure compatibility with polymeric material/coatings [47]. The inorganics in organosilane coatings contain silicon atoms that strongly bind to

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sulfide minerals and eliminate H2 O2 requirements for deposition. However, organosilane coatings require higher dosages due to their lack of selectivity towards sulfide minerals. Additionally, the sol-gel process used for their curation requires solvents (e.g., ethanol) and high curing temperatures (> 80°C). Nonetheless, organosilane coatings remain a research interest due to their ease of design and resiliency under different pH levels and temperatures. For instance, Diao et al. studied the efficacy of tetraethyl orthosilicate (TEOS) and N-propyltrimethoxysilane (NPS) as organosilane coatings to reduce pyrite oxidation (Fig. 2.6a, b) [47]. TEOS formed Fe–O–Si and Si–O–Si bonds along the surface of pyrite as a protective coating. It ultimately degraded due to its susceptibility to nucleophilic attacks from OH− groups on the Si atom. NPS was fabricated with propyl groups instead of Si atoms to create a more hydrophobic, flexible passivation layer resilient against cracking. Accordingly, NPS created a dense network around pyrite, reducing its exposure to air, water, and bacteria [47]. Overall, NPS improved the suppression of pyrite oxidation from 58% (TEOS) to > 95% in 120 days. Scanning electron microscopy (SEM) confirmed that NPS created a smooth, crack-free surface resistant to oxidation from the iron-oxidizing bacteria, A. ferroxidans. Although promising, 120 days is an insufficient testing time to assess its efficacy and therefore requires further investigation with different acid-producing waste rocks. Different synthesis tactics should also be investigated to improve their environmental impact. Other organosilane coatings investigated are U-mercaptopropyltrimethoxysilane (PropS-SH), c-aminopropyltrimethoxysilane (APS), and vinyltrimethoxysilane (VTMS) [64]. Electrochemical tests (i.e., cyclic voltammetry) confirmed that each could suppress pyrite oxidation, while Fe-leaching tests confirmed that PropS-SH (89.2%) > VTMS (71.4%) > APS (49.4%). Due to its promise, PropS-SH was studied in a composite with SiO2 nanoparticles (Fig. 2.6d) [65]. 2 wt% of SiO2 improved Fe release by ~ 71.4% compared to the PropS-SH coating. The passivation layer efficiency increased from Si-O-Si covalent bonds created in the crosslinking of PropS-SH’s and SiO2 nanoparticles’ silanol and hydroxyl groups, respectively. Similarly, Gong et al. used sepiolite, a naturally fibrous clay mineral, as a filler for PropS-SH to create the complex PSPT [66]. Sepiolite was chosen as a cheap, nontoxic filler since its enriched with Si–OH groups along its exterior that can polymerize with organosilane coatings to create dense and stable passivation networks. Fourier transform infrared spectroscopy (FTIR) and X-ray photoelectron spectroscopy (XPS) confirmed sepiolite’s ability to embed itself within PropS-SH’s crosslinked network via oxygen bridging. Accordingly, 0.8 wt% of sepiolite created a uniform surface coating that enhanced its coverability and improved Fe dissolution by approximately 40% compared to PropS-SH. Surpassed this dosage, aggregation resulted, which disrupted the passivation layer’s integrity and, therefore, the ability to suppress its oxidation. Additionally, hydroxyapatite has been investigated as a PropS-SH filler since it can increase the network’s stability, capacity, and hydrophobicity by forming Si-O-Si

2.4 Passivation/Microencapsulation

25

a

b

c

d

e

f

Fig. 2.6 Mechanism and formation of a tetraethylorthosilane (TEOS) and b npropyltrimethoxysilane (NPS) along with their SEM images, reproduced with permission from [47]; c methyltrimethoxysilane (MTMS), reproduced with permission from [36]; d mercaptopropyltrimethoxysilane (PropS-SH) with SiO2 , reproduced with permission from [38]; and self-healing agents, halloysite nanotubes (HNT) and benzotriazole (BTA) with gammamercaptopropylmethoxysilane (PropS-SH) microencapsulation illustrating its formation of e a passivation layer and f its self-healing mechanism. Reproduced with permission from [68]

and Fe-O-Si bonds [67]. Nonetheless, SiO2 nanoparticles crosslinked with PropSSH performed best in reducing pyrite oxidation due to the SiO2 nanoparticles’ size, which can embed into voids and create denser crosslinked networks. Investigating nanoparticles with different sizes and shapes with varying degrees of crosslinks in the network can push this technology further. The micro properties of these nanoparticles may enable better interparticle bonding and tighter packing that can be advantageous to suppress sulfide mineral oxidation. Organosilane coatings are susceptible to microcrack formation that compromise the integrity of the passivation layer. To address this issue, Li et al. looked at the addition of self-healing agents, halloysite nanotubes (HNT), and benzotriazole (BTA) to extend the lifetime of PropS-SH coatings [68]. Fig. 2.6e, f illustrate the

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overall structure of the passivation layer and its healing mechanism. As such, HNT acted as a nanofiller due to its ability to embed within PropS-SH coatings and was further used to load, store, and release passivation agents. BTA was the passivation agent used due to its ability to form protective films and its sensitivity toward pH levels. Specifically, BTA does not release from the protective capsules under neutral to alkaline levels (pH ~ 6.1–10). Once pH levels decrease to 4, 3, and 2, BTA releases more rapidly at 27%, 35%, and 91%, respectively. Thus, there is a delay in BTA release until pH levels decrease or equivalent H+ concentrations increase to decompose Cu-BTA complexes and, in turn, open HNT plugs and release the BTA agent. Released BTA agents then repair microcracks by forming new protective coatings. Li et al. demonstrated that this could extend the efficacy of PropS-SH protective layers by nearly 30% when tested over 60 days [68]. A cost analysis revealed that this regenerable coating is more expensive than alternative passivation layers (e.g., PropS-SH), but its long-term protection may offset long-run costs. Similarly, Yu et al. investigated the microcrack regeneration abilities of a pHsensitive attapulgite clay carrier (ATP) containing an 8-hydroxyquinoline (8-HQ) nanofiller within a PropS-SH and TEOS organosilane coating layer [69]. First, the hydrolysis of organosilanes synthesized active silanol groups. Next, a dehydration condensation reaction created Si-O-Si crosslinked networks between ATP-HQ and the silanol groups. Finally, an inert and dense passivation film with Fe-O-Si bonds formed by Si–OH bonds reacting with pyrite’s hydroxyl groups. The network releases 8-HQ upon local damage to quickly bind and repair cracks. This system suppressed pyrite oxidation by 88.1% but longer durations under different conditions are required to assess its efficacy [69]. Overall, the continual regeneration of passivation layers is an exciting development and should extend to other materials. This strategy may solve major hurdles source control faces in the long term, where research should focus on improving the fabrication methods (i.e., eliminating solvents and high curing temperatures) while reducing the complexity and cost of the regeneration mechanism. Tannic acid was fabricated into an organosilane coating with PropS-SH under mild conditions to mitigate the environmental impact of the organosilane coatings’ fabrication process [70]. A dense, scale-like, hydrophobic film was fabricated on pyrite by co-polymerizing tannic acid’s benzoquinone derivatives with PropS-SH using Michael addition or Schiff base reactions that do not require high curing temperatures. The formed organosilane coating’s hydrophobicity and greater charge transfer resistance enabled it to outperform its counterparts and create a passivation efficiency of approximately 86% based on Fe releasement. However, its synthesis used solvents. Methyltrimethoxysilane (MTMS) is the only identified organosilane material synthesized with ambient temperatures and without solvents. MTMS was first fabricated under a water environment using a sol–gel process with high curing temperatures. Like TEOS, MTMS created Fe–O–Si and Si–O–Si bonds along the surface of pyrite that crosslinked to create a dense network (Fig. 2.6c) [71]. Its performance experienced approximately an 80% reduction in pyrite oxidation in a 7-h testing period. To improve its environmental impact, MTMS was curated with 3aminopropyltrimethoxysilane (APTMS) to create an APTMS-MTMS matrix with

2.4 Passivation/Microencapsulation

27

curing temperatures between 15–40 °C [72]. APTMS accelerated Si–O–Si formation via amino protonation, which enriched the crosslinked network with Si–O– Si and Fe–O–Si bonds and created a superhydrophobic surface. Using a 25 °C curing temperature, the APTMS-MTMS coating achieved a comparable passivation efficiency (78%) to its original fabrication process. Overall, the fabrication of organosilane coatings and their susceptibility to microcracks are organosilane’s drawbacks. Further research should continue to look for additional materials that synthesize under environmentally benign conditions. Other organosilane materials that may achieve this are the water-soluble organosilane reagents γ-glycidyloxypropyltrimethoxysilane and tetraethoxysilane [71]. The performance of these synthesized materials may further improve by SiO2 nanoparticle addition under different sizes and degrees of crosslinking. Furthermore, these newly synthesized materials require microcrack regeneration capabilities and reduced complexity and cost.

2.4.4 Carrier Microencapsulation Carrier microencapsulation (CME) is metal(loid)-oxyhydroxide coatings formed by adding metal(loid) ions Al3+ , Fe3+ , Ti4+ , or Si4+ with an organic carrier such as catechol. Organic carriers increase metal(loid) solubility to enable their transfer to the sulfide minerals’ surfaces. Once the metal(loid) reaches the sulfide mineral’s surface, it oxidizes along its surface and forms oxide and hydroxide precipitates that act as a protective barrier against oxidizing agents [73]. Catechol is the preferred organic carrier since its redox sensitivity allows for both greater metal(loid) ion solubilities and higher specificity towards sulfide minerals [74, 75]. Catechol’s specificity reduces dosage requirements and, therefore, costs. However, catechol is expensive itself and has been identified as a major disadvantage to carrier microencapsulation. To address this hurdle, Yuniati et al., investigated catechol from hydrothermally treated low rank coal waste streams in silicon-catechol (Si-Cat or equivalently Si(Cat)3 2− ) complexes [75]. Silicate polymerized along the surface of pyrite, as determined by FTIR and XPS. Specifically, it formed a stable, silica-quinone coating along the surface of pyrite and suppressed oxidation by 78% when treated for 6 h. Increasing treatment times created catecholate–Fe complexes while less resulted in incomplete coatings. Nonetheless, a stable layer with Fe–O–Si bonds formed along the mineral’s surface with a pH of 9.5 [75]. Its overall reaction commences by the hydrolysis of the Si-Cat complex into silanol and quinone. Silanol subsequently forms a passivation layer by forming hydrogen bonds with the surface of pyrite. The bulky nature of quinone likely reduced its passivation efficiency as it can reduce the degree of packing between chemical structures. Nonetheless, the low rank coal wasted stream catechol used was not compared with pristine catechol. Despite its cost, catechol remains the most common organic carrier used in research due to its high specificity towards sulfide minerals. Accordingly, Ti4+ was synthesized as a CME (Ti(cat)3 2− ) using catechol due to its stability under a broad

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range of pH levels (3–12) [76]. Figure 2.7a illustrates the CME’s decomposition mechanism, involving its adsorption, partial oxidation-intermediate formation, nonelectrochemical dissociation, and hydrolysis-precipitation [76]. Although successful in suppressing arsenopyrite oxidation, its slow decomposition of 25 days limited its feasibility. Thus, Park et al., used Al3+ -catechol and Fe3+ -catechol complexes as passivation layers on arsenopyrite due to their quicker dissolution kinetics (< 2 d) [74, 77]. Under a pH of 5, Al(cat)+ complexes formed and quickly adsorbed onto the arsenopyrite surface. Once the mineral oxidized, the complex decomposed and released Al3+ which formed Al-oxyhydroxide precipitated along the surface (Fig. 2.7b). Although precipitates formed along the surface, it only suppressed oxidation by approximately 83%. It has further experienced instabilities from acid generation when used with arsenopyrite for more than 15 days [78]. Thus, Al3+ precipitates are not an optimal material choice for CMEs.

a

b

c d

Fig. 2.7 Mechanism and formation of a iron-based carrier microencapsulations (CMEs), reproduced with permission from [76]; b aluminum-based CMEs, reproduced with permission from [74]; and c titanium-based CMEs, reproduced with permission from [73, 74]; d Cyclic voltammetry measurements for Fe-cat and Fe–Ti-Cat passivation layers, reproduced with permission from [80]

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Fe3+ was tested as an alternative material [73, 79]. Fe3+ -catecholate complexes oxidized on the surface of the sulfide minerals and, in turn, decomposed and formed iron oxyhydroxide passivation layers (Fig. 2.7c). Fe3+ is a peculiar choice, as it is an oxidizing agent with sulfide minerals and may have competing effects. Nonetheless, Fe3+ -catecholate complexes formed solid precipitates along the surface and suppressed oxidation between pH levels of 5–10 [73]. However, the Feoxyhydroxide/oxide layers were susceptible to dissolution under acidic conditions. To overcome the susceptibilities of Fe3+ -catecholate passivation layers, Park et al. tested its stability when added with phosphates simultaneously [79]. Using this method, Fe(cat)2 and PO4 formed stable complexes that oxidated and decomposed on arsenopyrite. Subsequently, FePO4 coatings formed on arsenopyrite with its oxidized products, Fe(cat) and PO4 . FePO4 coatings were stable under low pH and could suppress arsenopyrite oxidation by 65% after three days, where this suppression increased with time [79]. The efficacy of FePO4 coatings evolved over time since Fe(cat)2 would continually decompose and release Fe3+ for further FePO4 precipitation on arsenopyrite. As such, the FePO4 could suppress pyrite more than other tested Fe-catecholate complexes tested. It would be of interest to assess its efficacy in a long-term test (> 100 d) and ensure the stability of FePO4 since its destabilization would release Fe3+ (i.e., a potent oxidant) and PO4 that can cause eutrophication. Therefore, more stable CMEs comprised of Fe-Ti-O and Fe-oxyhydroxides were synthesized by adding Fe3+ and Ti4+ catechol complexes simultaneously [80]. Ti4+ precipitation was accelerated from 25 to 7 days using a molar ratio of 1:1:6 of Fe:Ti:Cat, where the production of Fe2 TiO5 dominated the CME layer. Cyclic voltammetry measurements in Fig. 2.7d demonstrated that the Fe–Ti-Cat complex eliminated the CME’s response to anodic and cathodic currents, indicating that the layer is electrochemically inert [80]. The electrochemical resistance of Fecat decreased from the control by 5-fold as confirmed by electrical impedance spectroscopy. Although promising, leaching experiments must be conducted with different acid-producing waste rocks under long term tests to assess the feasibility of the CME layer. Future studies should couple other CME materials, e.g., Al-cat and Ti-cat, to see which performs best and extend studies for longer durations and under different acid-producing waste rocks.

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43. S. Wang, K. Zheng, H. Li et al., Arsenopyrite weathering in acidic water: humic acid affection and arsenic transformation. Water Res. 194, 116917 (2021). https://doi.org/10.1016/J.WAT RES.2021.116917 44. H. Yang, W. Luo, Y. Gao, Effect of Acidithiobacillus ferrooxidans on humic-acid passivation layer on pyrite surface. Minerals 8, 422 (2018). https://doi.org/10.3390/MIN8100422 45. J. Wang, Y. Liu, W. Luo et al., Inhibition of humic acid on copper pollution caused by chalcopyrite biooxidation. Sci. Total Environ. 851, 158200 (2022). https://doi.org/10.1016/J.SCI TOTENV.2022.158200 46. D. Zhang, H. Chen, J. Xia et al., Humic acid promotes arsenopyrite bio-oxidation and arsenic immobilization. J. Hazard. Mater. 384, 121359 (2020). https://doi.org/10.1016/J.JHAZMAT. 2019.121359 47. Z. Diao, T. Shi, S. Wang et al., Silane-based coatings on the pyrite for remediation of acid mine drainage. Water Res. 47, 4391–4402 (2013). https://doi.org/10.1016/j.watres.2013.05.006 48. D.M.C. Huminicki, J.D. Rimstidt, Iron oxyhydroxide coating of pyrite for acid mine drainage control. Appl. Geochem. 24, 1626–1634 (2009). https://doi.org/10.1016/J.APGEOCHEM. 2009.04.032 49. C.B. Tabelin, R.D. Corpuz, T. Igarashi et al., Acid mine drainage formation and arsenic mobility under strongly acidic conditions: importance of soluble phases, iron oxyhydroxides/oxides and nature of oxidation layer on pyrite. J. Hazard. Mater. 399, 122844 (2020). https://doi.org/10. 1016/J.JHAZMAT.2020.122844 50. G. Qian, R.C. Schumann, J. Li et al., Strategies for reduced acid and metalliferous drainage by pyrite surface passivation. Minerals 7 (2017). https://doi.org/10.3390/min7030042 51. R. Fan, G. Qian, M.D. Short et al., Passivation of pyrite for reduced rates of acid and metalliferous drainage using readily available mineralogic and organic carbon resources: a laboratory mine waste study. Chemosphere 285, 131330 (2021). https://doi.org/10.1016/J.CHEMOS PHERE.2021.131330 52. E. Nyström, H. Kaasalainen, L. Alakangas, Prevention of sulfide oxidation in waste rock by the addition of lime kiln dust. Environ. Sci. Pollut. Res. 26, 25945–25957 (2019). https://doi. org/10.1007/s11356-019-05846-z 53. R. Fan, M.D. Short, S.J. Zeng et al., The formation of silicate-stabilized passivating layers on pyrite for reduced acid rock drainage. Environ. Sci. Technol. 51, 11317–11325 (2017). https:// doi.org/10.1021/acs.est.7b03232 54. J. Pope, D. Trumm, Controls on Zn concentrations in acidic and neutral mine drainage from New Zealand’s bituminous coal and epithermal mineral deposits. Mine. Water. Environ. 34, 455–463 (2015). https://doi.org/10.1007/s10230-015-0372-2 55. L. Dyer, P.D. Fawell, O.M.G. Newman, W.R. Richmond, Synthesis and characterisation of ferrihydrite/silica co-precipitates. J. Colloid. Interface Sci. 348, 65–70 (2010). https://doi.org/ 10.1016/J.JCIS.2010.03.056 56. Z. Tu, Q. Wu, H. He et al., Reduction of acid mine drainage by passivation of pyrite surfaces: a review. Sci. Total Environ. 832, 155116 (2022). https://doi.org/10.1016/J.SCITOTENV.2022. 155116 57. K. Mylona, E. Adam, K. Papassiopi, N. Xenidis, Suppression of pyrite oxidation by surface silica coating. J. Geosci. Environ. Protect. 2, 37–43 (2014). https://doi.org/10.4236/gep.2014. 24006 58. C.U. Kang, B.H. Jeon, S.S. Park et al., Inhibition of pyrite oxidation by surface coating: a long-term field study. Environ. Geochem. Health 38, 1137–1146 (2016). https://doi.org/10. 1007/s10653-015-9778-9 59. C.U. Kang, B.H. Jeon, R. Kumar et al., Stability of coatings on sulfide minerals in acidic and low-temperature environments. Mine. Water Environ. 36, 436–442 (2017). https://doi.org/10. 1007/s10230-017-0437-5 60. K. Kollias, E. Mylona, N. Papassiopi, S. Thymi, Application of silicate-based coating on pyrite and arsenopyrite to inhibit acid mine drainage. Bull. Environ. Contam. Toxicol. 108, 532–540 (1234). https://doi.org/10.1007/s00128-021-03310-8

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61. S. Wang, Y. Zhao, S. Li, Silicic protective surface films for pyrite oxidation suppression to control acid mine drainage at the source. Environ. Sci. Pollut. Res. 26, 25725–25732 (2019). https://doi.org/10.1007/s11356-019-05803-w 62. K. Kollias, E. Mylona, K. Adam et al., Characterization of phosphate coating formed on pyrite surface to prevent oxidation. Appl. Geochem. 110, 104435 (2019). https://doi.org/10.1016/j. apgeochem.2019.104435 63. Y. Zhou, R. Fan, M.D. Short et al., Formation of aluminum hydroxide-doped surface passivating layers on pyrite for acid rock drainage control. Environ. Sci. Technol. 52, 11786–11795 (2018). https://doi.org/10.1021/acs.est.8b04306 64. Y. Ouyang, Y. Liu, R. Zhu et al., Pyrite oxidation inhibition by organosilane coatings for acid mine drainage control. Miner. Eng. 72, 57–64 (2015). https://doi.org/10.1016/j.mineng.2014. 12.020 65. Y. Liu, X. Hu, Y. Xu, PropS-SH/SiO2 nanocomposite coatings for pyrite oxidation inhibition to control acid mine drainage at the source. J. Hazard. Mater. 338, 313–322 (2017). https://doi. org/10.1016/j.jhazmat.2017.05.043 66. B. Gong, D. Li, Z. Niu et al., Inhibition of pyrite oxidation using PropS-SH/sepiolite composite coatings for the source control of acid mine drainage. Environ. Sci. Pollut. Res. (2020). https:// doi.org/10.1007/s11356-020-11310-0 67. S. Yang, T. Luo, J. Fan et al., Performance and mechanisms of PropS-SH/HA coatings in the inhibition of pyrite oxidation. ACS Omega 6, 32011–32021 (2021). https://doi.org/10.1021/ ACSOMEGA.1C04793/SUPPL_FILE/AO1C04793_SI_001.PDF 68. D. Li, B. Gong, Y. Liu, Z. Dang, Self-healing coatings based on PropS-SH and pH-responsive HNT-BTA nanoparticles for inhibition of pyrite oxidation to control acid mine drainage. Chem. Eng. J. 415, 128993 (2021). https://doi.org/10.1016/j.cej.2021.128993 69. M. Yu, J. Feng, Q. Yang et al., Inhibition of organosilane/ATP@HQ self-healing passivator for pyrite oxidation. Chemosphere 287, 132342 (2022). https://doi.org/10.1016/J.CHEMOS PHERE.2021.132342 70. D. Li, X. Chen, C. Liu et al., Suppression of pyrite oxidation by co-depositing bio-inspired PropS-SH-tannic acid coatings for the source control acid mine drainage. Sci. Total Environ. 862, 160857 (2023). https://doi.org/10.1016/J.SCITOTENV.2022.160857 71. Y. Dong, W. Zeng, H. Lin, Y. He, Preparation of a novel water-soluble organosilane coating and its performance for inhibition of pyrite oxidation to control acid mine drainage at the source. Appl. Surf. Sci. 531, 147328 (2020). https://doi.org/10.1016/j.apsusc.2020.147328 72. Y. Dong, Z. Liu, W. Liu, H. Lin, A new organosilane passivation agent prepared at ambient temperatures to inhibit pyrite oxidation for acid mine drainage control. J. Environ. Manage. 320, 115835 (2022). https://doi.org/10.1016/J.JENVMAN.2022.115835 73. X. Li, N. Hiroyoshi, C.B. Tabelin et al., Suppressive effects of ferric-catecholate complexes on pyrite oxidation. Chemosphere 214, 70–78 (2019). https://doi.org/10.1016/j.chemosphere. 2018.09.086 74. I. Park, C.B. Tabelin, K. Seno et al., Simultaneous suppression of acid mine drainage formation and arsenic release by carrier-microencapsulation using aluminum-catecholate complexes. Chemosphere 205, 414–425 (2018). https://doi.org/10.1016/j.chemosphere.2018.04.088 75. M.D. Yuniati, T. Hirajima, H. Miki, K. Sasaki, Silicate covering layer on pyrite surface in the presence of siliconcatechol complex for acid mine drainage prevention. Mater. Trans. 56, 1733–1741 (2015). https://doi.org/10.2320/matertrans.M-M2015821 76. I. Park, C.B. Tabelin, K. Magaribuchi et al., Suppression of the release of arsenic from arsenopyrite by carrier-microencapsulation using Ti-catechol complex. J. Hazard. Mater. 344, 322–332 (2018). https://doi.org/10.1016/j.jhazmat.2017.10.025 77. X. Li, M. Gao, N. Hiroyoshi et al., Suppression of pyrite oxidation by ferric-catecholate complexes: an electrochemical study. Miner. Eng. 138, 226–237 (2019). https://doi.org/10. 1016/J.MINENG.2019.05.005 78. I. Park, C.B. Tabelin, K. Seno et al., Carrier-microencapsulation of arsenopyrite using Alcatecholate complex: nature of oxidation products, effects on anodic and cathodic reactions, and coating stability under simulated weathering conditions. Heliyon 6, e03189 (2020). https:// doi.org/10.1016/J.HELIYON.2020.E03189

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79. I. Park, K. Higuchi, C.B. Tabelin, S. Jeon, N.H. Mayumi Ito, Suppression of arsenopyrite oxidation by microencapsulation using ferric-catecholate complexes and phosphate. Chemosphere 269, 6 (2021) 80. X. Li, I. Park, C.B. Tabelin et al., Enhanced pyrite passivation by carrier-microencapsulation using Fe-catechol and Ti-catechol complexes. J. Hazard. Mater. 416, 126089 (2021). https:// doi.org/10.1016/j.jhazmat.2021.126089 81. K. Kollias, E. Mylona, N. Papassiopi, A. Xenidis, Development of silica protective layer on pyrite surface: a column study. Environ. Sci. Pollut. Res. 25, 26780–26792 (2018). https://doi. org/10.1007/s11356-017-0083-2

Chapter 3

In-Situ Remedies

Abstract In situ remedies refer to processes that treat mine drainage at the source, whether directly adding amendments as in-pit treatments or backfilling old mining pits to create in situ bioreactors (e.g., saturated rock fills). In-pit treatments can be further subdivided by the following amendment types and removal mechanisms: alkaline; adsorptive; biological, or mixed treatments. This section will outline the mechanisms of each technique, current research in their subfield, literature gaps and potential methods to advance each field, with a focus on waste by-products. Keywords Saturated rock fill · Biostimulation · AMD · Passive · Bioaugmentation · Precipitation · Adsorption · Alkaline

3.1 In-Pit Treatments In-pit treatments remove target contaminants and achieve neutralization by backfilling end pits directly with amendments. Strategies include neutralization, immobilization, bioremediation, and mixed treatments. Typically, in-pit treatments reduce remediation costs and extraction times by targeting contaminants at the source and become further attractive when there are non-accessible or non-excavatable drainage areas [1, 2]. Herein material research advances and their optimization for each treatment type is discussed. Overall, research trends focus on reducing the environmental impact and cost of amendments by sourcing and testing waste materials available near the affected mining site. Additionally, research focuses on optimizing amendments’ mix ratios, particle sizes, thermal and chemical modifications, and determining potential valueadded products from formed precipitates. The following section discusses typical alkaline amendments used and their efficacy. Table 3.5 summarizes highlights, mechanisms, and targets found from in-pit studies. Tables 3.1, 3.2, 3.3 and 3.4 are summaries of the in-pit technologies’ target metal(s) initial concentration(s) (mg/L), achieved removal percentage, and removal capacity (mg/g). The removal capacity was calculated from reported values if not given explicitly. © The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 C. Chidiac et al., Passive Treatments for Mine Drainage, SpringerBriefs in Applied Sciences and Technology, https://doi.org/10.1007/978-3-031-32049-1_3

35

36

3 In-Situ Remedies

3.1.1 Alkaline In-Pit Treatments Alkaline amendments neutralize mine drainage and remove metal(loid)s by inducing metal hydroxide precipitation. These amendments are generally added by plowing up to depths of 1 m in the mining pit and then tilling these materials into the tailings [3]. Alkaline substances increase the pH of the contaminated water by consuming H+ ions from solution through hydroxyl or carbonate addition. Increasing water pH levels reduces the targeted metals’ solubilities and invokes their precipitative removal. Trivalent metal ions (e.g., Fe3+ ) tend to precipitate around pH levels of 4–6 and are, therefore, well suited for alkaline treatments, while divalent ions (e.g., Ni, Cu) precipitate at higher pH values such as 8–10 [4]. Since divalent metals require pH levels above circumneutral levels, their removal often depends on co-precipitation or an additional post-treatment neutralization step to eliminate any disruption to receiving waters’ health. Co-precipitation is the binding of soluble metals (e.g., Ni at pH < 8) to precipitates by surface adsorption, mixed-crystal formation, occlusion, and mechanical entrapment. Drawbacks of co-precipitation include dosage requirements, cost, treatment longevity, and environmental impact [5]. Specifically, the produced precipitates and co-precipitates cause secondary pollution by forming toxic sludge. Sludge production is costly as it requires proper transport, treatment, and disposal. Sulfide precipitation is preferred over hydroxide precipitation since sulfide precipitates tend to be less dense and ease sludge transport and handling [6]. Limestone is the conventional alkaline amendment used for neutralization and precipitation [7]. However, limestone contributes to greenhouse gas emissions upon its production and requires continual dosing from armoring, a process in which the surface of limestone becomes coated with metal hydroxides [8]. Therefore, there is a present need to find suitable alternatives to limestone. Amendments derived from industrial wastes with carbonate moieties can overcome the shortfallings of limestone. Carbonate moieties are desirable since it enables lime-based materials to have long-term buffering capabilities [3]. Accordingly, carbide lime, a waste product generated from acetylene gas, made of 85–95% Ca(OH)2 , 1–10% CaCO3 , and 1–3% unreacted carbon and silicates, demonstrated sustained circumneutral levels that could reduce metal concentrations (Cu, Ni, Zn) to meet regulations [9]. Additionally, there was no armoring, marking it as a suitable lime-based alternative. Table 3.1 summarizes the performance of lime-based materials.

3.1.1.1

Industrial By-Products

Alkaline amendments derived from industrial byproducts create opportunities to repurpose waste and create a circular economy in the mining industry. Hurdles faced in implementing industrial byproducts include their research and development, sourcing, long-term performance, and potential contaminant leaching. For instance, off-speciation fly ash from coal power plants, cement kiln dust, and recycled concrete

3.1 In-Pit Treatments

37

aggregate (RCA) has shown promise in AMD treatment [10–13]. RCA and cement kiln dust are construction materials rich in CaO content with additional surface functionalities (e.g., K–O, Si–O, Ca–O) that enable adsorptive and precipitative removals. Their fine particulate sizes quickly neutralize AMD and can remove metal targets (e.g., Fe, Cu, Zn, Cr) [10, 11] RCA and cement kiln dust’s performance is in Table 3.1. Similarly, fly ash is a coal combustion product comprised of fine particulates capable of generating alkalinity. Fly ashes’ high surface can quickly achieve neutralization in AMD and form synergies with other amendments [12, 13]. It further creates opportunities for synthesizing pervious geopolymers, which are alkaline aluminosilicate materials with enhanced porosity and adsorption capacities [14]. Successful geopolymers for Cu remediation include those derived from bagasse and coal-fly ash [14]. However, fly ash and its derived geopolymers commonly leach contaminants into treated waters and therefore require preliminary testing. Oxygen furnace slags are waste products from steelmaking that exhibit high CaO and MgO content and, therefore, neutralization capacities [15]. Their structure undergoes a slow dissolution process that can sustain elevated pH levels for three months and remove hard-to-treat metals (Co, Ni, and Zn) by metal hydroxide precipitation [15]. Its MgO functional groups form hydroxides, carbonates, and magnesiumsilicate-hydrate that are effective in removing a diverse range of contaminants from mine drainage (e.g., As, Cd, Ni, Zn, and Co) by sorption, co-precipitation, and precipitation methods (see Table 3.1) [16]. Basic oxygen furnace (BOF) slags, submerged arc furnace (SAF) slags, and BOF sludge have both been identified as effective neutralizers (pH > 9) and capable of removing Ni, Cu, and Co [17]. In particular, BOF sludge is an excellent candidate as an alkaline amendment since it removes sulfates effectively, transforms hazardous waste into inert waste, and incorporates circular economy practices for BOF use. Furthermore, syntheszing slags into geopolymers increases their porosity and sorption capacities, but they quickly saturate [18]. Overall, oxygen furnace slags offer a feasible option for remediation if local waste materials are available. It further creates circular economy since its sludge is a promising amendment. However, Mn or sulfates may require additional treatment, while Si, Mg, and Ca leaching can occur [15, 17]. The regeneration of geopolymers is further needed since they saturate in the short term. Specific performance of slags as an alkaline amendment is in Table 3.1. Bayer precipitates, a waste material from alumina refineries, are excellent alkaline amendments due to their high abundance of CaCO3 and hydrotalcite (magnesium– aluminum hydroxycarbonate). Bayer precipitates can remove a range of metals (Cu, Al, Zn, Fe, Ni) since they form hydroxide layers in the solution that can adsorb metals and generate alkalinity at the same capacity as lime [4]. Additionally, the thermal treatment of Bayer precipitates reduces hydrotalcite crystallinity and creates layered double hydroxide (LDH) structures with increased surface area capable of higher metal(loid) removal [19]. These phenomena collectively eliminated Al dissolution at higher pH levels and created hydrotalcite and LDH components that are more effective than lime.

38

3 In-Situ Remedies

LDH and hydrotalcite precipitates have caught the attention of researchers as an effective means to enhance metal removal with the potential for recovery. Hydrotalcite precipitates can form by adjusting the divalent to trivalent metal ratios (2 to 1). Their formation can improve removal capacities by surface adsorption and anion exchange [20]. Similarly, another study increased pH levels using NaOH and trivalent metal concentration to promote hydrotalcite production instead of gypsum and further produced a recoverable Cu/Zn ore-grade hydrotalcite precipitate capable of offsetting remediation costs [8]. Future research should investigate alkaline waste materials’ ability to produce ore-grade precipitates and increase pH sufficiently, as NaOH is not economical for large-scale treatment. Additionally, modifying the cation ratio by adding salts could be detrimental if salinity levels reach toxic levels and can be hard to control if continual metal dissolution occurs in the tailings. Bayer precipitates and hydrotalcite precipitation is in Table 3.1. Agri-foods are waste materials derived from agricultural food sources that are attractive as AMD amendments due to their high carbonate content. For instance, eggshells, a waste material directed to landfills, are a calcareous material made of 95% CaCO3 [21]. Their high porosity and enrichment in calcium and hydroxide moieties make them effective AMD buffers, particularly when amended with cement byproducts and MgO [21]. Their thermal activation by calcination can further improve their sorption capacities and surface areas by modifying their pore structures [22]. In particular, these amendments exhibited higher removal with increasing calcination temperatures and can meet targets for specific metals by adjusting liquid-tosolid ratios. Thus, eggshells offer a “green”, cost-efficient method that can neutralize AMD and remove metal(loid)s by precipitation, co-precipitation, and adsorption. Similarly, biomass ashes produced from food and crop incineration are promising amendments due to their high phosphorus and calcium content. For instance, straw, meat, bone meal, and poultry litter-derived biomass ashes from a biomass power plant showed promise for AMD treatment [23]. Straw and poultry ash maintained circumneutral pH levels and removed Zn and Fe. Both materials contained high amounts of CaO capable of neutralization and phosphorous compounds capable of initiating the precipitation and co-precipitation of target metal(oid)s. While all of the tested biomass ashes showed potential for AMD treatment due to their richness in CaO and phosphate groups, straw and poultry litter were the most effective. Overall, agri-food waste materials rich in CaO, carbonates, and phosphates are excellent candidates for AMD treatment. Future research should speciate agricultural wastes to determine the complex system of mechanisms and include trials with economic evaluations in lieu of typical alkaline materials. Agri-food alkaline material performance can be seen in Table 3.1.

3.1.1.2

Earth Minerals

Earth minerals are attractive amendments for mine drainage due to their adsorptive properties, alkalinity, and ability to invoke the precipitation and co-precipitation of contaminants. Thus, calcite (CaCO3 ), amorphous Fe(OH)3 , goethite (FeO(OH)), and

3.1 In-Pit Treatments

39

rhodochrosite (MnCO3 ) were assessed for contaminated neutral drainage (CND) treatment [24]. Calcite effectively co-precipitated and adsorbed targeted metals, where its carbonate moieties’ strong affinity for trace metals yielded its performance. However, calcite was less effective than goethite and amorphous Fe(OH)3 at removing Fe and Zn via precipitation. Thus, although calcite can remove metal(loid)s, alternatives can have beneficial differences in metal affinities. This subsection, therefore, discusses earth mineral types that show promise for mine drainage treatment. The specific performance of promising materials can be seen in Table 3.1. For instance, cryptocrystalline magnesite increased AMD pH levels and effectively removed Cu, Zn, Co, and Ni through hydroxide and sulfate precipitation [25]. Commercial value metals were also recovered by selectively precipitating metals in the solution using pH adjustments. Moreover, marl, sandstone, and calcareous crust could treat highly acidic AMD with their carbonate content enabling precipitation [26]. Other researchers look to modify minerals to improve their properties for AMD treatment. For example, polyoxymethylene sorbitan fatty acid ester (i.e., Tween 80) and sodium dodecyl can alter the charge of Mg(OH)2 to create a stable suspension and prevent its retention in negatively charged porous media [2]. Modifying the charge of Mg(OH)2 allows its migration through contaminated media and increases pH steadily through the slow release of hydroxide ions in solution over time. Slowing the release of hydroxide ions in the solution enables circumneutral pH levels to be sustained and mitigates the volume of sludge waste produced. Waste by-products in the mineral industry have been a recent research interest to create a circular economy within the mining industry. For instance, alkaline clay and coal refuse (10:90 ratio) raised and sustained pH levels for > 2 years, completely removing Co, most of Ni, Zn, Cu, and a significant reduction in sulfates [27]. Additionally, the implementation of grass in a subsequent study further stabilized and immobilized metals by phytoaccumulation and phytoextractions mechanisms; while implementing sand reduced neutralization times by increasing retention times and reducing oxygen and water diffusion (Fig. 3.1) [28]. Combining waste byproducts with sand and plants maintained low metal concentrations for 25 months by integrating in-pit alkaline treatments with a capping mechanism (See Chap. 2). Moreover, low-grade nickel ore (LGO), containing goethite, calcium oxide, and magnesium aluminum oxide, has similar properties to fly ash and limestone from its high specific surface area and site density [29]. Thus, LGO increased AMD pH levels to 5.4 and removed > 93% of Al, Fe, Ni, and sulfates by co-precipitation of Al hydroxides. The sulfate removal of LGO superseded limestones and therefore is a preferred substitute when sulfate removal is a goal. Alternative mine wastes studied are phosphate-limestone wastes, raw-grade phosphate ore, and phosphate mine tailings [30]. The combined wastes increased pH levels from 3 to 8.5 and removed metals through precipitation and sorption. Finally, charcoal ash leachate has been investigated for AMD and resource recovery [31]. Charcoal ash leachate released soluble salts that achieved neutralization when introduced to AMD sparingly. Iron, sulfate, and magnesium nanoparticles

40

3 In-Situ Remedies

Fig. 3.1 An example of different treatment combinations that can improve metal removal from AMD, where CR represents coal refuse and AC represents alkaline clay. Adapted with permission from [28]

were recovered from their formed sludge by precipitation and adsorption mechanisms, which reduced sludge volumes. Future work should identify materials and processes that can recover metals to create a circular economy, such as selective precipitation, hydrotalcite, or LDH adsorption. Table 3.3 shows the initial concentrations (mg/L) of each target metal, and the percent removal, and removal capacity (mg/g) achieved by each material. In summary, alkaline in-pit treatments effectively remove Fe, Al, and Zn. Since Al redissolves under high pH increases, modifications like thermal activation can be used to prevent redissolution, such as in the case of Bayer precipitates [19]. Materials that induce high pH influxes (>10) and have high sorption affinities can remove Co effectively. On the other hand, Ni exhibited limited removal, where quicklime, hydrotalcite precipitation, cryptocrystalline magnesite, and cement kiln dust

3.1 In-Pit Treatments

41

performed best. Cd and Cu were removed with hydrotalcite precipitation and limestone amendments, while sulfate removal was variable, with LDH showing promise. Chemical modifications may also serve as sulfate enhancers to alkaline amendments if necessary. However, if insufficient, biological in-pit treatments offer a viable alternative for sulfate removal. Furthermore, while traditional chemicals like quicklime and NaOH tend to remove large amounts of contaminants, their economic and environmental impact are burdensome. Promising alternatives are available, dependent on water chemistry and the location of the mining site. Ongoing research focuses on waste materials and their modifications that can be effective neutralizers for AMD. Laboratory column and in situ, scale-up tests are recommended for locally available materials to assess their ability to remediate mine drainage under representative conditions. Running preliminary tests reduces costs and the use of virgin materials in the remediation effort while still meeting guidelines. Further research is trending toward technical and economic evaluations of locally available materials for specific mine waste treatment, their removal mechanisms, resource recovery by selective precipitation, LDH, or hydrotalcite formation, improving long-term capacities, modifying promising materials, and using geochemical modeling to understand long-term performance.

3.1.2 Immobilization In-Pit Treatments In-pit immobilization uses materials with high adsorptive and ion-exchange capabilities to immobilize target contaminants. Adsorption occurs physically and chemically. Physical adsorption involves physical and van der Waals forces of attraction to collect ions on the surface of adsorbents. Chemical adsorption relies on chemical bonding (e.g., dipole–dipole interactions) specific to the target ions and the adsorbent’s surface. Desired adsorbent properties include regeneration capabilities, high porosity, and small pore diameters to enable reusability, more available surface area, and adsorptive sites. Metal(loid) immobilization by adsorption is generally suitable for CND and NMD (i.e., pH ≥ 4) due to its susceptibility to desorption and ion competition with H+ ions. Therefore, adsorbents treating AMD are typically alkaline to decrease H+ ions and, thus, ion competition. Furthermore, treating dilute NMD and CND is recommended since adsorbents’ active sites quickly saturate and can experience desorption with highly concentrated mine drainages. At the same time, their regeneration with benign chemicals reduces the treatment’s costs while improving its repeatability and environmental impact. Overall, the current research is testing adsorbents’ regeneration, coupling synergetic adsorbents, modifying surface chemistry, targeting trace metal(loid) removal, and investigating organic and waste substrates with high capacity for immobilization. Adsorbent materials that have been investigated for mine drainage are summarized in Table 3.2.

–

27.5 mg/L

99.99% –

–

201 mg/L

99.98%

–

–

98.3%

–

0.3 mg/L –

–

–

94%

–

–

0.0029 mg/g

628 mg/L

–

Cement kiln dust (pH = 3 → 8–10.8)

–

–

97.5%

0.3 mg/L

12.5 mg/g

45–100%

1–1000 mg/L

–

0.3986 mg/g

100%

100%

0.58 mg/L

94 mg/g

84%

5.6 mg/L

35 mg/g

> 95%

52 mg/L

–

99.5%

1.1 mg/L

–

99.5%

1.1 mg/L

Cu

–

–

–

–

100%

2.6 mg/L

Pb

–

–

–

–

88.2%

0.17 mg/L

–

–

–

–

–

67–87%

25 mg/L

–

–

–

0.00475 mg/g 1.225 mg/g 0.0029 mg/g 0.013 mg/g

95%

1 mg/L

245 mg/L

99.65%

100%

100%

80 mg/L

116 mg/L

546 mg/L –

1805 mg/g

4950 mg/g

–

Basic oxygen Furnace Slag (pH = 5.9 → 9.6)

–

6.67 mg/g

100%

10 mg/L

–

100%

169 mg/L

–

97.6%

169 mg/L

Zn

96%

–

–

–

99.3%

0.7 mg/L

–

96.9%

0.7 mg/L

Cd

94 mg/L –

6.67 mg/g

> 95%

10 mg/L

–

98.77%

1.62 mg/L

–

90.12%

1.62 mg/L

Ni

99%

–

–

–

99.74%

1.91 mg/L

–

89.01%

1.9 mg/L

Co

250 mg/L

49.3 mg/g

> 95%

74 mg/L

99.5%

99.91%

–

27.5 mg/L

Al

201 mg/L

Fe

Fly ash pervious geopolymer (pH = 4 → 9.37–9.76)

Power generation fly ash (pH = 3.57 → 6/12.5)

Recycled concrete aggregates (pH = 2.02 → 8.07–11.23)

Carbide lime (pH = 2.6 → ~ 7.5)

Quicklime (CaO) (pH = 3.9 → 10.9)

Calcite (CaCO3 ) (pH = 3.9 → 6.6)

Material

–

–

–

–

–

–

–

–

73%

2903 mg/L

SO4

(continued)

[15]

[11]

[14]

[13]

[10]

[9]

[3]

[3]

References

Table 3.1 Tested alkaline amendments with the initial concentrations of target contaminants (mg/L), percent removal (%), and removal capacity (mg/g)

42 3 In-Situ Remedies

Eggshells activated at 900 °C (pH = 2.5 → 6.9/9.2)

Chicken eggshells mixture (pH = 2.94 → ~ 7)

Hydrotalcite precipitation (pH = 2.9 → 9.2)

Bayer precipitates (pH = 3.74 → 8)

Basic oxygen furnace sludge (pH = 2.5 → 10)

Submerged arc furnace slag (pH = 2 → 9.2)

Magnesium oxide (pH = 7.46 → 12.75)

Material

Table 3.1 (continued)

99.9%

20.5 mg/g

70 mg/L

99.91% –

0.3 mg/g

59 mg/L

92%

–

–

99.98%

22 mg/L

0.81 mg/g

99.9%

49 mg/L

–

–

0.07%

1.5 mg/L

99.8%

100%

72,997 mg/g 20,415 mg/g

61.4 mg/L

219 mg/L

–

76.0 mg/g

98.54%

0.23 mg/L 36.0 mg/g

78.8%

0.14 mg/L

3.5 mg/g

97.9%

0.01 mg/L

534.3 mg/g

99.6%

1.6 mg/L

–

BDLa –

–

99.7%

0.3 mg/L

–

–

–

mg/L

29 > 97.5%

Zn

520 mg/L –

–

98.6%

0.36 mg/L

0.02 mg/g

96.8%

1.54 mg/L

–

0.02 mg/g

90%

0.5 mg/L

> 97%

mg/L

0.00198

Cd

BDLa –

–

99.07%

0.54 mg/L

–

–

< 0.0 mg/g

64.3%

0.03 mg/L

mg/L > 97%

0.0179

mg/L

Ni

> 99.5%

0.0191

Co

> 10 mg/L –

1233 mg/L

99.7%

2.5 mg/g

350 mg/g

16.7 mg/L

99.8%

100%

–

50 mg/L

7000 mg/L

–

–

Al

–

Fe –

236 mg/g

99.2%

0.71 mg/L

–

–

99.99%

40 mg/L

1.3 mg/g

99.9%

77 mg/L

–

0.03 mg/g

88.5%

0.65 mg/L

Cu

–

–

–

–

–

–

–

96.2%

0.13 mg/L

Pb

–

–

–

10.8%

1247 mg/L

–

133 mg/g

70%

4752 mg/L

214 mg/g

66.7%

6436 mg/L

> 41%

mg/L

210.7

SO4

(continued)

[22]

[21]

[8]

[4]

[32]

[17]

[16]

References

3.1 In-Pit Treatments 43

–

–

Cd –

Zn

0.75 mg/g

–

–

< 0.0 mg/g

2 mg/L –

0.14 mg/g

94%

300 mg/L

> 99.9%

–

0.13 mg/g

1.24 mg/g

–

99.9%

99%

99%

0.0 mg/g

600 mg/L

260 mg/L

2500 mg/L

1.7 mg/g

70%

–

0.47 mg/L –

1.65 mg/g

100%

4.11 mg/g

84 mg/L –

19 mg/g

0.02 mg/g

97.2%

20 mg/L

–

0.22 mg/g

69%

160 mg/L

0.78 mg/g

98.73%

26.0 mg/g

–

7.90 mg/L

97.06%

17 mg/L

99.98%

99.99%

99.52%

190 mg/L

260 mg/L

41.3 mg/L

–

–

100%

–

Ni

100%

Co 39 mg/L

Al

310 mg/L

Fe –

0.03 mg/g

> 99.9%

25 mg/L

–

0.01 mg/g

99%

3.3 mg/L

0.78 mg/g

99.36%

7.80 mg/L

Cu –

< 0.0 mg/g

96.3%

0.6 mg/L

–

< 0.0 mg/g

100%

0.04 mg/L

0.61 mg/g

96.83%

6.30 mg/L

Pb –

–

–

3700 mg/L

28 mg/g

93%

61000 mg/L

–

236 mg/g

58.48%

4600 mg/L

SO4

[30]

[29]

[26]

[25]

[23]

References

represents below detection limit The initial and final pH of the tested mine drainages are highlighted with the material Removal capacities were calculated based on the mass of the target removed and the amendment dosage used in each study a. –/–/– refer to initial concentration of the target contaminant (mg/L)/% Removal/Adsorptive capacity (mg/g) b. If adsorptive capacity is bolded it is given in the literature, else it was calculated by total metal removed in treatment divided by total dosage of immobilization material

a BDL

Phosphatic limestone wastes (pH = 3.08 → 8.47)

Low–grade nickel ore (pH = 2.2 → 5.36)

Marl (pH = 2.3 → 8.1)

Cryptocrystalline magnesite (pH = 2 → 10)

Biomass ashes (pH = 3.83 → 5.7–7.8)

Material

Table 3.1 (continued)

44 3 In-Situ Remedies

3.1 In-Pit Treatments

3.1.2.1

45

Zeolites

Zeolites are aluminum- and silicon-containing minerals that are common commercial adsorbents for mine drainage due to their physiochemical properties (e.g., high surface area, porosity) that enable precipitation, electrostatic adsorption, and ionexchange removal. As shown in Table 3.2, a natural zeolite tuff and a fly ash zeolite (i.e., an alkaline material synthesized by hydrothermal activation of fly ash) have shown promise in mine drainage treatment [33, 34]. Specifically, fabricating fly ash into a zeolite increased its adsorption and cation exchange sites for Pb, Zn, Cu, Fe, Cd, and Ni (>80%), while its CaO components enabled quick AMD neutralization. On the other hand, natural zeolite tuff demonstrated synergism when employed with siliceous sand [34]. Zeolite-sand combinations removed Zn, Cu, and Ni to < 0.1 mg/L, where zeolites immobilized Mn, Fe, and Ti, and siliceous sand immobilized Cr, Fe, Zn, and Cu. These materials worked synergistically from their different cation affinities and broadened the range of treatable contaminants. Furthermore, the natural zeolite tuff was abundant in functional moieties capable of generating alkalinity (e.g., CaO, MgO, Al2 O3 ) and, therefore, maintaining circumneutral levels. The synergisms of zeolite and sand blends demonstrates the importance of adsorbent mixtures when treating complex mine drainages and the potential of using cheap, natural materials to derive zeolites.

3.1.2.2

Bio-based Materials

Researchers tested bio-based waste materials for AMD treatment to reduce costs and create a circular economy. For instance, repurposed sludge and manure with high carbonate content and adsorption capabilities (e.g., digested sewage sludge, activated sludge, sheep manure) are effective neutralizers that remove metals by Feprecipitation, divalent metal co-precipitation, divalent precipitation as carbonates (Pb, Zn), and adsorption (Table 3.2) [35–37]. As shown in Fig. 3.2, spent activated sludge could recover 39% of Cu and 100% of U using a 2-step desorption process with Na2 CO3 by exploiting its carbonate moieties. After their recovery, Cu and U were 10-folds more pure than the influent. Future research should investigate the recovery of differently sourced sludges and use a design of experiments to optimize the recovery of a broader range of contaminants under different chemistries. Agricultural organics (Table 3.2), such as peat and wood ash, contain heterogenous surfaces and irregular crevices that render them with large sorption site densities, fast sorption kinetics, recalcitrance to biodegradation, and resiliency to cationic competition [35, 38]. However, agricultural organics like peat and wood ash are more effective at treating CND or NMD since they are incapable of neutralizing or buffering AMD. Nonetheless, wood ash can sequester Ni with its large abundance of calcite moieties, Ca-, Na-, K-oxides, hydroxides, and carbonates [38]. It further exhibited a highly mobile portion of Ni, which may enable metal recovery. Future studies should assess potential sequestration with carbonate species (e.g., Na2 CO3 ) using a design of experiments for optimization, as well as their efficacy in different blends.

46

3 In-Situ Remedies

Fig. 3.2 The regeneration capacity of spent waste activated sludge by benign reagents (H2 O, Na2 CO3 ); Reproduced with permissions from [36]

Biochar, adsorbents synthesized by biomass pyrolysis, are another promising class of adsorbents. Biochars are highly porous, carbon-rich materials comprised of coarse particulates that range in size and contain minerals such as Ca, K, and P [39–42]. They further exhibit a high surface area with alkali salt carbonates (e.g., CaCO3 , MgCO3 ), making them an ideal alkaline adsorbent for AMD (Fig. 3.3a) [43]. Table 3.2 demonstrates the effectiveness of biochar derived from poultry litter, lucerne shoot, vetch shoot, canola shoot, wheat straws, and sugar-gum wood for Cd and Cu removal [43]. Poultry litter was most effective due to its surface functional groups that enabled surface complexations, electrostatic interactions with oxygen functional groups, and precipitation with CO3 . Nonetheless, forest and agricultural-derived biochar suffer from low adsorption capacities and are more susceptible to acidic conditions when compared to conventional amendments. A pivot towards chemical modifications of biomass waste or fishery-derived biomasses has occurred from their susceptibilities to acidic pH levels.

Fig. 3.3 a Removal mechanisms for biochar derived from lignocellulose biomass treating heavy metals, b scanning electron microscopy image of biochar derived from wood residue after KOHsteam treatment; reproduced with permission from [43] and [44], respectively

3.1 In-Pit Treatments

47

Accordingly, porosity modifications (i.e., change in pit, holes, and cavetype openings) using steam activation with CO2 or KOH on wood residues by fast pyrolysis have improved the performance of biochar [44]. KOH-activated biochar performed best, removing Cu concentrations (5–20 mg/L) below guidelines (0.3 mg/L). The high porosity and O-functional groups of the KOH-activated biochar were vital features that enhanced the removal capabilities of biochar. Figure 3.3b illustrates the highly porous and heterogenous surface of the KOH-treated biochar captured through SEM imaging and its performance compared to non-activated biochar. Furthermore, the modified biochar’s regeneration was possible under three cycles using HNO3 (98.6%) and remained relatively high after six cycles (76.2%), suggesting reusability is possible. However, using HNO3 creates secondary pollution, and future research needs to assess benign chemicals or methods for biochar regeneration (e.g., electrochemical regeneration). At the same time, the KOH-activated biochar could not neutralize AMD nor sequester Ni and Co. Using mixed amendments with wood residue biochar or replacing it with alkaline biochars may overcome these drawbacks. For instance, crab-derived biochar is enriched in CaCO3 (i.e., alkalinity) and exhibits a high surface area ideal for metal sequestration [45]. Its CaCO3 component can induce a two-step removal with precipitation followed by adsorption, which has proven effective for Cu and exceeded the adsorption capacities of other conventional adsorbents. Future research should assess their performance using different derived biomasses and their treatment potential with different metal(loid) targets with varying concentrations.

3.1.2.3

Muds and Clays

Muds and clays offer cheap, locally available amendments that have shown promise in mine drainage treatment. For instance, attapulgite, a natural low-cost magnesium aluminum phyllosilicate clay with the formula (MgAl)2 Si4 O10 4H2 O, has been studied as an alternative amendment for AMD [46]. Attapulgite can neutralize and sequester several metals (Co, Cu, Fe, Ni) while exhibiting regeneration capabilities. Red mud has also shown promise in As remediation by creating nucleation sites for Fe3+ precipitation and adsorption [47]. Deposited Fe3+ and its precipitates provided positive functionalities to attract As(III) while enabling co-precipitation of As-iron oxyhydroxides, respectively. Red mud-based geopolymer pervious concrete complexes has also removed metals (Cu, Mn, Cd, Zn), performing better than raw pervious concrete due to the portlandite composition in red mud (Ca(OH)2 ) [48]. Overall, earth minerals and organic materials are suitable adsorbents for AMD. Research should investigate recovery options for valuable materials, where unmodified, modified, virgin, and waste products are all potentially of interest, depending on site requirements and economics. Overall, highly heterogeneous, porous materials that contain CO3 -based functionalities serve as superior adsorbent materials for mine drainage treatment. However, drawbacks include poor pH control, desorption potential, inability to operate under low pH conditions due to H+ competition, and the immobilization of sulfate and certain metals (e.g., Co, Ni) proving to be more

Compost (pH = 7.25)

Waste-activated sludge (pH = 2.0)

Biofert granules (pH = 2.8)

Digested secondary’ sludge (pH = 2.8)

Cattle slurry (Liquid manure) (pH = 2.8)

Power generation fly ash (pH = 3.57 → 6/12.5) (capacity improves with less loading)

Zeolite (Lignite fly ash) (pH = 2.7)

109 mg/g

24.9 mg/g

33%

–

281.4 mg/g

–

–

38.4%

8527 mg/L

100 mg/L

90.8 mg/g

97.2 mg/g

25%

88.3%

270 mg/L

100 mg/L

23.8 mg/g

20.6 mg/g

15.8%

98%

61.7%

270 mg/L

100 mg/L

270 mg/L

100%

94% 0.43 mg/g

–

7.45 mg/L

2.29 mg/L –

–

0%

1952 mg/L

–

–

–

0.3986 mg/g

–

~ 0%

44 mg/L

–

–

–

0.0029 mg/g

100%

99.65%

80 mg/L

116 mg/L

546

mg/L

–

0.03 mg/g

3.24 mg/g

–

–

–

–

–

0.00475 mg/g

95%

1 mg/L

< 0.00 mg/g

–

–

~ 0%

3709 mg/L

–

10.6%

30 mg/L

1.83 mg/g

29.9%

30 mg/L

0.76 mg/g

62.6%

30 mg/L

1.225 mg/g

100%

245 mg/L

0.06 mg/g

100%

1.3 mg/L

22%

0.01 mg/L

80%

–

0.79 mg/L

–

99.94%

–

64.9 mg/L

– –

–

≥ 76.7%

–

Zn

–

Cd

≥ 99.3%

Ni 0.43 mg/L

Co

10.4 mg/L

Zeolite rich-tuff and sand (75:25 vol/vol) (pH = 2.89)

Al

Fe

Material

–

13.2 mg/g

95%

139 mg/L

3.2 mg/g

50.2%

15 mg/L

3.0 mg/g

98.1%

15 mg/L

0.6 mg/g

98.4%

15 mg/L

0.0029 mg/g

100%

0.58 mg/L

0.02 mg/g

99.3%

0.32 mg/L

–

–

< 0.10 mg/L

Cu

–

–

0.5 mg/g

77.7%

2 mg/L

0.7 mg/g

100%

2 mg/L

0.081 mg/g

100%

2.0 mg/L

0.013 mg/g

100%

2.6 mg/L

< 0.0 mg/g

100%

0.01 mg/L

–

–

< 0.10 mg/L

Pb

–

–

–

8800 mg/L

–

–

–

–

0.04 mg/g

–

–

–

34.1%

203.4 mg/L

SO4

Table 3.2 Tested in-pit adsorbents with the initial concentration of contaminants (mg/L), percent removal (%), and removal capacity (mg/g)

(continued)

[38]

[36]

[37]

[37]

[37]

[13]

[33]

[34]

References

48 3 In-Situ Remedies

–

Biochar (Crab by-product) (pH = 2)

–

–

–

–

–

–

–

–

–

84.1%

21.5 mg/L 86% 5.48 mg/g

17.2 mg/g

184.8 mg/g

99%

100 mg/L

11.69 mg/g

99.9%

2.3 mg/L

137 mg/g

99%

1.75 mg/L

26 mg/g

58%

–

–

–

–

–

150 mg/L

Cu

0.24 mg/L

–

4.62 mg/g

> 99%%

0.58 mg/L

–

6%

–

0%

< 0.05 mg/L

4.9 mg/L

0%

0.367 mg/L

0%

13%

95%

141 mg/L

468 mg/L

9.4 mg/L

129.2 mg/g

–

Ca-alginate beads (pH = 3)

–

164.2 mg/g

–

–

–

1.78 mg/L

–

–

88%

–

–

–

1.78 mg/L

Zn

150 mg/L

–

0.70 mg/g

–

Cd

96%

99%

7.45 mg/L

2.29 mg/L –

69% 0.43 mg/g

–

7.45 mg/L

Ni –

2.29 mg/L

Co

500 mg/L

Biochar (Poultry litter) (pH = 3.2)

Biochar (Wood residue, KOH modified) (pH = 2.7)

Mixture: soil, peat, wheat straw (pH = 3.0)

–

HD-peat-calcite (pH = 7.25)

Al

–

Fe

Wood ash (pH = 7.25)

Material

Table 3.2 (continued)

–

–

–

–

–

–

43%

0.14 mg/L

Pb 901

–

–

–

–

901 mg/L

–

mg/L

SO4

(continued)

[50]

[45]

[43]

[43]

[49]

[38]

[38]

References

3.1 In-Pit Treatments 49

–

0.01 mg/g

100%

532 mg/L

Fe

Al

–

–

–

0.00 mg/g

93%

21.2 mg/L

Co

–

0.01 mg/g

95%

0.79 mg/L

Ni –

–

100% –

–

16 mg/L

Zn

100%

1.6 mg/L

Cd

–

100%

10 mg/L

0.01 mg/g

100%

15.65 mg/L

Cu

Pb

–

–

SO4

–

–

[48]

[46]

References

The initial pH of the mine drainage is highlighted in the table a. –/–/– refer to initial concentration of the target contaminant (mg/L)/% Removal/Adsorptive capacity (mg/g) b. If adsorptive capacity is bolded it is given in the literature, else it was calculated by total metal removed in treatment divided by total dosage of immobilization material

Red Mud Pervious Geopolymer (pH = 4)

Attapulgite (pH = 2.84)

Material

Table 3.2 (continued)

50 3 In-Situ Remedies

3.1 In-Pit Treatments

51

difficult than other species. While modifications to some adsorbents may create more effective sulfate removal, an approach using SRB may be a more viable option if sulfate is a target.

3.1.3 Biological In-Pit Treatments Biological in-pit treatments rely on microorganisms, including SRBs, acidophilic heterotrophic bacteria, and algal species, to remove metal(loid)s, sulfate, and acidity in AMD. For instance, SRBs are heterotrophic anaerobic bacteria that use sulfate ions as electron acceptors to drive their metabolism and growth. The generation of bicarbonate ions and alkalinity occur as SRB reduces sulfate to disulfide ions in their metabolic pathways. Bicarbonate ions can induce increases in pH that can stimulate hydroxyl precipitation of metal(loid)s and generate pH levels suitable to increase the number of microbes in situ. Simultaneously, there is a production of disulfide ions that can quickly couple with metal cations to form metal sulfide deposits that precipitate from the solution. Metal sulfide precipitates are preferred since they are denser than their hydroxide counterparts, which reduces sludge volumes and handling costs. However, metal sulfide precipitation is only effective at removing divalent metals Cd, Zn, Pb, and Cu. Metal sulfide precipitation can remove limited amounts of Ni and Co, where pH levels are integral to their solubilities and removals. Nonetheless, biological in-pit treatments with SRB are an attractive strategy when targeting certain pH-insensitive metals (e.g., Zn) and solutions with high metal(loid)s concentrations that do not readily precipitate or adsorb from the solution. Alternatively, algal and acidophilic heterotrophic microbes act as metal scavengers in AMD. Their microbial membranes contain negatively charged moieties, such as carboxyl and phosphorus functional groups, that can electrostatically attract and form complexations with neighboring metals in solution. These electrostatic interactions enable the binding and removal of diverse metals. However, like adsorption, removal is most efficient when treating dilute solutions. Desorption of metal(loid)s in solution will incur once the active sites on microbes membranes saturate from concentrated solutions. On the other hand, many metal(loid)s in mine drainage are trace nutrients and minerals for microbial species. Thus, microbes metabolically uptake trace amounts to drive their metabolic routes and functions. There are currently two methods of stimulating the biological removal of contaminants in environments: biostimulation and bioaugmentation. Bioaugmentation involves inoculating the affected sites with cultured microorganisms to improve the diversity and growth of microbes in situ. The inoculated cultures must be acidophilic microbes that maintain viability and functionality under acidic conditions (pH < 4) and high metal concentrations. Alternatively, biostimulation involves nutrient additions (carbon, nitrogen, phosphorus) in situ to stimulate the growth of indigenous bacteria or algal species. Nutrient additions often reach maximum thresholds and therefore require care when selecting dosages. Table 3.3 summarizes the bioaugmentation and biostimulation studies used to treat mine drainage.

52

3 In-Situ Remedies

Limited studies have tested the effectiveness of bioaugmentation due in part to the high cost associated with its implementation. For instance, microalgae N. oculata was employed for Cu removal in AMD [60]. N. oculata removed 99.9% of Cu with concentrations less than 1010 mg/L, where 88.92% and 10.7% of removal were from metabolism and adsorption, respectively. Metabolic removal by microalgal species offers long-term removal as it prevents the occurrence of desorption and redissolution. Furthermore, increasing the operation time improved lipid accumulation, a value-added product (i.e., biodiesel), with Cu concentrations less than 6.3 mg/L. Likewise, A. manzaenis YN25, a thermoacidophilic archaeon, was studied as a potential specimen for mine drainage bioaugmentation [51]. As demonstrated in Fig. 3.4a, A. Manzaenis YN25 adsorbed Cu (45.1 mmol/kg), Ni (43.2 mmol/kg), Cd (36.6 mmol/kg), and Zn (35.5 mmol/kg) through electrostatic interactions and by forming complexations with metals using its membranal carboxyl and phosphorus functional groups. Ligand exchange dominated metal removal at a wide range of concentrations (0.1–1 mM) and was enhanced at higher temperatures and pH levels [51]. Although demonstrating the principle of bioaugmentation for AMD treatment, further experiments are warranted with industrial drainage to assess the feasibility of bioaugmentation. To test the efficacy of bioaugmentation with pure cultures, D. desulfurican was tested as a specimen and compared against a mixed indigenous microbial community (Fig 3.4b) [52]. Both the bioaugmentation and biostimulation tests added lactate to promote biological growth. Overall, the bioaugmented sample with D. desulfurican exhibited better removal rates and extent of removal for Cu and Zn compared to the mixed indigenous culture. Specifically, the bioaugmented test using D. desulfurican removed Zn in 24 h, superseding the indigenous microbial community that took 360 h to remove Zn [52]. At the same time, D. desulfuricans could remove Cu to certain thresholds unattainable for the indigenous microbial community. To our knowledge, these are the only studies implemented for bioaugmentation with cultivated microbes. Bioaugmentation incurs costs and operational complexities associated with microbial cultivation. Instead, biostimulation is an attractive approach since it can promote microbial growth in situ using waste materials that can create a circular economy. Thus, research has primarily focused on repurposed waste recalcitrant organics to stimulate the growth of indigenous microbes at the mine-impacted sites. For instance, green waste and sewage treatments promoted microbial growth capable of increasing pH levels from 2.4 to 5.5, decreasing sulfate concentrations 7-fold, and removing Fe, Zn, Mn, Cd, and Al [53]. Sewage sludge served as a labile organic supplement and inoculated the affected site with sewage-borne microbes. Simultaneously, the green waste provided a substrate for microbial attachment and, concurrently, biofilm formation [53]. Biofilms are advantageous in AMD treatments since they create interconnected microbial communities resistant to harsh environments and more metabolically active (i.e., capable of faster bicarbonate and disulfide ion generation). On the other hand, sewage sludge provides an opportunity to integrate bioaugmentation with biostimulation without the necessary costs of microbial cultivation. Accordingly, sewage sludge also stimulated microbial growth capable of removing

3.1 In-Pit Treatments

53

Fig. 3.4 a Removal comparisons by bioaugmentation (D. desulfurican) and b biostimulation to treat Zn-contaminated AMD, reproduced with permission from [52]

>94% of Ni and >97% of Zn from AMD in another study [54]. These studies demonstrate the efficacy of using bioaugmentation and biostimulation in conjunction and realize the potential of labile and refractory waste materials to achieve both mechanisms. Thus, although bioaugmentation of pure cultures is practically infeasible, researchers should assess microbe-rich waste materials to achieve biostimulation and bioaugmentation concurrently. The following section highlights the promising amendments found for biostimulation. For instance, peat soil and limestone demonstrated high Fe removal, where Fe removal increased in the bed depth of the peat soil since anaerobic conditions were established for SRB growth [57]. Thus, it may be beneficial to combine labile or recalcitrant organic fractions with peat to stimulate ORP gradients required to treat a multitude of metal(loid)s. A potential labile carbon source is molasses, capable of supporting SRB growth at concentrations of 1–5 g/L [54]. Its stimulation of SRB has shown success in Zn removal as ZnS precipitates, dropping Zn levels from 30 mg/L to 12 months), small surface area to volume ratios, and circumneutral pH. Furthermore, precise stoichiometric fertilizer additions can induce quicker Se removal by 1.5 years as it prevents nutrient sinking. Overall, it is apparent that fertilizer addition in-pit lakes hold their merits but require long treatment times and are costly to supply in the long run. Future studies should use waste amendments rich in nitrogen and phosphorus (e.g., animal manure) to induce primary production of algae. To further improve biostimulation in in-pit lakes, an in-situ passive fixed film bioreactor was designed using polypropylene plastic pieces [66]. Methanol proliferated SRB growth and supported biofilm formation along the polypropylene substrates. The fixed film reactor eliminated acidity and sulfates in the mine drainage, where their removal efficiencies were proportional to the HRT of the bioreactor. The electricity generated by wind and solar energy enabled the reactor to operate passively. Thus, forming a fixed filmed bioreactor in situ is a promising opportunity if operated with a waste organic substrate. Overall, trends within biological in-pit treatments include material selection, passive bioreactor design, microbial selection, resource recovery, and optimal carbon dosing. Organic waste should be sourced nearby mining sites that provide both labile and non-labile fractions, where additional cellulosic waste can further act as a bacterial substrate to promote biofilms. Additionally, bioaugmentation using extremophiles can improve biological treatment but incurs high costs when employed on a large scale and remains largely understudied. Thus, biostimulation with a blend of cellulosic and organic-rich amendments is a suitable option for mine drainage treatment; bioaugmentation can supplement treatment using microbe-rich wastes (e.g., sewage) when this first attempt is subpar in treatment.

56

3 In-Situ Remedies

Fig. 3.5 The generation of phytoplankton blooms in a Se-contaminated pit lake upon nutrient additions, reproduced with permission from [65]

3.1.4 Mixed In-Pit Treatments Mixed in-pit treatments amalgamate biological, immobilization, or alkaline treatments to remediate complex mine waters that require multiple removal mechanisms concurrently (Table 3.4). Immobilization techniques can concentrate and localize metals for their biological reduction and biosorption. Localization by alkaline adsorbents optimizes performance since it ensures suitable pH conditions for SRB (pH ~ 5–8) growth and metabolism. For instance, alkaline zeolites were applied with complex organic waste to treat weathered tailings [67]. The combined amendments synergistically augmented metal sulfide and hydroxide precipitation and adsorptive removal. Compared to the control (pH 3.2–3.6), the mixed amendments increased the pH to circumneutral, providing favorable conditions for SRB growth, hydroxide precipitation, and adsorptive removal that enabled the removal of Cd, Cu, Ni, and Zn to exceed 88%. Zeolites have also been combined with market and food waste compost to treat complex mine drainage [68]. The zeolites and lime in the market compost maintained circumneutral pH levels ideal for biological activity and reduced Cr leaching potential. Similar results occurred with crushed limestone, spent mushroom compost, activated sludge, and woodchips treating AMD [59]. The spent mushroom substrate adsorbed metals and provided a carbon source to support SRB growth and bacterial sulfate reduction. Alternatively, Maifan stones, an abundant adsorbent, have immobilized SRB to remediate complex AMD [69]. The amendments combined were synergistically superior at remediating Fe and sulfate and showed neutralization abilities. The Maifan stones adsorbed and concentrated metals nearby the SRB and provided bacterial attachment sites that improved SRB activity to remove metals and sulfates. These studies demonstrate the potential benefits of combining adsorbents with organics and alkaline materials to enhance the biological removal and precipitation of target metal(loid)s to meet regulatory standards.

Sugarcane Bagasse (pH = 2.3)

Molasses (pH = 3.3)

Sewage sludge (pH = 2.8)

Sewage and green waste (pH = 2.4)

Lactate (pH = 5.8)

N. oculata (pH− = 8.5)

D. desulfuricans (pH = 5.8)

A manzaensis YN25 (pH = 4.0)

Material

100% 0.83 mg/g

2.33 mg/g

9.21 mg/L 74%

150.71 mg/L

94.4%

–

–

–

–

–

–

–

0.01 mg/L –

0.02 mg/g

90.9%

100 mg/L

0.08 mg/g

–

–

–

–

–

–

–

12.6 mg/L

–

–

14 mg/L –

102 mg/L

99.89%

3.2 mg/g

11.03 mg/g

380 mg/L

99.99%

99.89%

95.4%

5.2 mg/L

200 mg/L

690 mg/L –

–

– 2.6 mg/L

–

–

260.8 mg/L

–

–

–

–

–

–

–

2270 mg/L

1.102 mg/L

1.47 mg/L

–

–

12.6 mg/L

– 4.11 mg/g

– 2.54 mg/g

–

–

–

Cd

Zn

Cu

–

81.3%

1.18 mg/L

6.0 mg/g

~ 100%

31 mg/L

0.02 mg/g

96%

20 mg/L

0.25 mg/g

98.6%

16 mg/L

9.45 mg/g

100%

18.9 mg/L

–

–

100%

18.9 mg/L

2.32 mg/g

–

–

–

–

66.67%

0.3 mg/L

Pb

–

~ 100%

0.86 mg/L

–

–

0.089 mg/L

0.21 mg/g

100%

25 mg/L

–

98.7%

4.58 mg/L

–

–

0.087 mg/L

–

< 0.00 mg/g < 0.00 mg/g

96.5%

0.3 mg/L

0 mg/g

0%

9.9 mg/L

–

99.92%

15.9 mg/L

–

100%

9.9 mg/L

2.87 mg/g

–

5.9–59 mg/L 11–112 mg/L 6.5–65 mg/L 6.3–63 mg/L

Ni

260.8 mg/L –

–

Co

–

–

Al

2270 mg/L

Fe –

–

11.25 mg/g

75%

3000 mg/L

36.02 mg/g

86.6%

2600 mg/L

1739 mg/g

47%

7400 mg/L

–

–

47%

7400 mg/L

SO4

(continued)

[55]

[54]

[54]

[53]

[52]

[60]

[52]

[51]

References

Table 3.3 Tested in-pit biological amendments with the initial concentration of target contaminants (mg/L), percent removal (%), and removal capacity (mg/g)

3.1 In-Pit Treatments 57

–

–

Cd –

–

–

7.5 mg/g

– –

–

6.70 mg/L 0.87–1.8%

26–14%

0.90 mg/L –

2.5–20%

0.058 mg/L

–

–

19–64%

13.4 mg/L

2.64 mg/g

100%

–

0.88 mg/L –

0.67 mg/g

96.9–99.5%

65.45%

0.02 mg/g

–

171.57 mg/L

Zn

3.82 mg/L

14.25 mg/g

4.48 mg/L –

–

Ni

97.3–99.5%

–

–

Co

99.8%

Al

3570 mg/L

–

95%

75 mg/L

Fe –

–

11.3 mg/g

99.95%

3.77 mg/L

–

–

12.13 mg/L

Cu –

–

5.02 mg/g

98.99%

1.69 mg/L

0.03 mg/g

93.2–95.9%

7.49 mg/L

Pb –

–

2700 mg/g

56.25%

1600 mg/L

42.72 mg/g

83.7–89.2%

12,760 mg/L

SO4

[63]

[59]

[58]

[57]

References

The initial pH conditions of the mine drainage are highlighted a. –/–/– refer to initial concentration of the target contaminant (mg/L)/% Removal/Adsorptive capacity (mg/g) b. If adsorptive capacity is bolded it is given in the literature, else it was calculated by total metal removed in treatment divided by total dosage of immobilization material

Nitrogen and phosphorus (pH = 7.56)

Mix (limestone, spent mushroom compost, activated sludge, woodchips) (pH = 2.8)

Sheep manure & compost (pH = 2.4)

Peat soil (pH = 3.5)

Material

Table 3.3 (continued)

58 3 In-Situ Remedies

–

–

–

–

100%

–

–

17–43%

87.6 mg/L

–

–

96–97%

–

–

44–91%

3.1 mg/L

Zn

Cu

–

–

100%

–

–

–

–

22–73%

–

–

–

61.6 mg/L

Pb

–

–

–

6000 mg/L

SO4

30 mg/L

> 99.3%

2.979 mg/g

200 mg/L

> 98%

19.6 mg/g

0.204 mg/g

0.0978 mg/g

97.8%

1 mg/L 0.2 mg/g

100%

mg/g

2 mg/L

1 mg/L 86.7 ± 10.5%

10 mg/L 0.904 mg/g

90.4 ± 9.1%

2 mg/L 0.189 mg/g

94.5 ± 2.6%

–

95.6 mg/g

95.6%

1000 mg/L

–

93.61%

–

–

88–100%

–

–

Cd

834.5 mg/L –

–

–

Ni

91%

Co

14–50 mg/L

–

–

93–95%

–

–

–

0.5–24%

80–82%

152 mg/L

34–44%

Al

500 mg/L

Fe

[70]

[69]

[68]

[67]

References

The initial pH conditions of the tested mine drainage are highlighted a. –/–/– refer to initial concentration of the target contaminant (mg/L)/% Removal/Adsorptive capacity (mg/g) b. If adsorptive capacity is bolded it is given in the literature, else it was calculated by total metal removed in treatment divided by total dosage of immobilization material

Crab shell

Maifan stones and SRB (pH = 4.1)

Food waste and zeolite (pH = 3.4)

Compost and zeolite (pH = 3.0)

Material

Table 3.4 Tested mixed amendments with initial target contaminant(s)’ concentration, percent removal, and removal capacity (mg/g)

3.1 In-Pit Treatments 59

60

3 In-Situ Remedies

3.2 Saturated Rock Fills 3.2.1 Saturated Rock Fills (SRFs) Technology Description Saturated rock fills (SRFs) are in-situ bioreactors that capture water seepage by backfilling mining pits with waste rock (see Fig. 3.6) [71]. The backfilled mining pits capture rain and snow as time progresses, creating saturation conditions within the waste rock pores. Saturation conditions mitigate oxygen diffusion into the rock pores that assist in maintaining anoxic conditions. Simultaneously, oxygen consumption occurs by aerobic microbial respiration supported by carbon or nutrient additions. Carbon sources support microbial growth by serving as electron donors and, in turn, oxidize to bicarbonate and dissolved CO2 . Alternatively, nutrient sources, such as nitrogen and phosphorus, allow for the primary production of algae that can act as carbon sources for microbial growth. These induced algal blooms may also serve as metal scavengers in mining pits, but caution is needed to prevent eutrophication. Since the present mine sediments are the SRF inoculum, the carbon and nutrient amendments proliferate microbes pre-acclimatized to target contaminants (e.g., Se, NO3 − ). Overall, the present microbial species and their diversity determine the success of SRFs. Beneficial microbial species include anaerobic SeRB, such as obligate Se6+ and Se4+ reducers and SRB that utilize sulfate and Se as terminal electron acceptors. SRFs remove target contaminants by hosting a diverse microbial community that uses contaminants (e.g., Se) as electron acceptors to fuel their anaerobic microbial growth. Thus, the redox potentials of contaminants determine their degree of removal since each yield different energy gains for microbes. For instance, since nitrate yields the most energy for microbes, it is preferentially removed by microbes over other contaminants. Nitrate removal commences with its biological reduction to N2 (g), which subsequently releases into the atmosphere. Once nitrate concentrations deplete, anaerobic microbes switch to using Se as their terminal electron acceptor. Se is then biologically reduced to selenite or elemental selenium and removed from the solution by co-precipitation with reduced sulfates [73]. SRB stimulation by carbon or nutrient addition can also remove other metals by metal sulfide precipitation once Se and nitrate levels diminish. Although beneficial for metal precipitation, H2 S, a corrosive and toxic gas, may generate in the waters from SRB metabolism. Thus, once SRF treatments are complete, the treated effluent enters an aeration polishing step to remove residual H2 S produced from SRBs’ metabolisms. SRFs expedite the metabolic uptake of target contaminants since pit lakes are characteristically deeper than natural lakes, more saline, and less susceptible to wind mixing. These characteristics create meromictic conditions and suboxic zones that enable Se and sulfate precipitation. The successful implementation of SRFs requires the following considerations: 1. Mining pit should be connected hydraulically to downstream flows [71]. 2. SRFs must operate under long HRTs that are resilient under changing temperatures and flow conditions [71].

3.2 Saturated Rock Fills

61

Fig. 3.6 Schematic of saturated rock fills and its mechanism to microbially reduce Se in its saturated pores. Adapted with permissions from [72]

3. Waste rock must be placed strategically for hydraulic control and seepage capture within the SRF [71]. 4. The rate at which suboxic conditions develop determines the time required for target removal and, therefore, HRT [71]. 5. SRFs require an understanding of process stoichiometry, kinetics, and performance in response to carbon source types and dosages [71]. 6. SRF hydraulics depend material porosity, which invokes uncertainty in flow estimates as porosity can vary. These uncertainties must be considered within its design [74]. 7. Competing ions (e.g., nitrates, Se, oxygen) suppress sulfate reduction. 8. Oxidative-reductive potential (ORP) gradients are required to remove Se and nitrates simultaneously. SRFs offer several advantages as a passive treatment strategy for mine-impacted sites. First, SRFs lower treatment costs since indigenous waste rock and mine pit sediment are used as substrate and inoculum, respectively [71, 75]. Second, SRFs mitigate sulfide mineral oxidation. Specifically, backfilling mine pits with waste rock mitigates sulfide mineral oxidation since oxygen levels within the saturated pore spaces deplete from diffusion-controlled mass transport [74]. At the same time, indigenous microbes mitigate sulfide mineral oxidation by forming biofilms that act as secondary diffusion barriers for oxidants and consume oxidants, oxygen and Fe3+ , through their respiration [76]. Lastly, the suboxic zones within SRF-saturated waste rock pores create localized ORP gradients. The ORP gradients enable the simultaneous removal of nitrates, Se, and sulfates. The co-removal of Se and sulfates broadens Se removal mechanisms to include adsorption, precipitation of Se as an inorganic or organic selenide, and co-precipitation of Se with reduced sulfur. Figure 3.4a illustrates an overall schematic of SRFs, and Table 3.5 summarizes SRF’s effective targets, removal mechanisms, and highlights.

62

3 In-Situ Remedies

3.2.2 Saturated Rock Fill Studies Initially, small-scale, saturated bioreactors were studied to assess SRF’s feasibility. Tanks filled with rock particles collected from a mine waste near Leadville, CO, simulated one SRF design [77]. The constructed bioreactors treated AMD contaminated with Zn, Cd, and Cu and were operated under both unsaturated and saturated conditions. Glucose was the electron donor which induced a microbial shift from iron-oxidizing bacteria to heterotrophic iron- and sulfur-reducing microbes. The resulting microbes’ respiration caused pH levels to elevate to circumneutral levels under both flow conditions. However, saturated conditions reduced carbon amendment requirements, attained greater pH levels, and larger hydraulic conductivities. This performance enhancement was due to the proliferation of anaerobic bacteria from oxygen-limited conditions in the saturated pores which can produce alkalinity and remove metals. The anaerobic bacteria create higher hydraulic conductivities due to their smaller cell yields and thinner biofilms than that seen for aerobic microbes. The saturated conditions also created suitable growth conditions for ironand sulfur-reducing bacteria, enabling more than 85% of Zn, Cd, and Cu to be removed over 300 days. Subsequently, wooden boxes were used as bioreactors to treat Se-contaminated waters [75]. The bioreactors operated under an HRT of 2 days and utilized an active substrate mixture comprised of mulch, manure, limestone, and bone-meal to sustain microbial growth. The resulting design removed more than 95% of Se, demonstrating resiliency to seasonal temperature changes (~ 2 to 14 °C) and to excess sulfate concentrations. Unexpectedly, there was the removal of nitrates and nitrites with Se. Generally, nitrates and nitrites compete as terminal electron acceptors in anaerobes’ respiration but concurrently diminished with Se in this study from an ORP gradient created in the saturated waste rock. This study concluded that SRFs should use end pit lakes as in-situ bioreactors since they contain suitable inoculum for metal removal. Additionally, the authors recommend liquid fertilizer or cheap carbon sources (e.g., old hay or leaf litter) to stimulate microbial growth. Following these preliminary studies, Teck conducted up-flow column tests as a proof-of-concept for their SRF design [78, 79]. The columns were packed with waste rock and operated with up-flow hydraulics under aerobic, microaerobic, and suboxic conditions [79]. The efficacy of glycerol and methanol were tested and compared to a control without carbon source. Se and nitrate removal increased with decreasing dissolved oxygen levels, with nitrate removal superseding Se removal. Both contaminants’ removal required carbon amendments; reactors without carbon sources experienced sporadic trends in their contaminants’ concentration unless there were suboxic levels with a 16-h HRT. Lastly, glycerol was the preferred carbon substrate under low oxygen conditions since it proliferated SRB and SeRB and induced the highest removal of both contaminants. However, glycerol is readily biodegradable and incurs high operating costs with its dosage requirements. Moreover, a small-scale test using plastic barrels backfilled with waste rock assessed the efficacy of different carbon amendments and their dosages [71]. The reactors operated under a two-week HRT to treat coal mine water from southern

3.2 Saturated Rock Fills

63

Alberta, contaminated with Se and nitrate. Methanol, molasses, and their combination were tested under different dosages and compared with a control without a carbon substrate. A 50:50 mixture of methanol and molasses enabled more than 98.8% of Se removal. Se removal required carbon amendments in which their dosage requirements were stoichiometrically proportional to the oxygen and nitrate levels in the water (e.g., 3 g methanol per 1 g of NO3 − ). Utilizing dosages that surpassed their stoichiometric requirement showed no additional benefits for both amendments, and Se removal was independent of the carbon amendment type. The performance assessed with small-scale reactors translated well to the efficacy of SRFs when applied in situ at small-scale mine-impacted sites. For instance, an SRF treated Se in a small mine pit within the Rocky Mountain Foothill region (British Colombia, Canada) [74]. The SRF was operated with an HRT of 0.5 to 3 years in the winter and 0.3 to 2 years in the spring and summer months. The HRT was controlled by balancing rainfall infiltration with groundwater exits over bedrock ridges and downward flows through underlying bedrock. Se levels remained < 1 µg/L within the saturated pores, owing to its adsorption, precipitation as elemental Se and precipitation of metal or organic selenides. There was also a shift of Se speciation from Se6+ to Se4+ . However, there was no indication of sulfate reduction. Similar results occurred when treated Se in a coal mine in Northern Canada, as Se levels remained < 3 µg/L after SRF treatments [80]. Furthermore, carbon amendments are not required when treated mine sites with organic-rich waste rocks under saturated conditions, as indicated by a study treating phosphate mine waste [81]. The waste rocks were strategically placed in short, compacted lifts to ease water retention and reduce water flux rates. The saturation conditions preferentially proliferated native SeRB, as there were 300 times more SeRB in the saturated waste rock. Due to this microbial shift, oxygen levels depleted, and nitrate, Fe, and Mn concentrations simultaneously reduced with Se. However, due to competing ions, there was no sulfate reduction. Due to its success in small-scale studies, Teck installed two SRFs to treat the Elk Valley mines [82]. The mining sites were backfilled with water to establish saturation conditions and were shown capable of treating 20 million liters of water per day and removing > 95% of Se. The removed Se was shown to be successfully reduced to an immobile, solid state and was securely stored within the mine pit’s retention time. In addition, there was near-complete nitrate removal from their mine waters. In summary, SRFs can remove Se and nitrate to meet regulations in large-scale operations. However, only one proof-of-concept study has assessed its feasibility of removing divalent metals and its ability to generate alkalinity. In addition, there have been no reports indicating the degree to which SRF can suppress sulfide mineral oxidation. Research should assess whether SRFs can treat a broader range of contaminants, generate alkalinity and mitigate sulfide mineral oxidation. Additionally, research should investigate the efficacy of different carbon and nutrient sources simultaneously, as it remains unclear which substrates are best suited for SRF technology. Finally, the resilience of reduced Se deposits is unknown in SRFs. The stability of Se deposits is required to understand long-term implications and to uncover the risk of washout events as time evolves.

Materials

Neutralizing substances (Portland cement, phosphatic limestone wastes, waste slags, recycled concrete aggregates, calcite, etc.)

Type

Alkaline Zn, Ni, Co, Cu, Pb

Target Metals

References

• Alkaline materials increase pH and generates [2, 3, 8–10, 15, 21, 26, 29] metal hydroxide precipitates • Concrete materials (e.g., fly ash) release Cr, Cu and elevates pH > 9 • Alkaline materials produce toxic sludge (i.e., secondary pollution) • Calcite materials have long-term buffering capabilities (> 107 days) • Ore grade hydrotalcite can form in mining water containing Mg and Al using a M2+ :M3+ ratio of 2.5:1 • Surface modifications with SDS and Tween 80 improves dissolution • Passivation of alkaline amendments can act as source control • Calcination can improve pore structure of materials • Selective precipitation, double-layered hydroxides, and hydrotalcite formation useful for resource recovery • There are opportunities to turn waste and hazardous wastes into treatment materials or value-added products

Mechanisms/Highlights

Table 3.5 A summary of in-pit treatments, their associated materials, target metals, and mechanisms/highlights

(continued)

64 3 In-Situ Remedies

Target Metals Cu, Cd, Ni, Zn, Co, Pb

Materials

Adsorbents (Zeolites, sewage sludge, activated biochar, etc.)

Type

Immobilization

Table 3.5 (continued) Mechanisms/Highlights

References

(continued)

• Metals are removed by ion-exchange, [17, 25, 28, 33, 34, 37, 43, 44, 46] surface complexation and precipitation • Suffers from desorption and poor pH control • KOH increases the number of oxygenated groups and therefore the sorption capacity of biochar • Alkaline clay and coal refuse field study shows most promise: circumneutral pH > 7 yrs and supports vegetation for enhanced metal immobilization • Not suitable for mine drainage with high acidity unless adsorbent has neutralization properties • Animal based biochar has calcite moieties which counteracts H+ competition at low pH • Granular size and activation temperature can be modified to improve activated carbon • Research focus on selectivity of desired (valuable or toxic) metals for recovery • Na2 CO3 shown capable of regenerating spent biosorbents

3.2 Saturated Rock Fills 65

Mixed Cu, Cd, Ni, Zn

Adsorptive Material (Zeolites, Maifan stones) Recalcitrant Material (Food waste, compost)

Bioremediation

Target Metals

Materials

Slowly biodegradable organics Cu, Zn, Cd, Pb, Se (Sewage, green waste, molasses, manure, etc.) N/P source (Fertilizers for algal growth)

Type

Table 3.5 (continued) Mechanisms/Highlights

References

• Metals removed via adsorption and metal sulfide bioprecipitation • Prevents metal leaching from adsorbents • Shown promise in maintain circumneutral pH with high metal removal (> 88%) • Immobilization of SRB on adsorbent materials effective

[67, 68]

(continued)

• SRB induce metal sulfide bioprecipitation [1, 26, 51, 53, 54, 56, 59, 63, 65, 70] and generate alkalinity • Algal species can metabolize, assimilate, and adsorb metals • Algal removal is less impactful than SRB removal apart from metals insensitive to SRB bioreduction (e.g., As, Co, Ni) • There are successful pilot studies (> 172 days) with SRB growth • Green waste enhances removal via promoting biofilm formation • Mitigates secondary sludge volume • Bioaugmentation shown to improve rates and capacity compared to biostimulation but is costly • Biostimulation eases treatment processes and costs • Low temperature and salinity likely to inhibit treatment • Research focuses on effective recalcitrant organic carbon sources • Cellulose shown to be promising as a long-term C source (> 4 wks)

66 3 In-Situ Remedies

Materials

Organics (molasses, methanol, glycerol, glucose) Substrate (Waste rock)

Type

Saturated Rock Fills (SRF)

Table 3.5 (continued) References

Mechanisms/Highlights • Mined out pit lakes are backfilled with waste [71, 74, 77–79, 81] rock and are saturated to reduce O2 exposure • Backfilled waste rock requires hydraulic control • Microbial activity within indigenous sediment is stimulated via organic and nutrient amendments • Algae production via nutrient addition acts as a carbon source for microbial growth and acts as a metal scavenger • Biofilms create protective microenvironments for microbes • Suboxic zones allow for dissimilatory metal reduction • SRB growth generates alkalinity • Requires long residence times • Metal precipitation onto waste rock introduces uncertainty in residence time • Redox gradients allow for simultaneous removal of nitrates and Se

Target Metals Se4+ , Zn, Cu, Cd

3.2 Saturated Rock Fills 67

68

3 In-Situ Remedies

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71. S. Jensen, J. Foster, M.-C. Noel, M. Bartlett, Mine design for in-situ control of selenium and nitrate (2018) 72. L.B. Kirk, C. Hwang, C. Ertuna et al., Column tests of selenium biomineraliation in support of saturated rockfill design. in Mine Water and Circular Economy (vol. II) (2017), pp 1191–1195 73. S.G. Deen, V.F. Bondici, J. Essilfie-Dughan et al., Biotic and abiotic sequestration of selenium in anoxic coal waste rock. Mine Water. Environ. 37, 825–838 (2018). https://doi.org/10.1007/ s10230-018-0546-9 74. M. Bianchin, A. Martin, J. Adams, In-situ immobilization of selenium within the saturated zones of backfilled pits at coal-mine operations 75. A. Luek, C. Brock, D.J. Rowan, J.B. Rasmussen, A simplified anaerobic bioreactor for the treatment of selenium-laden discharges from non-acidic, end-pit lakes. Mine Water Environ. 33, 295–306 (2014). https://doi.org/10.1007/s10230-014-0296-2 76. S. Jin, P.H. Fallgren, J.M. Morris, J.S. Cooper, Source treatment of acid mine drainage at a backfilled coal mine using remote sensing and biogeochemistry. Water Air Soil Pollut. 188, 205–212 (2008). https://doi.org/10.1007/s11270-007-9536-4 77. J.D. Jenkins, Role of flow and organic carbon on acid mine drainage remediation in waste rock (2001) 78. C. Hwang, L.B. Kirk, B. Peyton, Changes in microbial community structure in response to changing oxygen stress in column tests of denitrification and selenium reduction (2017) 79. L.B. Kirk, C. Hwang, C. Ertuna et al., Column tests of selenium biomineralization in support of saturated rockfill design (2017) 80. A. Martin, R. Goldblatt, J. Stockwell, In situ attenuation of selenium in coal mine flooded pits (2015) 81. L.M.B. Kirk, In Situ Microbial Reduction of Selenate in Backfilled Phosphate Mine Waste (Montana State University, S.E. IDAHO, 2014) 82. MARIAAN WEBB, Teck doubles water treatment capacity at Elkview, in Mining Weekly (2021). https://www.miningweekly.com/article/teck-doubles-water-treatment-capacity-at-elk view-2021-02-17/rep_id:3650. Accessed 29 Nov 2022

Chapter 4

Ex-Situ Remedies

Abstract Ex-situ remedies refer to treatments that use natural hydraulic gradients or pump-and-treat methods to remediate mine drainage or developed plumes offsite; these include anoxic limestone drains, permeable reactive barriers, constructed wetlands, and gravel bed reactors. The following chapter delves into each of their technologies in terms of their removal principles, research trends, and identified literary gaps. Keywords ALD · Permeable reactive barrier · Passive · AMD · Gravel bed reactor · Constructed wetland

Ex-situ treatments refer to methods not conducted on-site, such as technologies that intercept plumes (e.g., anoxic lime drains, permeable reactive barriers, constructed wetlands) and semi-passive pump-and-treat methods (e.g., gravel bed reactors). The following chapter describes each technologies components, removal mechanisms, efficacies, and research gaps.

4.1 Anoxic Limestone Drains (ALD) Anoxic limestone drains (ALD) encompass a bed of crushed limestone (CaCO3 ) buried underneath a soil or clay cover with plastic liners. The crushed limestone increases the alkalinity as AMD flows through while the clay or soil covers devoid oxygen permeation (Fig. 4.1) [1, 2]. After passing through the ALD, the alkalinetreated AMD enters a pond or wetland. The atmospheric oxygen in the ponds or wetlands oxidates and precipitates Fe3+ as Fe(OH)3 and Al3+ as Al(OH)3 . The precipitation of Fe and Al as hydroxides removes these trivalent metals from the solution and enables the co-precipitation of neighboring divalent metals [3]. ALDs negate premature oxygen exposure since Fe(OH)3 and Al(OH)3 precipitation occurs upon oxygen contact, which in turn, creates depositions on limestone known as armoring (Fig. 4.1) [4]. Armoring reduces limestone dissolution rates and the overall efficiency © The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 C. Chidiac et al., Passive Treatments for Mine Drainage, SpringerBriefs in Applied Sciences and Technology, https://doi.org/10.1007/978-3-031-32049-1_4

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Fig. 4.1 An anoxic limestone drain schematic with an SEM image showing limestone armoring. Adapted with permissions from [2, 4]

of the ALD. To mitigate armoring, ALDs should operate with < 1 mg/L of dissolved oxygen and 1–20 mg/L of Fe and Al [4]. For ALDs to be efficient, they further require high-purity limestone (> 90% CaCO3 ) with high hydraulic conductivity, porosity, surface area, and a long detention time with flushing capabilities to remove sludge or armoring. ALD sizing must account for limestone dissolution over time, limestone armoring, and potential environmental impacts [3]. Design recommendations for ALDs include a twenty-year system life; an implementation of a two-foot clay or clayey soil cover to mitigate oxygen and protect ALD covers from erosion or root growth; perforated pipes to remove excess sludge; and to use a 6-inch safety factor of limestone to account for embedment from excavation [5]. To improve ALD system life, oxygen, Al, and Fe can be partially removed by flowing the AMD through a stage preceding the ALD comprised of compost. Additionally, parallel ALDs can be implemented under horizontal and vertical flow regimes to improve the effective surface area and maintenance adaptability. However, even with treatment optimized, ALDs are only partially effective at removing sulfate, Fe, Al, and Mn [3]. Thus, ALDs are often implemented as pre-treatments to create suitable pH conditions for constructed aerobic wetlands [5, 6].

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Alternatively, oxic limestone drains (OLD) control the oxidation of trivalent and divalent metals to form oxide precipitates capable of adsorbing trace metals and sulfates [1]. Although OLD broadens the removal mechanisms available, it also requires low metal concentrations to prevent drain clogging. Since AMD contains relatively high metal concentrations, this technology is ineffective as a stand-alone removal technique. For these reasons, OLDs and ALDs are absent from the subsequent recommendations section (Chap. 5).

4.2 Permeable Reactive Barrier (PRB) Permeable reactive barriers (PRBs) use natural hydraulic gradients to drive contaminated plumes toward impermeable barriers installed perpendicular to their stream. As demonstrated by Fig. 4.2a, the contaminated plume flows toward the “capture zone” containing reactive material, which either removes (i.e., chemically or biologically transformed) or retains the target contaminant. The reactive material requires a higher hydraulic conductivity (≥ 10x) than the surrounding material to ensure the plume flows through the capture zone. While installing an impermeable lower bound prevents lower-ground influxes [7]. Figure 4.2b, c illustrate the impermeable barrier designs for PRBs: continuous wall or trench-based PRBs and funnel-and-gate PRBs, respectively. Funnel-and-gate PRBs use impermeable, angled walls to redirect the flow toward the capture zone for

Fig. 4.2 a The interception of a contaminated plume by a permeable reactive barrier (PRB); the configuration of a b continuous PRB and c a funnel-and-gate PRB. Images reproduced with permission from [8]

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treatment. The reactive material is placed in the capture zone using a steel caisson to reduce material dosages and is preferred when the reactive material is expensive. Although using steel caissons and steel, angled walls minimizes the amount of reactive material required, it incurs high construction costs and adverse environmental impacts that exceed that seen for continuous wall PRBs. Thus, funnel-and-gate PRB designs must mitigate the lengths of the funnel walls to reduce their environmental impact and cost [9]. The second PRB design is the continuous wall or trench-based PRBs. This design is most common and involves the installation of a perpendicular, subparallel trench downstream from the contaminated plume [8]. Its implementation requires a clear understanding of the hydrology of surrounding groundwaters and large areas not spatially impeded by existing construction structures [7]. Additionally, trench-based PRBs necessitate more reactive media than funnel-and-gate designs and thus use cheaper materials within their capture zone (e.g., organics, recycled- or wastematerials). Since a trench filled with reactive material replaces angled walls, trenchbased PRB designs negate steel use for the caissons and funnel walls. By eliminating steel requirements, the environmental impact associated with steel manufacturing and transport is reduced [9]. Additionally, less equipment is required for continuous wall PRB installments, further reducing its environmental impact and creating a more sustainable design [10]. Nonetheless, a life cycle assessment can objectively compare the sustainability of the two PRB designs.

4.2.1 PRB Design: Associated Materials and Removal Mechanisms PRBs designs invoke removal mechanisms necessary to remove the target contaminants within a given plume while remaining economically and environmentally advantageous. The removal mechanisms of PRBs depend on (1) influent conditions (Eh, composition, pH), (2) reactive substrate material, and (3) HRT [11]. Thus, the main components of PRBs designed to target specific contaminant(s) are the reactive material and its hydraulic conductivity. When selecting reactive material, one must consider the material(s) porosity to sustain flow, lifespan, environmental friendliness, cost, and availability from local sources. Recent research has been focused on recycled or waste materials within PRBs, as these materials can reduce their carbon footprint while achieving the same treatment targets. Combined media, implemented as a blend or as stratified layers of different reactive materials, should be considered when dealing with complex plumes to invoke better removal rates and to provide lifetime extensions [8]. The hydraulic conductivity of the material will influence the contact time the contaminant has within the given material (i.e., HRT). Long HRTs are required to sustain microbial sulfate reduction and metal sulfide precipitation [11]. Short HRTs (i.e., a few minutes to hours) are implemented to promote the sorption or

4.2 Permeable Reactive Barrier (PRB)

77

precipitation of oxidized species and are more economical for lower metal and sulfate concentrations [11]. Therefore, the HRT of the PRB design should be chosen based on the mine drainage composition. Determining factors governing the HRT of the PRB design are the porosity and total volume of the reactive material(s). Factors influencing the flow capacity of PRBs and HRT are the characteristics of the influent water and the metal removal rates of reactive barriers. Although variable, the quickest flow rate reported for PRBs is one pore volume per week (i.e., seven-day residence time) [12]. There is a potential for higher flow rates, as modeling simulations have demonstrated flow rates as high as 120 L/s are possible with a 12 m thick PRB and 7-day residence time [12]. However, further experimental demonstrations are required to confirm whether this flow is possible. The following sections outline the different classifications of reactive materials used in PRBs and their advantages and disadvantages. Table 4.1 summarizes the highlights and mechanisms for each classification and their effective targets.

4.2.2 Reactive PRBs Reactive PRBs use materials that undergo reactions with the contaminant for their removal or containment. Zero-valent iron (ZVI) is the most common material implemented since it has a negative reduction potential (–0.44 V) and oxidizes quickly. For instance, when exposed to oxygen, ZVI undergoes oxidation in the following manner, 2Fe0 + O2 + 2H2 O → 2Fe2+ + 4OH−

(4.1)

thereby generating electrons and alkalinity as hydroxide ions. ZVI is a strong reductant and can donate its electrons to divalent dissolved metals with more positive reduction potentials. As ZVI donates electrons to divalent metals, their valency reduces; reducing the valency of divalent metals decreases their mobility in mine drainage and, therefore, their toxicity. Similarly, ZVI oxidizes with sulfate under anaerobic conditions in the following manner, + − 2+ Fe0 + SO2− + 4H2 O 4 + 9H → HS + 4Fe

(4.2)

further consuming H+ and producing corrosion products. The corrosion product, HS– , can remove heavy metals by complexation and sorption mechanisms [45]. However, corrosion products and precipitate deposition reduce the porosity and hydraulic conductivity of ZVI. ZVI corrosion also expands its structure and initiates gas formation detrimental to its porosity and hydraulic conductivity. These phenomena collectively reduce the permeability of the reactive material and cause short-circuiting issues that reduce contact times between the reactive material and the contaminant.

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Table 4.1 A summary of the classifications of permeable reactive barriers with their reactive materials, target metals, and their associated removal mechanisms and highlights Type

Materials

Targets

Mechanisms/highlights

Diffusion-active

Reactive Layer: recalcitrant organics, ZVI, limestone, anaerobic sludge Conductive Layer: coarse sand

Cu, Zn

• Intercalating reactive [13–18] and conductive layers • Reactive layer harbors and isolates microbes from mine drainage in the conductive layer • Improves microbial diversity and metal sulfide removal • Precipitation at interlayers’ interface eliminates clogging • Principles extend to alkaline-reactive materials

Biological

Recalcitrant Organics (e.g., algae, cellulose, sludge, coconut husks) ZVI (e− donor) Packed material (limestone, pea gravel)

Cu, Zn, Cd, Pb, Se, As

• Organics commonly mixed with ZVI and limestone • SRB growth for MeS precipitation (> 95% removal) and neutralization • ZVI removes metals and H2 S (> 400 d) • Algae carbon sources sustain COD levels for > 20 months • Ineffective for Co and Ni • Sulfide precipitates can adsorb As

Alkaline

Neutralizing Zn, Ni, Co, substances Cu (e.g., Pervious concrete, fly ash, MgO, recycled concrete aggregate, dregs, grits)

References

[19–27]

• Neutralization induces [28–35] metal hydroxide precipitation • SO4 − removal requires calcium-based materials • Concrete materials/fly ash release Cr and elevate pH > 9 • Susceptible to clogging in the long run • Effective for Co and Ni (pH ≥ 9) (continued)

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Table 4.1 (continued) Type

Materials

Targets

Mechanisms/highlights

Dispersed alkaline substrate

Neutralizing substances (e.g., Limestone, BaCO3 , Clam shells) Granular material (e.g., wood chips)

Cu, Zn, Cd, Co

• Granular and alkaline [13, 36–39] materials blended to prevent clogging • Calcite-rich wastes are suitable substitutes for limestone • Limestone incapable of inducing metal hydroxide precipitation of divalent metals • Effective for Co and Ni (pH ≥ 9)

Adsorptive

Adsorbents Cu, Cd, Ni, (Apatite, activated Zn, Co, Pb, carbon, bentonite, waste Se foundry sand, etc.)

• Metals removed via electrostatic attraction, ion-exchange, chemisorption • Small, porous particles exhibit high metal removal • Susceptible to metal desorption • Not suitable for mine drainage with high acidity • Apatite removes metals via sorption, ion-exchange, and precipitation

Reactive

Reactive Material (ZVI) Dispersant Material (Sand, zeolite, pumice)

• ZVI removes metals [20, 26, 45–54] via reductive precipitation and complexation/sorption on its corrosion products • Dispersants stores ZVI corrosion products and prevents clogging • ZVI’s pre-acidification, sulfidation, and Fe dosing increase Se4+ reactivity • Cd and Zn cannot be reduced by ZVI

Cu, Cd, Ni, Zn, Pb, Se

References

[11, 40–44]

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Despite these limitations, ZVI remains a research interest owing to its negative reduction potential and strong reducing powers. One study by Statham et al. assessed the performance of ZVI under freeze-thawing events to assess its applicability in colder climates [55]. The first freeze-thawing event reduced the HRT by 15–18% from induced particle agglomeration but subsequently stabilized. Minimal hydraulic conductivity losses were observed and thus were deemed suitable for colder climates. These results contradict alternate studies demonstrating severe hydraulic losses in ZVI from operation [56–58]. Thus, recent research has pivoted away from using traditional ZVI as a stand-alone media for PRBs. ZVI is commonly blended with granular material to improve its long-term permeability and cost [45]. The implementation of granular material with ZVI additionally broadens the potential removal mechanisms within the capture zone and enables better accessibility to ZVI reaction-sites. For instance, Bilardi et al. and Madaffari et al. used lapillus as a cheap dispersant for ZVI [45, 46]. Madaffari et al. first demonstrated the efficacy of lapillus a dispersant for Ni removal under different ZVI to lapillus weight ratios (10:90, 30:70, and 50:50) [46]. Ni removal and hydraulic conductivity were positively and negatively correlated with ZVI content, respectively. Lapillus blends with ZVI exhibited competing effects as the loss of hydraulic conductivity created preferential flow patterns that reduced the residence time of the PRBs in the long run. Thus, a 50:50 volume ratio was optimal for maximizing metal removal and hydraulic conductivity. Similarly, Bilardi et al. used a ZVI lapillus mixture to treat a contaminated plume with Ni, Zn, and Cu [45]. Like that observed by Madaffari et al., a 50:50 volume mixture was deemed superior in removal but still suffered from hydraulic conductivity loss. A 30:70 volume ratio maintained a hydraulic conductivity that was four orders of magnitude greater than that seen for the 50:50 mixture. Thus, a 50:50 volume ratio of ZVI and lapillus requires close monitoring of HRT. Alternatively, a 30:70 volume ratio of lapillus and ZVI requires greater reactive media depths and area to meet target removals. All three metals were removed by adsorption onto iron corrosion products when using the recommended compensations. Another common dispersant for ZVI is pumice [47]. Accordingly, Moraci and Calabrò studied a 30:70 weight ratio of ZVI to pumice for Ni removal from a contaminated plume [59]. Unlike lapillus, pumice demonstrated synergism with ZVI by storing its reaction products in its pores and removing pollutants using Fe2+ or electron transfer on the corrosion products’ surface. Over a 58-week testing period, there were no hydraulic conductivity losses, a remarkable improvement from the four-order magnitude loss seen for the ZVI control. Additionally, Cu could easily be removed from the mixture, while Ni required a longer residence time for removal [59]. Although a 30:70 weight ratio demonstrated resiliency in Moraci and Calabrò’s study, Bilardi et al. removed Cu, Ni, and Zn sufficiently with only a ten wt% [47]. Increasing ZVI content surpassed ten vol% had marginal improvements in removal, as all columns achieved > 99% removal. However, only the 10 wt% columns could maintain good hydraulic conductivity after 90 days. When implemented at higher dosages, ZVI can aggregate and reduce the number of active sites available and overall porosity in the reactive zone. These phenomena collectively reduce metal

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removals and flow throughputs. The discrepancy among trials is likely from variation within the mine drainage itself, as pH and other drainage constituents can alter the degree to which contaminants react with ZVI and the electrostatic interactions among ZVI components. Zeolites as ZVI dispersants are of interest since they are cheap materials that are widely available and are capable of sorption and ion exchange with surrounding metals [48, 49]. Additionally, zeolite surface modifications with iron or iron oxides by ZVI exposure exhibits synergistic effects for their metal adsorption capacity. Thus, zeolites were used as skeletal structures to stabilize and disperse ZVI and micro-ZVI (m-ZVI) in PRBs [48]. Zeolites and ZVI worked synergistically in PRBs to improve Cd, Cr, and Pb removal capacities [48, 50]. Similarly, modified commercial zeolites with NaCl, HCl, and HDTMA-Br were used as nano-ZVI (n-ZVI) stabilizers [51]. Sodium-doped zeolites with n-ZVI outperformed all other pre-treatments, creating composites with higher specific surface area than n-ZVI and zeolite alone. The overall surface area of the composite increased from ZVI forming Fe(II) phase porous layers and structural channels on the zeolite surface. The ZVI-zeolite composite enabled the ORP of the solution to remain constant throughout testing, demonstrating that n-ZVI maintained its reduction abilities when stabilized with sodium-doped zeolite. Its strong reduction potential enabled the removal of Cd, Cr, Pb, Zn, Ni, and Cu (> 90%) and eliminated the rapid agglomeration of n-ZVI [60]. However, Cd and Zn were insensitive to n-ZVI reactions, and instead, the metals adsorbed and formed surface complexation on iron oxyhydroxides. Thus, zeolites can provide additional adsorption sites to remove metals insensitive to ZVI reduction reactions (e.g., Cr). Nonetheless, n-ZVI implementation is not ideal since nanomaterial effects in aquatic environments are not well understood. Leaching studies should be conducted priorly to its adoption to ensure long-term n-ZVI stability. Granular-activated carbon (GAC) has also been used as a ZVI dispersant [52]. GAC was first compared with ZVI as a granular reactive material for PRBs due to its ability to remove metals by complexation and electrostatic attraction. GAC maintained higher hydraulic fluxes than ZVI but did not exhibit the same degree of removal. There was synergy with a 30:70 weight ratio of ZVI:GAC that improved the reactivity and hydraulics of both materials. Like that seen for zeolites, GAC acted as a skeletal support for ZVI while providing additional adsorption sites for metal remediation. Thus, GAC is an alternative material that can successfully branch ZVI along the capture zone of PRBs. Decisions on the choice of material should be according to their availability, regeneration abilities, carbon footprint, and price. Moreover, Dong et al. used bentonite and sand to disperse ZVI for Se removal [53]. Bentonite was chosen as a dispersant since it is easily functionalized and can reduce iron passivation by transferring corrosion products from ZVI onto its surface. The ZVI and bentonite mixture completely removed Se and demonstrated synergy since Al-functionalized bentonite electrostatically attracted Se in solution and stored ZVI precipitates and corrosion products. However, competing ions (SO4 , Cl− , NO3 − , and humic acids) reduced their performance. The pre-corrosion of ZVI using H2 O2 and HCl can eliminate the effect of competing ions for Se removal [61]. Specifically, H2 O2 and HCl pre-treatments acidified ZVI with elements Fe0 , hydrated FeO, and

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Fe3 O4 , which increased Se removal from sulfate-rich water from ~ 60% to > 90%. Additionally, Se removal was sustained over pH levels from 3 to 9, demonstrating strong buffering capabilities without competing ion effects (e.g., Cl, CO3 , NO3 , Na, K, Ca, and Mg). The enhancement in Se removal stemmed from a stepwise reductive pathway of Se6+ to Se4+ and subsequently to elemental Se. The pre-treated ZVI showed higher selectivity toward Se oxyanions, causing an enrichment on the ZVI surface before their reduction. Fan et al. used sulfidation and Fe dosing as pre-treatments for ZVI before treating Se under aerobic conditions [62]. Se in solution adsorbed onto the ZVI surface and reduced to elemental Se, with Se removal and rates increasing by 1.8–32.8 times and 11.7–194 times, respectively, compared to pristine ZVI. The sulfidation of ZVI accelerated its corrosion which augmented Se removal, while Fe dosing facilitated electron transfer by forming the semiconductor Fe3 O4 . Thus, ZVI surface modifications are integral to Se treatment in complex mine drainage, where the latter approach minimizes environmental impact since it negates the use of harsh chemicals (i.e., secondary pollution). As demonstrated in Fig. 4.3, several life cycle assessments have been conducted on PRBs utilizing ZVI as the reactive material [9]. As shown, ZVI contributes to > 43% of each LCA category (e.g., global warming, human health), exceeding the contributions from steel ( 9). Moreover, since alkaline PRBs increase pH levels to precipitate and remove dissolved metals, they will likely be unsuitable for metals (e.g., Co, Ni) with high solubility under near-neutral pH levels (6–9). These metals may be removed via co-precipitation by their inclusion, adsorption, and occlusion onto precipitate surfaces. However, co-precipitation offers little control and is unsuitable for treating mine drainage with high metal concentrations. Recycled construction materials permit the repurposing of highly produced waste by-products that meet the treatment targets attained for the benchmark, lime. Integrating their use in PRBs can offer a sustainable solution that solves two waste problems simultaneously. Accordingly, one promising recycled alkaline material is spent-magnesia (MgO)-carbon refractory bricks [28, 29]. Once introduced into water, this recycled material produces Mg(OH)2 , achieving quick equilibria (~ 8 to 24 h) and a sustained pH level of ~ 9.8. Mg(OH)2 sustained alkaline pH levels that completely removed Fe, Ni, and Co by precipitating them as metal hydroxides

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with high crystallinity. Although this material provides a cheap means to remove metals, its column experiment failed due to the quick passivation of Fe precipitates. Thus, de Repentigny et al. tested centripetal flow regimes to reduce passivation [28]. Unexpectedly, this flow regime accelerated the breakthrough in the column due to preferential flows created from short circuiting. The centripetal flow regime should be re-assessed using more coarse materials, as this may alleviate short-circuiting issues and reduce metal passivation. One example of a coarse, alkaline waste material is pervious concrete. Pervious concrete is a Portland cement mixture containing single-size coarse aggregates comprised of water and minimal amounts of sand. Typically, pervious concrete acts as a pollution sink in city operations (e.g., city drains), retaining particles by filtration mechanisms. The alkaline and porous nature of the material is advantageous for mine drainage remediation since these attributes increase pH levels and hydraulic conductivities, respectively [30–32]. Once in contact with water, pervious concrete releases CaCO3 and, in turn, precipitates dissolved metals as hydroxide precipitates. CaCO3 necessitates strong acid waters and is unsuitable for neutral and slightly acidic mine-impacted waters. Nonetheless, its performance was comparable to ZVI under strongly acidic water, removing Al, Fe, and Zn completely while achieving 97%, 93%, and 70% removal of Ni, Co, and Cu, respectively. The porous nature of the pervious concrete further improved the PRBs lifetime twice as much as ZVI. There were no signs of exhaustion after 26 weeks, which could alleviate one of the critical hurdles of PRBs [58]. However, its fly ash component released Cr and created alkaline pH levels unsuitable for receiving waters (10–12); pre-treatments or blends may eliminate these hindrances. Jones and Cetin observed the same issues with fly ash [33]. The original motive was to use fly ash, a coal combustion waste product, as it contains fine particles and is produced at > 50 million tons in the United States with little reuse. Although the fly ash had sorption abilities for Cr and Fe, the material desorbed metals from its structure into the treated mine drainage. Thus, fly ash must be pre-characterized to ensure its suitability for treatments. Cement kiln dust (CKD) is a construction waste material that dissolves in water, releasing CaO and increasing solution pH levels to alkaline [34, 63]. CKD is a byproduct found in cement kiln exhaust gas and offers a sustainable solution since it is a waste product produced in quantities of millions of tons. CKD is a dust mixture of incompletely calcine and unreacted raw materials (e.g., clay, limestone, iron ore, silica) with augmented alkali functionalities (e.g., sulfates and halides). Thus, it is fine-grained (i.e., high surface area and porous) with a high abundance of carbonates (i.e., neutralization capacity) [34]. In batch testing, CKD removed 80–100% of Ni in a simulated CND, where precipitation and sorption dominated its removal at high and low concentrations, respectively. Additionally, CKD exhibited high Cu and Zn removal (> 98%) but experienced too high pH levels and ultimately failed due to clogging [63]. Coupling CKD with a hydraulically conductive material (e.g., sand) can alleviate clogging and will be discussed in the next section [63]. Similarly, recycled concrete aggregate (RCA) is a construction waste material that is strongly porous with multiple alkaline components (e.g., limestone, Cabearing minerals, and Portland cement) produced in large volumes in the United

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States (> 50 million tons) and with low reuse (< 50%) [33, 64]. Four grades of RCA tested could elevate AMD pH levels to > 11 from their releasement of CaO and MgO [33]. Reducing conditions were further established, achieving variable removals for Cr (22–62%), Fe (> 98%), Cu (33–80%), Mn (80–100%), and Zn (> 95%) removal [64]. Thus, the effectiveness of RCAs in PRBs depends on the metal speciation and may not be ideal for specific targets (e.g., Cr). The RCAs also removed sulfates from AMD significantly and may be a viable option for sulfate remediation. Overall, although RCAs were effective for some metal removals and sulfate, they did not target all metals with the same efficiency, and increased pH levels above acceptable levels. A blend with RCA or different grades may experience better efficacies. Similarly, dregs and grits from chemical liquor recovery in kraft pulp mills are suitable alkaline materials for PRBs (pH ~ 10.5–13) due to their CaCO3 and CaO (grits) composition [35]. However, only dregs achieved high Cu removal (99%) in column tests using a solid-to-liquid ratio of 1:10, likely stemming from its pH induction of 10. Dregs further removed sulfates by adsorption in simple batch tests due to its high point-zero-charge (~ 9.5–9.75) that correlates to positively charged moieties (i.e., Ca2+ ). However, sulfate removal failed (< 5%) when treating complex mine drainage from competing ion binding and sorption. Therefore, dregs may offer a cheap solution for Cu and Cu-like metals but requires blending to achieve appropriate sulfate removals. Dregs also have plastic characteristics and low permeability; incorporating porous material with dregs can prevent clogging and preferential flows. Overall, alkaline PRBs invoke metal precipitation using cheap waste material containing CaCO3 , CaO, and MgO components. Waste materials require inert support materials with high surface areas to prevent clogging and improve their applicability in the field. Support materials can also buffer pH levels and broaden metal targets. A common drawback of alkaline PRBs is their inability to remove sulfates from water, where RCA shows the most promise, and metals with high solubility under nearneutral pH levels (6–9). It is imperative to test the ability of alkaline amendments to adsorb or precipitate sulfates and co-precipitate soluble heavy metals before their implementation if these target(s) are treatment goal(s). Alternatively, a polishing step can remediate sulfates post-treatment, or a biological PRB can treat sulfates.

4.2.4 Dispersed Alkaline Substrate (DAS) PRB Dispersed alkaline substrate (DAS) PRBs integrate alkaline-reactive material with hydraulically conductive material (i.e., coarse-grained, inert material) with higher surface area (e.g., sand, wood chips) to mitigate clogging from passivation and increase the number of exposed sites on alkaline amendments for precipitation. For example, Schwarz et al. dispersed limestone with wood shavings in an 88 to 12 weight ratio to reduce limestone passivation [13]. Although demonstrating high Zn and Cu removal (98–100%), passivation of the reactive material persisted, causing a reduction in limestone dissolution (i.e., limestone armoring persisted) and, in

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Fig. 4.4 a A calcite hardpan and b Fe- and Al-based precipitate distribution formed in a DAS PRB with a limestone and wood shaving blend. Reproduced with permissions from [13] and [36], respectively

turn, delaying pH increases. The reactor failed from the production of a hardpan comprised of calcite, wood chips, and precipitates (Fig. 4.4a). The hardpan may be harvested periodically, or the PRB design configuration can adjust to emulate a diffusive exchange system discussed later. Dispersants in this study were ineffective at improving the longevity of alkaline PRBs from the high limestone weight ratio that could not disperse adequately. Accordingly, limestone passivation was limited to the first 20 cm of the column once limestone was blended in a 20:1 mass ratio using wood shavings (Fig. 4.4b) [36]. As mentioned in the preceding section, CKD is a promising alkaline material but quickly passivates. To overcome the clogging challenges observed in the alkaline PRB design, CKD was tested in a DAS configuration with sand as the hydraulically conductive material [63]. The CKD DAS PRB treated artificial drainage with Cu and Zn (100 mg/L) using a 20:80 weight ratio of CKD to sand, achieving >98% removal of Cu and Zn. There was no exhaustion of the reactive material in the 99-day testing period, and the sand created synergism with the CKD by dispersing it and exposing additional sorption sites. Thus, CKD is a promising material in a DAS configuration and warrants an assessment with industrial mine drainage. Furthermore, Millán-Becerro et al. tested a multitude of alkaline reagents (limestone, BaCO3 , biomass ash, fly ash, MgO, Mg(OH)2 , Ca(OH)2 , and calcium-enriched seashells) with woodchips in a 20:80 weight ratio for phosphogypsum water (i.e., acidic water with high metal and sulfates content) [37]. The hydraulics of the tested

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reactors showed no adverse change in the 556-h tested period, further demonstrating the importance of weight ratios between the alkaline and dispersant materials. However, limestone was insufficient at removing divalent metals (i.e., Zn, Cu) and sulfate in the tested reactor since its pH levels (pH = 4.8) were insufficient for metal hydroxide precipitation. BaCO3 , a carbonate source, increased the solution pH to 6.8 and achieved 95% removal of both divalent metals. Nonetheless, both amendments were still insufficient at removing sulfates (≤ 30%) and thus should be avoided or supplemented when sulfate removal is the treatment goal. The low sulfate removal with BaCO3 was unexpected as Ba-ions complex with sulfate and precipitate as BaSO4 in solution through the following reaction, BaCO3 + Ca2+ + SO2− 4 → BaSO4 + CaCO3

(4.3)

Sulfate removal with BaCO3 increases with increasing BaCO3 concentrations, increased acidity (i.e., improves dissolution), and can only remove sulfates associated with Ca and not that found with Mg (i.e., MgSO4 ) [65]. The hindrances observed from Mg are because of the low solubility observed with BaCO3 . Specifically, CO3 needs to be removed as CaCO3 to continue the dissolution of BaCO3 and therefore the removal of BaSO4 ; if introduced as MgSO4 CO3 cannot be removed as MgCO3 . This phenomenon can extend to other cationic species that couple with sulfates and impede the dissolution of BaCO3 . Thus, the presence of competing ions (e.g., Mg), within the phospho-gypsum wastewater may inhibit BaCO3 from precipitating sulfates in the treated water, or perhaps the dosage of BaCO3 implemented was insufficient. On the other hand, to improve the environmental impact of DAS PRBs alkaline industrial wastes, biomass ashes, and fly ashes were tested [37]. Biomass ash was subpar in divalent metal removal, while fly ash nearly removed all tested metals apart from Cu (> 87%). The metal removal seen for fly ash contradicts previous studies and thus may be highly dependent on the fly ash source and its interactions with the constituents in mine drainage. Moreover, oxide and hydroxide reagents (i.e., MgO, Mg(OH)2 , Ca(OH)2 ) removed most metals (Cu, Zn, Cr, Al, and U) but, like limestone was insufficient at removing sulfates (< 30%). Overall, Ca(OH)2 -DAS removed the highest amount of net acidity and total removal of dissolved metals but will incur higher costs and should be replaced with waste materials abundant in Ca(OH)2 moieties. Complementary to the results of Millán-Becerro et al., Torres et al. investigated the efficacy of limestone and BaCO3 as reactive material in DAS PRBs [38]. The limestone and BaCO3 were mixed with wood shavings in a 50:50 weight ratio and operated as a sequential treatment with limestone preceding BaCO3 . BaCO3 was the polishing step since it can remove sulfates and other minerals in solution. Limestone increased pH levels to circumneutral levels (pH = 6) and removed Fe, As, and Al < 1 mg/L, but could only remove divalent metals Cu and Pb. The second stage with BaCO3 increased the pH further (pH = 9), removed sulfates to < 22 mg/L, and divalent metals effectively (Cd > 99.4%, Co > 99.6%, Cu > 99.9%, Ni > 97.9%, Zn > 99.9%). Thus, BaCO3 is more effective than limestone and can remove hardness generated from limestone dissolution (e.g., Ca, Mg). The contradiction between

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Torres et al. and Millán-Becerro et al. results is likely from additional competing ions in phosphor-gypsum water (e.g., PO4 , F) that can impede sulfates removal. To improve DAS PRBs economic feasibility, Larraguibel et al. investigated a sequential PRB technique with one stage comprised of waste material: calciteenriched clam, egg, or mussel shells in a DAS configuration with wood shavings in a 1:4–1:5 volumetric ratio; and the second stage used BaCO3 to remediate Mn and sulfates from mine drainage [39]. The tested materials in the first stage (i.e., clam, egg, and mussel shells) exhibited near complete removal of trivalent (Al, Fe) and divalent metals (Cu, Zn) and achieved circumneutral pH levels (6.8–7.3); nonetheless, the seashells dissolved more quickly and were therefore deemed most compatible for DAS PRBs. The post-treatment stage with BaCO3 agreed with Millán-Becerro et al. since there was only moderate removal of sulfates (~ 0.7 g/d) from low sulfate loading. Thus, the determining factors for sulfate removal are the loading of BaCO3 , sulfate loading, and competing ions. Regardless, BaCO3 availability is scarce and therefore expensive. Future research should look at BaCO3 blends with additional, more economical amendments and potential, more robust substitutes. Overall, the DAS design eliminated hydraulic losses in alkaline PRBs using optimal weight or volumetric ratios of alkaline and hydraulically conductive material. Optimal volume and mass ratios vary among reactive materials and should be pretested, where promising waste alkaline materials (e.g., fly ash, CKD, and seashells) should take precedence. Nonetheless, a consistent trend among these trials is the issue of sulfate removal. Currently, BaCO3 is the only alkaline amendment that can remove sulfates, and its performance is sensitive to sulfate loading, BaCO3 loading, and mine drainage composition. It is important to note that although limestone and other CaCO3 -derived materials are capable of adsorbing sulfates due to their high point-zero-charge, it is evident that the complexity of mine drainage often limits this removal capacity and can be further limited by armoring (i.e., precipitation of iron- and aluminum-oxyhydroxides or calcium sulfates) [66]. Biological PRBs can alleviate the limitation of DAS PRBs when sulfate becomes a persistent contaminant within the treated mine drainage. They also become attractive when treating divalent metals soluble at near-neutral pH levels.

4.2.5 Adsorptive PRBs Adsorptive PRBs utilize high surface area reactive materials with moieties that can either ion exchange, form complexations, or adsorb heavy metals onto their surfaces by electrostatic interactions [7]. Accordingly, the suitability of adsorptive amendments for mine drainage treatment increases with increasing specific surface area and negatively charged functional groups with ion exchange capabilities. Adsorptive PRBs are suitable for short HRTs and waters with lower acidity, metal, and sulfate content; since short HRTs and lower metal concentrations mitigate the potential of metal saturation and desorption from active sites. The negatively charged ions on adsorptive materials will cause repulsion on sulfate ions, thereby limiting the ability

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of adsorptive PRBs to remediate sulfates. Likewise, unless blended or created with an alkaline material, there will be minimal acidity consumption with adsorptive PRBs. Nonetheless, its use eases PRB operation by eliminating chemical requirements and producing minimal sludge [40]. The factors for adsorptive material selection include the targeted contaminant, cost, shelf life, and regenerability [35]. Economically, it is advantageous to use on-site adsorbents as numerous reports have demonstrated waste materials to be efficient metal(loid) scavengers. However, one pertinent drawback is the secondary pollution associated with the regeneration of adsorptive material and metal(loid) desorption upon their saturation. The sourcing of novel, cheap adsorbents for PRBs has been a research interest. For instance, activated carbon is inexpensive and can remove Cd from groundwater to permissible levels for seven months [41]. Thermally activated dolomite and wood ash have also shown promise for neutral drainage treatment containing Ni and Zn [67–69]. The thermal activation of the materials improved their ion-exchange capacities and surface area, leading to improved sorption and removal capacities (Fig. 4.5). This is required as Ni and Zn removal is difficult due to their high solubilities under near-neutral pH (6–9) [68]. Modified wood ash demonstrated the greatest Ni and Zn removal owing to its larger specific surface area of 22.6 m2 /g and greater cation exchange capacity. Unlike dolomite, these modified wood ashes maintained circumneutral pH levels for 112 days. More impressively, activated wood ash was regenerated using a selective precipitation method with environmentally benign Na2 S and a pH control method. Although thermal modifications require intensive energy inputs, it is desirable in hard-to-treat drainages since it creates amendments superior to typical PRB amendments (e.g., fly ash, biomass, activated carbon). Nonetheless, a life cycle assessment is needed to assess whether this material is economical and environmentally friendly. Apatite, a calcium phosphate with a hexagonal dipyramidal crystal structure (Ca5 (PO4 )3 OH) is another promising amendment known for its adsorptive properties towards radionuclides and metals [42, 43, 54]. Once exposed to mine Fig. 4.5 Enhancement in adsorption capacities upon thermal pre-treatments of dolomite. Adapted with permission from [68]

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drainage, Apatite quickly dissolutes under low to moderate pH environments (2.2–7.1) in the following manner, − Ca5 (PO4 )3 OH ↔ 5Ca2+ + 3PO3− 4 + OH

(4.4)

Thus, Apatite generates alkalinity and phosphates in situ to induce metal phosphate precipitation while providing a high surface area for surface adsorption. Oliva et al. first demonstrated Apatite II™, a biogenic hydroxyapatite, as a suitable adsorbent for the removing Cd, Cu, Co, Ni, Pb, Zn, and Hg contaminated water [42, 43, 54]. It elevated slightly acidic pH levels (5) to circumneutral levels while releasing PO4 that induced metal precipitation as metal phosphates and provided active sites for surface adsorption (Fig. 4.6). This enabled near complete removal of Cd, Cu, and Ni as Cd3 (PO4 )2 (s), Cu3 (PO4 )2 (s), and Ni3 (PO4 )2 (s) precipitates, but was ineffective for Co and Hg. Additionally, model predictions demonstrated that this material is efficient for 5–10 years when treating mine drainage with pH > 5 using a 1 m thick layer. Fig. 4.6 a Cd- and b Cu–PO4 precipitates deposited on Apatite II™. Reproduced with permission from [42]

a

b

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Moore et al. demonstrated Apatite as a suitable material for Se adsorption once integrated with carbonate functional groups [54]. The carbonates positively influenced Apatite’s Se removal, increasing its uptake by a magnitude, and improving kinetic rates by 20x. Future research should test its efficacy in long-term experiments, assess its regenerability, and reduce cation competition upon treatment of complex mine drainages (e.g., surface modifications, blending). To reduce costs, bentonite, volcanic ash soil, and red soil were studied as adsorbents for PRBs [40]. Bentonite and red soils were redeemed as suitable adsorbents due to their small particle sizes and porous structures with high specific surface areas, giving them a high affinity for adsorption and cation exchange capacity. Bentonite exhibited the best sorption capacity when tested for Cu, Zn, and Ni contaminated water, achieving 99.9%, 89.2%, and 99.9% removal of Cu, Zn, and Ni and neutralization. Similarly, waste foundry sand and Kerbala sand were investigated for Cu removal in a 1:3 mixture [44]. Waste foundry sand is a green waste sand that can be a cheap adsorbent for Cu removal, while Kerbala sand maintains its hydraulic conductivity. Accordingly, Cu adsorbed onto the carboxylic and alkyl halide groups on the waste foundry sand, achieving up to 93% removal. Although these findings from batch studies show promise for the waste materials, column tests are required to assess their feasibility in PRBs. Several waste adsorbent materials have shown promise, exhibiting strong metal affinities and adsorption capacities to several metal contaminants. However, these materials are often conducted under batch mode and using simplified systems that do not contain the amount of diverse cationic species generally present in mine drainage. These promising amendments require assessments under long-term column tests using industrial mine drainage. Surface modification considerations can also reduce cationic ion competition and improve their regeneration abilities without harsh chemical additions.

4.2.6 Biological PRBs Biological PRBs utilize organic materials or reductants (e.g., ZVI and H2 ) as electron donors to support SRB growth. SRBs reduce sulfates in the influent water to H2 S using organic electron donors. As H2 S reduces, the water alkalinity increases from bicarbonate production in the following reaction, − SO2− 4 + 2CH2 O → H2 S + 2HCO3

(4.5)

The H2 S generated by SRB acts to precipitate dissolved heavy metals as sulfide precipitates using the following mechanism, Me2+ + H2 S → MeS(s) + 2H+

(4.6)

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Utilizing SRB to generate metal sulfide precipitates and alkalinity offers several advantages. First, the growth conditions for SRBs are diverse, ranging from temperatures −5 to 50 °C and pH levels of 2.6–9.5, and are, therefore, robust in the harsh conditions invoked by mine drainage [70]. Although, HRTs can incrementally increase until alkalinity generation is sufficient since SRBs thrive under neutral-alkaline pH levels [71]. Furthermore, the ideal growth for SRB tend toward sulfide conditions with a COD:S ratio of 1.5; COD:S ratios greater than 6 proliferate methanogens rather than SRBs. Nonetheless, with ideal SRB growth conditions, biological PRBs typically exhibit sulfate removal rates of approximately 100 mg/L/day of residence time [12]. Additionally, metal sulfide precipitates are denser and have lower solubilities than their hydroxide counterparts; this makes their sludge handling costs cheaper (i.e., easier thickening and dewatering) and eases their removal from solutions. Metal sulfide removal has different efficacies depending on the solubilities of the target metals. Specifically, the differences in solubility trend as CuS < PbS < CdS < ZnS < FeS < MnS, meaning Cu and Pb are the easiest to remove via metal sulfide precipitation and that Co and Ni are insensitive to metal sulfide precipitative removal. Although metal sulfide precipitation has advantages, the use of SRB in PRBs makes it susceptible to clogging due to the passivation of metal sulfide precipitation, biofilm formation, and gas (i.e., H2 S) production. Furthermore, biological PRBs require recalcitrant organics that sustain bacterial growth for at least five years [19]. The recalcitrant organics should be locally available in large quantities to reduces costs and environmental impact. Typical waste organics are nonporous and require dispersants (e.g., sand, woodchips) to maintain their hydraulic conductivity. The addition of cellulosic-based dispersants can further supply longterm carbon sources since cellulosic bacteria can slowly degrade these organics and supply labile carbon sources to neighbouring microbes. For instance, Tindade et al. utilized waste material, sugar cane bagasse, mixed with sand to remove Zn and Ni from a contaminated plume [20]. Sugar cane bagasse provided appropriate growth conditions for SRB, promoting pH increases from slightly acidic (5.5–5.8) to circumneutral (6.8–8). SRB induced metal sulfide precipitation that completely removed Ni and Zn from the solution without signs of substrate exhaustion at the end of the 145-day testing period. Similarly, Genty et al. mixed different proportions of organic wastes (chicken manure, leaf compost) and cellulosic waste (maple wood chips, maple sawdust) within a PRB operated under two different HRTs (5 and 7 days) [71]. The cellulosic waste dispersants maintained the hydraulic conductivity for over the 450-day period with > 18 wt%. Each combination promoted SRB growth, thereby increasing pH from 3.5 to 6 and achieving > 90% removal of Cd, Cr, Ni, and Zn. The tested organic wastes also did not deplete at the end of testing, marking them as suitable substrates for biological PRBs. Similarly, bamboo chips, rice husks, pig-farm wastewater treatment sludge, municipal wastewater sludge, and coconut husk chips were tested as potential waste materials to sustain SRB growth in PRBs (Fig. 4.7b) [21]. Their sulfate reduction potential was proportional to their lignin content since lignin is a complex phenolic

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b

a

c Fig. 4.7 Shifts in sulfate removal by a addition of ZVI to limestone, reproduced with permissions from [19]; b addition of whole-cell and lipid-extracted algae to sludge under different Cu concentrations (a = 0 mg/L, b = 10 mg/L, c = 30 mg/L, d = 50 mg/L and e = 0 mg/L), adapted with permissions from [22]; and c sulfate production rate with changing metal concentrations and ZVI addition, reproduced with permissions from [23]

polymer that protects plants against microbial biodegradation. The lignin content of the organics therefore correlate to their biodegradability via, B = −0.028X + 0.83

(4.7)

where B is the biodegradability on a volatile solids basis (VS/TS) and lignin content of VS (% dry weight). Overall, SRB growth was stimulated and sustained using a different blend of biodegradability organics in a 60:20:20 wt ratio of pig-farm wastewater treatment sludge, rice husks, and coconut husk chips, thereby removing > 95% of Cu and Zn over 200 days. This highlights the importance of characterizing the make-up of organics (e.g., lignin proportion) to better estimate their biodegradability and therefore suitable fractions in biological PRBs. A biological PRB with gravel, peat, lime, and limestone performed similarly [72]. The organic amendments established anaerobic conditions suitable for SRB proliferation and subsequent sulfite precipitation of Cu2 S and ZnS. The Co and Ni in the mine drainage preferentially bounded to the dissolved organic material, creating a secondary removal mechanism within the column, and concentrating metals near SRB for biosorption. A unique substrate used with PRBs was whole-cell algae and lipid-extracted algae [22]. Both substrates achieved > 99.5% Cu removal via metal sulfide precipitation

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and maintained circumneutral pH levels for four months. Both substrates exhibited a slow, sustained release of carbon, only demonstrating a 9.3% decline in COD levels. Their slow release can sustain SRB growth for ≥ 24 months and provide a means to remediate algal blooms. Another promising approach was using ZVI as an electron donor for SRB [19, 23]. ZVI donated electrons to SRBs as it oxidized and produced H2 gas that can directly transfer electrons to microbes via their c-type cytochromes [19]. The ZVI PRB outperformed a limestone and organic-mixed reactor in a 400day column experiment (Fig. 4.7a) [19]. Its success came from its proliferation of SRB that maintained circumneutral pH and > 99.7% removal of Cu, Cd, and Pb with HRTs as short as 1–3 days. ZVI further removed microbially toxic H2 S from the reactor and improved microbial activity. Likewise, a separate study using ZVI as a sole electron donor demonstrated its feasibility of maintaining > 99% removal of Cr, Mn, and Zn in contaminated ground waters. ZVI enhanced sulfate reduction rates by promoting anaerobic conditions as a reducing agent; it further provided Fe2+ upon its oxidation, an element that is beneficial for SRB dehydrogenase (Fig. 4.7c) [23]. Additionally, there were no hydraulic losses when used with microbes, indicating microbes may play a role in ZVI corrosion prevention. Another common trend is the utilization of ZVI with organics. For instance, Sasaki et al. used ZVI with an organic blend of compost leaf mulch, woodchips, sawdust, and trace limestone [24]. ZVI created reducing conditions for complete removal of Se by SRB, while sand additional to the upper and lower bounds maintained its hydraulic conductivity. The successful operation could treat 2 mg/L of Se for 75 years using 1546.8 cm2 of reactive material. Employing organics (compost leaf and woodchips) and ZVI with limestone and sand was further extended to As remediation [25]. The blend successfully increased pH levels from 3 to circumneutral and removed As via adsorption onto Fe(oxy)hydroxides and precipitates. A 20 wt% of organics was optimal, while ZVI played no integral role in As removal and should be added minimally at 10 wt%. Another study assessed ZVI with peat and limestone under different proportions [26]. Herein, ZVI induced additional removal mechanisms via Fe2+ generation that can act as a sulfate scavenger. The removal of Ni, Cu, Al, Co, and Zn increased from the additional adsorption sites and by precipitation and co-precipitation on its corrosion products. The success of ZVI and organic blends led to many pilot studies. For instance, a biological PRB in Aznalcollar, Spain remediated a contaminated aquifer [73, 74]. It utilized the reactive mixture of calcite, vegetable compost, ZVI, and sewage. Calcite chips raised pH levels and caused metal precipitations as hydroxides and carbonates, while the compost and sludge provided carbon sources to stimulate SRB growth. ZVI induced reducing conditions to proliferate SRB, but the sludge ultimately clogged the reactor resulting in its failure after three years of operation. Similar results occurred for a pilot test of a PRB comprised of ZVI, limestone, and pea gravel (volume ratios: 30:20:5:45) [27]. Metal sulfide precipitates on the reactive material reduced the hydraulic conductivity by an order of magnitude within 18 months. Thus, although the PRB converted a net acid-producing groundwater to a net acid-consuming water, its implementation failed due to the passivation of the reactive material. Failure likely occurred due to inadequate volume ratios between organic and cellulosic waste, as

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sufficient void volumes are required to ensure metal sulfide precipitates do not impede mine drainage flows. Overall, biological PRBs are promising for the remediation of mine drainages. The factors determining their success are the lifetime of the organics and the sustainment of hydraulic flows. Thus, complex organic substrates with sustained organics release should blend with cellulosic waste for success. Current research focuses on finding new, sustainable waste materials that fulfill these goals and optimizing their volumetric ratios. Additionally, ZVI can act as an electron donor capable of removing toxic, H2 S, that impedes biological activity and thus metal removal. However, due to the environmental burden of ZVI, its dosage must be mitigated. There has also been a negation of phylogenetic view of these designs. Identifying which microbes are proliferating under certain conditions and how their growth correlates to performance would be beneficial to optimization. Additionally, bioinformatics may yield information on how removal depends on the upregulation of microbial genes. These systematic investigations are required to understand differences in performance and have been negated from literature to date. Current trends for biological PRBs are further trending toward diffusion active or sulfidogenic exchange system PRBs which can protect SRB from mine drainage toxicity.

4.2.7 Diffusion-Active (DAPRB)/Sulfidogenic Exchange System (SDES) PRBs Schwarz and Rittman first proposed diffusion-active PRBs (DAPRB) or sulfidogenic exchange system PRBs (SDES) as a tactic to provide microbial protection from mine drainage toxicity [15]. As demonstrated by Fig. 4.8, DAPRBs consist of intercalating layers of reactive and conductive materials [14, 15]. The conductive material is coarse sand or gravel with hydraulic conductivities ≥ 100 times greater than the implemented reactive material. The hydraulics of the conductive materials ensures that mine drainage flows through the path of least resistance via The reactive lathe conductive material. By isolating the reactive layer from mine drainage flow path, microbes may take advantage of the diffusion-controlled, reactive layer to protect themselves from the drainage toxicity (i.e., metal(loid)s, oxygen, acidity, etc.). The reactive layer should include an electron donor (i.e., particulate organic matter) and an electron acceptor (sulfates) while blended with sand. Electron acceptors, such as gypsum or CaSO4, should be supplemented to ensure SRB activity is not dependent on the sulfates provided by the mine drainage to prevent diffusion limitations. Using these supplements, the protection of microbes within this layer has improved sulfide generation rates compared to biological PRBs (Fig. 4.9) [16]. To ensure the reactor is diffusion-controlled, a minimum reactive thickness layer is required to meet the diffusion-dominant stability criterion for the initial entrance zone where metal toxicity is most pertinent [15]. Assumptions to estimate this reactive layer thickness are that the pore water in the conductive layer has a metal affluent

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Fig. 4.8 Schematic of diffusion-active permeable reactive barriers (DAPRBs) for treating acid mine drainage (AMD) with a scanning electron microscopy (SEM) image illustrating metal deposition at the conductive and reactive layer interface. Adapted with permission from [14]

Fig. 4.9 The improvement of sulfide generation switching from a traditional PRB design to a DAPRB. Adapted with permissions from [16]

concentration (uMo ); sulfide generate in reactive layer at the volumetric rate, r; metal and sulfide gradients meet in between layers interfaces where precipitation occurs; metal diffusion to the reactive zone is proportional to liquid layers thickness (δ bl ), and there is a diffusion coefficient (DM ). By incorporating Fick’s First law, the resulting diffusion-dominant becomes, qD =

r L b δbl ≥1 u Mo D M

(4.8)

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97

where δ bl /L b ~ 0.25 from numerical estimations and the rest of the parameters are from experimental conditions. However, there are smaller empirical estimations for L b , meaning multiple approaches can estimate the desired reactive layer thickness. Another advantage is the mitigation of PRB clogging. Since SRB is limited to the reactive layer, H2 S travels down its concentration gradient toward the conductive layer as it generates from SRB metabolism. Once H2 S reaches the interface, it contacts cationic metals present in the mine drainage, where they will then react and form metal sulfide precipitates along the reactive layer-conductive layer interface. The precipitation along the layers interface prevents precipitates from accumulating along the mine drainage flow path. However, diffusion and transport limitations can occur along the interface if precipitates obstruct the interfacial area. Pérez et al. first demonstrated the efficacy of DAPRBs in a bench-scale study using a reactive mixture of pinus radiata compost, anaerobic sludge, ZVI and gypsum/lime, and sand as the conductive material [17]. Sulfate removal was the same under PRB and DAPRB configurations, demonstrating that sulfate transport to the reactive zone was not diffusion limited. ZnS was then deposited along the layers interface once there was a spike of Zn in the feed. Although demonstrating DAPRB principles, it did not show any enhanced Zn removal from the equivalent PRB design. The short testing period (60 days) may have interfered with the results, as DAPRB configuration showed improved microbial diversity and sulfide generation (Fig. 4.9). Additionally, differences in performance may further be more evident with more toxic metals (e.g., Cr) or with optimized reactive and conductive layers. To optimize the layers, reactive thicknesses 2.5, 5, and 7.5 cm layers comprised of cellulose fiber, leaf compost, bovine manure, and limestone were studied [18]. To achieve complete removal of Zn and Cu, a minimum thickness of 5 cm was required without experiencing breakthrough over the 225-day testing period. The minimum 5 cm thickness contradicted their predicted minimal thickness of 11.5 cm calculated using the bioprotection criterion (i.e., diffusion-dominated stability criterion). Thus, failure might have resulted if testing times increased. Nonetheless, greater reactive layer thickness created more reducing conditions (ORP < 0) and suitable pH levels (> 6) at > 160 d. The growth conditions provided by the greater layer thicknesses promoted higher cellulosic bacteria activity. Cellulosic bacteria activity is imperative as they break down recalcitrant cellulosic organics and sustain the release of electron donors for SRB. Schwarz and Norma further investigated the long-term efficacy of DAPRBs in a bench-scale study for 591 days [14]. The reactive layer was a mixture of pinus radiata compost, biodigested sludge, sand, and calcium sulfate, and the conductive layer was coarse sand. The pH of the mine drainage reached circumneutral levels, and Zn removal ranged from 76 to 99%. The acidity loading in the influent positively influenced the removal of Zn, while Zn removal and sulfate reduction rates were robust against increase metal loadings. ZnS was further deposited along the interface of the layers as a metal sink zone, as indicated by white coloration. The robustness of the DAPRB under increasing acidic and metal loadings demonstrate that its intercalating layers can protect microbes and maintain their metabolic rates. However, the extent to which DAPRB-protected microbes remains unknown as there was no comparison

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with a biological PRB. Furthermore the cost competitiveness of the DAPRB design is highly dependent on the sulfate reduction rate, reactive material lifetime, discount rate, and item cost items (e.g., excavation, reactive material, synthetic liner) [14]. Adjusting the reactive to conductive material distribution allowed for a competitive sulfate reduction rate of 0.6 mol SO4 2− /m3 /d, marking DAPRB a cheaper option at USD 0.62/m3 /d compared to conventional precipitation (USD 0.86/m3 /d). Lastly, Schwarz et al. extended the SDES principle to alkaline-reactive material (ADES) to reduce clogging within alkaline PRBs [13]. The ADES used limestone and coarse sand as reactive material. While an SDES utilizing a blend of cow manure, leaf compost, cellulosic fibers, and limestone was added a control to compare with the ADES performance. The ADES and SDES eliminated clogging issues observed in their PRB control design and achieved high removal of Cu and Zn. From the mentioned studies, it is evident that SDES designs can solve PRB clogging. However, longer-term studies with more diverse plumes and amendments are required to assess the feasibility of DAPRB designs in scaled-up operations.

4.3 Constructed Wetlands Constructed wetlands (CW) are engineered systems that emulate natural wetlands to treat contaminated waters (e.g., sewage, industrial wastewater). As presented in Fig. 4.10, CWs features include natural vegetation (i.e., macrophytes), porous supporting media, carbon sources, and microorganisms. Carbon sources are provided as complex substrates to sustain microbial growth while ensuring slow degradation kinetics for a sustained lifetime [75]. Since substrates can be dense, their hydraulics must ensure proper flows and contact times with mine drainage; and should not obstruct mine drainage flows facilitated by surrounding supportive media [75]. Given the constituents of CWs, its removal mechanisms include biological reactions, phytoextraction, adsorption, cation exchange, precipitation, and co-precipitation. The degree of these reactions depend on the given substrate used, sediment pH, and the nature of the water and macrophytes utilized [76]. Biological reactions driven by SRB growth are the major contributors to heavy metal removal within CWs. Specifically, SRB metabolism sustains sulfide metal precipitation and alkalinity production by in-situ sulfide and bicarbonate production, respectively. Thus, substrates should be selected to ensure optimal SRB growth conditions. These include higher pH levels (~ 7 to 9) and COD to S ratios of 1.5. If COD to S ratios exceed 6, methanogens will dominate microbial fractions and adversely affect the efficacy of the CW by producing methane, a harmful greenhouse gas. SRB usage enables CWs to operate under varying temperatures (− 5 to 50 °C) and pH levels (2.6 to 9.5). SRBs also catalyze metal sulfide precipitation which eases sludge handling and costs since sulfide precipitates have heavier densities than their hydroxide counterparts. They also exhibit lower solubilities over a broader pH range, enabling better removals of pH-insensitive metals. Metal sulfide precipitation, therefore, offer further advantages to using CW in addition to their simplicity, cost-effectiveness, and environmental friendliness.

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Fig. 4.10 Schematic of a typical constructed wetland configuration and its associated removal mechanisms. Reproduced with permissions from [77]

Macrophytes that withstand low pH levels and wide temperature ranges are integral to CW designs. First, macrophytes immobilize metals, removing them from environments and storing them in their roots and/or shoots (i.e., phytoextraction) [78]. Excluder species are macrophytes that accumulate metals along their roots, thereby protecting their rhizome and shoots from metal toxicity. These are opposite from accumulator species which localize and immobilize metals along their shoots. Excluder species are preferred, as it prevents the accumulation along the shoots and the potential of wildlife exposure. Bioconcentration factor (BCF) and translocation factor (TF) are metrics used to quantify metal distributions within macrophytes. BCF quantifies the ability of the macrophyte to accumulate metals within its roots using the following concentration ratio, BCF =

[roots] [soil]

(4.9)

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As such, macrophytes with a BCF > 1 are desirable. TF instead measures a macrophytes translocation efficiency of moving metals from the root to shoots using the following concentration ratio, TF =

[shoots] [roots]

(4.10)

TF < 1 is therefore required to ensure metals retain within their roots. Additionally, macrophytes exhibit different metal affinities based on the species, and their selection should be according to metal target(s). For example, T. angustifolia has a high Ni affinity and is beneficial for its sequestration [79]. Nonetheless, metal accumulation along the roots causes plaque formation, as shown in Fig. 4.11, which reduces their metal uptake. Thus, macrophytes require harvesting and replanting periodically to ensure continual metal removal. Standing stock is a metric utilized to quantify the total accumulation of metal within a particular macrophyte component and is calculated based on biomass production per unit area and metal concentration. It can therefore be an indicator of when macrophytes require harvesting. Additional phenomena occur with macrophytes in CWs [77]. First, they oxygenate the water column and upper sediments via oxygen release from their roots. Their oxygen release mediates the oxidation of dissolved trace metal(loid)s and their Fig. 4.11 SEM images of vetiver grass a with and b without Fe-plaque formation. Reproduced with permission from [78]

a

b

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101

deposition in the deeper sediment layers. It further creates oxygen gradients with anoxic sublayers since there is an oxygen demand from microbes and contaminants. The anoxic sublayers enable SRB to thrive and generate sulfide compounds that can easily bind to trace metals. Organics within this layer also act as metal scavengers with trace elements within the sediments, while the roots of macrophytes provide a surface for fixed-film SRB growth. Biofilms create microenvironments that protect microbes from harsh environmental conditions and improve their metal uptake capacity. Finally, macrophytes can generate organic carbon as exudates and detritus that feed into SRB growth. Nevertheless, there have been some drawbacks identified with macrophytes and SRB growth. These include a local pH reduction from organic acid release and oxygen transfer within the rhizosphere where SRBs reside [70, 78, 80]. Oxygen transfers within the rhizosphere are particularly problematic if there are insufficient carbon amendments since aerobic conditions can arise [81]. Although aerobic conditions may proliferate algal species as trace metal scavengers, they have been deemed subpar to anaerobic CWs that use SRB [82]. Thus, sufficient organics must be present, particularly with macrophytes present. With sufficient organic amendments, macrophytes accentuate CW performance through their metal uptakes and proliferation of biofilms. Macrophyte selection overall should according to their (1) resistance to AMD conditions (i.e., metals, acidic pH levels, salinity); (2) their affinity to target metal(loid)(s); (3) their TF/BCF values for target metal(loid)(s); (4) biomass production; and (5) ecological amplitude [83]. Commonly implemented macrophytes that have shown promise are from the Typha and Chrysonogan families [84]. Other considerations for CW are environmental factors, as CW performance is sensitive to cold climates and heavy rainfall events. For instance, heavy rain falls create washouts under drying and rewetting events [85]. These events release adsorbed metals and metal precipitates into nearby environments and cause further contamination. Thus, the severity of these events should be assessed with the supporting matrix before implementation; CWs may be unsuitable for areas with heavy rain falls. Furthermore, cold climates are problematic for numerous reasons. First, it reduces the kinetics of microbial metabolic processes that goven metal sulfide precipitation and bicarbonate production rates [70]. Stein et al. also showed that methanogens proliferate over SRBs under colder climates regardless of the organic carbon loading, where macrophytes further accentuated their depletion [70]. Second, cold climates can induce freezing. Freezing reduces nutrient availability for microbes. Nutrient depletions further reduce metabolic rates and cause metal leaching from organic matrices [86]. Lastly, freezing events reduce interpore spaces. Interpore spaces determine the contact area for water flow and thus cause an increase in water velocity upon their reduction. Increasing water velocity reduces contact times between mine drainage and soil particulates and, therefore, the physicochemical removal mechanisms reliant on contact times (i.e., adsorption, precipitation, biodegradation) [86]. Thus, CWs may not be suitable in colder climates and require assessment at the lowest conceivable temperatures before being implemented. Moreover, there was a comparison of the economic impact of CWs and conventional pump-and-treat options. CWs are economically advantageous for low-strength

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mine drainage waters [87]. Specifically, they are economically competitive against conventional chemical treatments under low flow (< 20 L/s) and low acidic conditions (< 300 mg/L). These limitations are from their adverse effects on microbial activity under high flows and acid loadings [88]. However, CWs may remain competitive when treating waters with higher acidities if there are low flows and vice versa. CWs are constructed in three flow configurations: horizontal subsurface, vertical subsurface, and floating wetlands. Horizontal and vertical subsurface are more common, using subsurface flows through porous media under parallel or perpendicular flows from its input. Floating wetlands negate the use of supporting media and instead rely on hydroponic macrophytes for treatment. The following sections will discuss each CW configuration with research advantages found in each. Table 4.2 summarizes the findings regarding their mechanisms, highlights, and targets.

4.3.1 Subsurface Flow Constructed Wetlands Subsurface CWs direct drainage flow through an underground supporting, porous media to introduce alkalinity, adsorption sites, and attachment sites for SRB growth and biofilm formation [90]. Traditional media that have been used for wastewater, such as sand or gravel, are unfit for mine drainage with high acidities. Instead, alkaline materials that achieve neutralization (e.g., seashell grit) are suitable to treat AMD [90]. Saturation conditions in the alkaline material can further reduce oxygen influxes and ensure anaerobic conditions for SRB growth. Macrophytes are often strategically placed in CWs to ensure the tight hydrological control necessary for maintaining saturation conditions. As shown in Table 4.2, there are two flow regimes associated with subsurface flow regimes: vertical and horizontal subsurface. The vertical subsurface flow introduces the mine drainage at the inlet surface of the CW and drains the effluent at the bottom end to encourage vertical flows. Despite the design being less commonly implemented, vertical subsurface flows have longer contact times and lower total area requirements with the same economical impact as horizontal subsurface flow CWs (i.e., same amendment dosing) [87, 90]. The horizontal subsurface reduces the degree of contact time by implementing parallel flows yet remain the most studied CW configuration. To successfully treat AMD with CWs, circumneutral pH levels must be quickly established to support SRB proliferation. Strategies to neutralize AMD include using alkaline media, neutralization pre-treatments, or adding organics rich in alkaline moieties. For instance, seashell grit and zeolites are potential alkaline-generating media that can replace the conventional amendment, limestone [80, 90]. Seashell grit are suitable supporting media due to its smooth surface of quarts and calcitetype materials that are resilient against acidic conditions. The exterior of the seashell grits provided binding sites for contaminants, and their calcite moieties generated alkalinity as bicarbonate ions. However, zeolites were ill-suited as a porous media

Materials

Supporting media: gravel, cellulosic waste limestone, zeolite, seashell grit

Organics: domestic wastewater, plant litter broth, urban waste green, compost, activated sludge, liquid brewery waste, sucrose, cow manure, spent mushroom compost

Floating macrophytes Supporting media: water treatment residuals, peat, soil, coconut fiber

Hydrology

Vertical subsurface flow

Horizontal subsurface flow

Floating wetlands

Vetiveria zizanioides, Chrysopogon zizanioides, Carex lacustris, Typha latifolia, Juncus canadensis, Phragmitess australis, Juncus effusus

Typha latifolia, Phragmites australis,

Pseudoacorus L., Schoenopectus acutus, Typha latifolia, Typha angustifolia, Semostachya binnata, Saccharum bengalense

Macrophytes

Zn, Cu, Ni, Pb

Ni, Co, Zn, Cu, Pb

Cu, Cd, Zn

Targets

• Supporting media maintains hydraulic conductivity, removes metal via sorption and ion-exchange mechanisms and can provide alkalinity • Walnut shells promote anaerobic conditions by restricting aeration • Alkaline media aid in start-up since higher pH values allow for insoluble precipitate formation (e.g., oxy(hydroxides), oxides, carbonates, etc.) and co-precipitation • Organics stimulate SRB growth and remove metals via complexation • Organics rich in CO3 and OH− functionalities can buffer pH during start-up • SRB generates alkalinity and removes metals via metal sulfide precipitation • Macrophytes (acid- and metal-tolerant): low metal removal, control system hydraulics, provide surfaces for precipitate collection and biofilm formation, a source of oxygenation at roots • Biofilms remove metals via complexation and physical entrapment and act as a protective barrier for SRB • Precipitation catalyzed by organic acid and cationic nutrient secretion from macrophytes • Floating wetland macrophytes have greater metal removal capacities than subsurface • Select macrophytes (e.g., vetiver grass) reduce sulfate for metal sulfide precipitation and alkalinity generation • Chelators (e.g., citric acid) enhance metal accumulation in plants and reduce Fe and Mn plaque formation on macrophytes • Macrophytes need to have BCF > 1 and TF < 1 to ensure root sequestration • Large macrophyte densities encourage SRB growth in floating wetlands • Vertical subsurface flows reduce surface area requirements • Unplanted subsurface CW more susceptible to heavy rainfall (i.e., metal re-immobilization, desorption, etc.) • CW sensitive to freeze/thawing and washout events • Pre-treatments with wastewaters can improve influent quality and therefore CWs’ efficiency for hard-to-treat AMDs (pH ≤ 3) • Macrophytes and high loading rates reduce sulfate reduction

Mechanisms/highlights

[78, 96–101]

[80, 85, 90, 93–97]

[70, 75, 79, 82, 90–92]

References

Table 4.2 Summary of constructed wetland findings, their materials, macrophytes, target metals, and mechanisms/highlights. Images reproduced with permission from [89]

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due to their inability to generate alkalinity in AMD. Instead, it experienced limited removal via adsorption under acidic conditions. Macrophytes adversely affected the zeolite’s performance, releasing organic acids and further propagating the reduction in pH. Nonetheless, different zeolite-derived materials (e.g., fly ash) may achieve neutralization and therefore require further investigation along with other calcite-rich, porous waste amendments. An alternative approach to neutralize AMD is to pre-mix AMD with another wastewater source [91]. For instance, domestic wastewater and AMD pre-mixing achieved and sustained throughout CW treatment. This pre-treatment may further alleviate clogging issues in CWs since AMD contaminants prematurely make complexes with suspended solids and precipitate with phosphates in the domestic wastewater. Thus, the pre-neutralization of AMD with domestic wastewater enabled high removals (> 96%) of all present metals (Zn, Cd, Fe, Cu) by inducing metal hydroxide, oxide, and phosphate precipitation. Thus, future research should extend this pre-mixing step with alternative waste streams nearby affected mining sites. Lastly, the most common approach to generating alkalinity is using organic sources rich in alkaline moieties (e.g., carbonates, hydroxides). For instance, Sheoran et al. constructed a CW with powdered goat manure and soil as supporting media, and woodchips (cellulosic waste) with gravel along the bottom layer [79]. Goat manure as a carbon source sustained SRB growth which removed more than 70% of Zn, Cu, and Co and neutralized pH levels after 24 h. Similarly, Singh and Chakraborty tested the efficacy of bamboo chips (i.e., cellulosic waste) and cow manure (organic waste) with the common cattail, Typha latifolia [93]. The system treated AMD contaminated with Ni and Co over six months. SRBs were the major contributor to neutralization and metal removal by producing bicarbonate and sulfides. The same system treated a Zn-contaminated drainage in a separate study, demonstrating the robustness of the organic blends [94]. Likewise, plant litter broth and domestic wastewater were successful organics in CW start-up and demonstrated synergism with walnut shells as cellulosic waste [75]. Walnut shells alone improved removal by providing additional adsorption sites through their hydroxyl and carboxyl functional groups; they further supported plant growth and microbial growth by serving as a biofilm carrier (Fig. 4.12). Reducing redox potentials, suitable for SRB growth, formed by supplementing domestic wastewater and plant litter broth. The walnut shells synergistically promoted reducing redox potentials by restricting aeration through their consolidated packing. Collectively, these factors enabled SRB to proliferate within 20 days rather than the 50 days required for the gravel bed matrix CW control. Additionally, there was an increase in microbial diversity and richness; owing to a shift from aerobic acidophilic sulfuroxidizing microbes to facultative and obligative anaerobic, fermentative, sulfate- and iron-reducing bacterial populations by supplementing two different carbon sources. Other successfully tested organic blends include activated sludge and mushroom compost, further illustrating the importance of blended organics to proliferate a more diverse portfolio of microbes for successful CW treatment [92, 95]. It remains a debate on the main functions of macrophytes within CWs. Thus, Collins et al. implemented CWs with real, fake, or no macrophytes [82]. The CWs

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105

a

b Fig. 4.12 a Enhancement of metal removal using cellulosic waste, i.e., walnut shells (W-W), as supporting media over gravel (G-G); b Additional removal mechanisms switching from a gravelbased (G-CW) to a walnut shell-based CW (W-CW). Reproduced with permission from [75]

with real and plastic macrophytes had similar microbial compositions, but there were lower pH levels and higher Fe and Mn concentrations with real macrophytes. These findings complement another study that observed declining microbial sulfate reduction in the presence of T. latifolia due to oxygen release from its roots [88]. Similarly, another study investigated the effects of different macrophytes’ (Typha angustifolia, Phragmites australis, Phalaris arunginacea, unplanted) rhizospheres on bacterial communities in CWs [102]. All tested macrophytes increased microbial density and activity, yielding aerobic and facultative anaerobic microbes. Macrophytes were found to promote greater bacterial densities due to their ability to act as a surface for biofilm formation and provide carbon and enzymes (e.g., protease, amylase, etc.) for microbial growth in the rhizosphere via root exudates. Furthermore, root surfaces correlated to bacterial density, aerobic respiration, dehydrogenase activity, and protein/enzymatic activities, while root and shoot morphology of Phalaris arunginacea supported greater microbial density and activity. Overall, subsurface CWs show promise in mine drainage remediation when using alkaline-generating media, pre-treatments with other wastewaters, or organics rich in alkaline moieties. CWs must have organic blends with cellulosic waste; cellulosic

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wastes maintain hydraulic conductivities and create synergisms by providing additional sorption sites, improving microbial diversity, and serving as biofilm carriers. Macrophytes exhibit little uptake of contaminants in CW systems and are beneficial for hydraulic controls and promoting robust and diverse biofilms. Research should continue to search for waste alternatives that can serve as alkaline media while sourcing alternative waste streams that can pre-mix with AMD priorly. Additionally, bioinformatic tools can enable a better understanding of the implications of materials regarding microbial profiles and functionality.

4.3.2 Floating Constructed Wetlands Floating CWs are constructed without porous media or carbon amendments and therefore depend on the metal uptake of hydroponically supported macrophytes and the biofilms enriched along their roots [78]. Biofilms are complex matrices capable of creating metal complexations and physical entrapments [98]. Macrophytes accentuate physical entrapments in CWs by forming metal complexations using phytochelatins and metallothioneins, where thiol groups in cysteine can bind to metals and transport them into their vacuoles [99]. Floating CWs are selected for their cost-effectiveness and for being environmentally friendly but there is a critical limitation [78]. The lack of carbon amendments in floating CWs promote aerobic environments that hinder SRB growth required for metal sulfide precipitation and bicarbonate production. Thus, this design is susceptible to washout events since physical entrapment is not a long-term solution. However, if designed with sufficient plant densities, anaerobic conditions can be promoted in deeper zones of the CW. While using more diverse plants can deepen these oxic zones to create lower metal sinks and reduce washout events [77]. Additionally, chelators, such as citric acid, can aid in creating anaerobic conditions as an additional organic substrate [100]. Chelators further reduce plaque formation on macrophyte roots by creating complexations with Fe and can extend metal uptake through their rhizosphere. Nonetheless, chelators incur costs and require an economical analysis. The importance of creating oxygen gradients along floating CWs was supported by Buddhawong et al., using Juncus effuses to treat As- and Zn-contaminated water [96]. Since there was an insufficient plant density of Juncus effuses, aerobic conditions prevailed with minimal removal of Zn and As. The same occurrence happened in a two-year field study, which constructed a CW with the hydroponic plant, Juncus effusus [97]. The design ultimately failed, exhibiting < 30% Zn removal, as it solely depended on adsorption and macrophyte metal uptake since reducing conditions were not established. Since floating CWs negate supporting media, research efforts have been put forth exclusively on the efficacy of different hydroponic macrophytes. For instance, Kiiskila et al. tested Chryospogon zizanioides (i.e., vetiver grass), a non-invasive perennial grass that is fast-growing with large biomasses

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[78, 99]. This species is advantageous for its self-generating and dense roots, which provide a large surface area for plaque formation and metal sequestration [78]. A plant density > 16 g/L created anaerobic conditions that proliferated anaerobic bacteria and enabled dissimilatory sulfate reduction and bicarbonate production. The vetiver grass remained viable and removed high amounts of Fe and Pb (> 80%) while removing moderate to low ( 90%), demonstrating macrophytes’ metal uptake limitations upon plaque formation. Another study tested Chryospogon zizanioides and Phragmites australis over six months utilizing 286 plants/m to treat mine drainage contaminated with Zn, Cu, Fe, and Mn [101]. These macrophytes were suitable due to their resistance to high metal concentrations and acidic pH levels. Both species exhibited a BCF > 1 and TF < 1, demonstrating their ability to accumulate metals along the root biomass. Additionally, the macrophytes showed assimilation capacities of the metals, where the order of metal accumulation in C. zizanioides and P. australis were Fe > Zn > Cu > Mn and Fe > Zn > Mn > Cu, respectively. Nonetheless, all floating CWs have failed from freeze-thawing events. Gupta et al. addressed these issues in a novel configuration using built-in shallow sediment profiles capable of withstanding the harsh climates of Sudbury, Canada (Fig. 4.13) [98]. As shown in Fig. 4.13, the CWs consist of wooden pallets with buoyant supports that overlay snow fencing for structural support and burlap and coconut coir layers to prevent soil losses. Juncus canadensis, Carex lacustris, and Typha latifolia were studied since their native to Sudbury, tolerant to AMD, and acclimatized to cold climates. The floating systems built with macrophytes experienced larger fluctuations in redox potentials due to oxygen release (i.e., radial oxygen loss) creating oxic and anoxic microenvironments. Nonetheless, moderate reducing conditions were established after two years, promoting oxygen depletion and circumneutral pH levels. Thereafter SRB richness increased by ~ 30% and the relative abundance against total microbial community increased by 100%. Furthermore, the design was robust against cold climates, maintaining its buoyancy abilities without any visible damage. The macrophytes developed sufficient below-ground biomass capable of regenerating in warmer climates. From the outline performance, one shallow floating CW could treat 61 m3 of water if its sulfate concentration is ~ 500 mg/L. However, this study did not attempt to quantify its metal removal capabilities, and, therefore, future studies are warranted under a broad range of metal contaminants.

4.4 Gravel Bed Reactors (GBRs) Gravel bed reactors (GBRs) are simple, engineered bioreactors that emulate SRFs by utilizing waste rock or gravel as a physical substrate for bacterial growth [103]. The waste rock or gravel placements facilitate water capture in void spacing to promote

108

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Fig. 4.13 Built-in buoyant, shallow sediment profiles to overcome freeze-thawing events in CW. Reproduced with permission from [98]

saturated conditions. Saturated conditions are desirable since they reduce oxygen permeation and promote anaerobic conditions suitable for SRB growth. GBRs supply electron donors to support SRB growth as readily soluble carbon amendments or nutrients which provide carbon substrates through primary algal production. At the same time, buffer additions support SRB growth if the mine drainage influent has a high acid loading [104]. Although more expensive, dolomite is a potential GBR matrix with neutralization capabilities under saturated and unsaturated conditions [105]. Thus, dolomite enables in-situ neutralization within the GBR and promotes metal precipitation as hydroxides and co-precipitation of minor elements. Due to its neutralization capacity, dolomite use may eliminate buffer requirements. Nonetheless, in both cases, the microbial respiration along saturated rocks in GBRs reduces oxygen concentration, creates suitable pH levels (> 5.5), and develops ORP for SRB (≤ − 100 mV). GRBs, therefore, operate under the same principles as SRFs via the stimulation of facultative anaerobic bacteria within waste rock for the bioreduction of nitrates, Se, and sulfates. ORP gradients enable the simultaneous removal of Se and nitrate to below detection levels (µg/L) [106]. While the biofilms formed along waste rock/gravel create microenvironments that enable microbial survival under mine drainage’s harsh environmental conditions. These microenvironments also increase the diversity of microbial communities, thereby broadening the range of treatable contaminants. Additional removal mechanisms for GBRs include (1) Apatite II addition for the precipitation of Pb, Zn, and Cd as phosphate precipitates; (2) acidic pH shifts for As adsorption; and (3) alkaline pH shifts for divalent adsorption [103]. Thus, different environmental conditions can preferentially target specific contaminants while the mass loadings, target concentrations, and treatment rate(s) govern removal rates.

4.4 Gravel Bed Reactors (GBRs)

109

Fig. 4.14 Gravel bed reactor (GBR) schematic, reproduced with permission from [106]

GBRs offer a pump-and-treat method capable of remediating in-pit contamination or developed plumes. Their pump-and-treat capabilities also provide a means to incorporate a polishing step. Specifically, treated waters can be pumped out of the GBR and enter an aeration polishing step. Aeration can remove residual H2 S produced upon SRB metabolisms and elevate oxygen levels before release. Fig. 4.14 presents the overall schematic of a typical GBR, and Table 4.3 summarizes GBR studies with their respective targets and mechanisms/highlights. Like SRFs, there have been limited reports for GBRs due to their proprietary nature. However, from that published, it is evident they hold their merits. For instance, Williams et al. used a proof of concept in an up-flow column to treat a Cr-contaminated site in South Africa [107]. The packed column contained hydraulically conductive dolomite and operated in an up-flow configuration to ensure saturated conditions. The additions of citric acid established reducing conditions that enabled the complete removal of Cr after seven pore volumes. Like Joy’s study, the enhancement in Cr removal correlated to a microbial community shift which maintained > 95% removal of Cr until there was a depletion of citric acid. Its success led to a pilot study, using concrete containment structures filled with dolomite stone and citric acid as the electron donor. A final aeration stage was used as a polishing step to remove any residual H2 S in the treated water. After 22 pore volumes, reducing conditions were established that enabled the removal of > 95% of Cr. Furthermore, there was >99% of Cr >90% of nitrate removed once the system reached a steady state. Lyon conducted a subsequent pilot study to remediate Se and nitrates from groundwater [106]. Due to its proprietary nature, designs and lessons were discussed solely. As such, the design intercepted the flow from Peter Canyon Channel and redirected it to a subsurface GBR geomembrane cell through the media bed using gravitational flow. An ORP gradient removed both Se and nitrate simultaneously to the ppb level using nitrate- and Se-reducing bacteria. Once treated, the water entered an aeration polishing step to remove residual H2 S. The aeration step also increased dissolved oxygen levels to appropriate levels before its environmental release. A soluble carbon

Materials

Supportive media: gravel, dolomite Organics: citric acid, methanol, ethanol

Nutrients: bone-meal, manure, liquid fertilizers Organics: mulch, hay, leaf litter Supportive media: limestone

Targets

Cr6+ , Se4+

Se4+ , NO3

Reactor

Gravel bed reactors

Anaerobic column reactors

References

• Pit-lake water used as inoculum source for selenium- [108, 109] and sulfate-reducing bacteria • Simple flow-through design placed in pit-lakes • Treatment performance resilient to temperature fluctuations • Se4+ removal dominates SO4 −2 removal • NO3 reduces selenium removal • Used as a proof-of-concepts for SRF and GBR

• Utilizes same principles as SRFs within a simple [103, 104, 106, 107] bioreactor • Packed gravel/dolomite beds act as substrate for biofilm formation • Gravel allows for sufficient hydraulic conductivity • Gravel bed should have high AMD loading: surface area ratio to allow for nutrients to be transferred adequately for SRB growth • Biofilms create protective microenvironments for microbes • Difficult to maintain anaerobic conditions • Backwashing is required if biofilms grow thick • Organics added to enable anaerobic conditions and facilitate dissimilatory metal reduction • Treated water is polished with aeration to remove H2 S • Additional removal mechanisms: (1) Apatite II addition for phosphate precipitation of Pb, Zn, Cd; (2) acidic pH shifts for As adsorption (e.g., CO2 ); (3) alkaline pH shifts for divalent adsorption

Mechanisms/highlights

Table 4.3 Summary of gravel bed reactors (GBRs) and proof-of-concept anaerobic column reactors’ targets, conventional materials used, removal mechanisms, and research highlights

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source maintained anaerobic conditions as an electron donor (carbon source unidentified). Furthermore, backwash and recycle systems were required in this upscale operation to reduce excess biomass in the GBR and handle heavy rainfall events, respectively. Lastly, it costs nearly USD 2.2 million for two years of operation with construction costs. Although several details are missing to make fair comparisons to other technologies, it does give a rough estimation of installation costs for GBR designs. Moreover, additional pilot tests by GeoSyntec™ in California and West Virginia, USA, were used to treat Se-contaminated waters. Both operations maintained Se levels at < 5 µg/L for over six years. While their design reduced land capita requirements compared to CWs (2200 m3 vs. 80,000 m3 ). Similarly, a pilot test remediated As-contaminated waters in a Cement Plan in the USA. By diffusing CO2 in the matrix, pH levels dropped, and As levels returned to lower levels. Thus, GBRs have shown promise in As, Cr, Se, and nitrate removal from bench-scale to pilot-scale studies. Nonetheless, future studies should assess its efficacy for a broader range of contaminants (e.g., Cu, Zn) using sulfate reduction pathways.

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Chapter 5

Recommendations and Challenges

Abstract This chapter delves into each passive treatment (i.e., source control, inpit treatments, saturated rock fills, permeable reactive barrier, gravel bed reactors, and constructed wetlands) discussing their most promising remediation amendments/mode of operation as well as their challenges under different metal drainage conditions. Keywords Source control · Reactive barrier · Constructed wetlands · Saturated rock fills · Gravel bed reactor · Passive

5.1 Prevention of Mine Drainage Production Construction techniques are implemented first during mining operations to mitigate mine drainage production. They work by strategically packing waste rock to mitigate preferential air flows and water infiltration and, therefore, oxidants’ exposure (i.e., air and water) to reactive sulfide minerals. Successful designs use multiple compacted benches with short lifts (4–6 m) slanted exteriorly with an 18° maximum slope. Push dumping is required when creating these lifts to reduce rubble zones and the interspersion of coarse and fine particles. However, additional source control is needed since construction techniques cannot prevent mine drainage production. Alternative source controls seal or blend waste rock pile to eliminate its exposure to oxidants (i.e., Fe3+ , O2 , H2 O). From mentioned studies, physical barriers are a viable option if it is not an arid region, as saturation conditions are required to prevent cracking. Nonetheless, sludge mixed with silty soil or cement are cheap and effective amendment mixtures. Shallow water tables should be further applied, if possible, to maintain saturation conditions. When dealing with arid regions, inorganic coating methods are well-suited. Specifically, lime kiln dust and limestone are cheap solutions with tested long-term stability. Choices between these options depend on costs, availability, and the reactivity and classification of the treated waste rock. Recommendations are limited to the outlined suggestions since long-term tests (> two years) have demonstrated © The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 C. Chidiac et al., Passive Treatments for Mine Drainage, SpringerBriefs in Applied Sciences and Technology, https://doi.org/10.1007/978-3-031-32049-1_5

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their efficacy without high expense. Nonetheless, this method has only been proven effective with moderately reactive waste rock. CME is the next promising source control tactic when considering highly-reactive waste rock. However, it is still in early development, and its long-term efficacy remains unknown.

5.2 On-Site Treatment of Mine Drainage If source control fails and mine drainage production persists, on-site treatment is the next mode of action. There are several viable passive treatments to use on-site or in-situ to remediate mine drainage if waters exhibit relatively low flows (< 150 L/s) and acidity levels (< 800 mg CaCO3 /L). These include in-pit treatments (e.g., alkaline, biological, immobilization, mixed), SRFs, GBRs, CWs, and PRBs (discussed in the subsequent section). There are numerous considerations when selecting the appropriate remediation approach, including mine drainage composition, available area, and area/land type. Due to the complexity, this chapter presents each design regarding situations in which they will thrive and fail in remediating mine drainage. Figure 5.1 illustrates the process considerations and applicable technologies concerning mine drainage composition, and Table 5.1 summarizes each of their advantages, challenges, and targets. In brief, alkaline solutions can treat pH-sensitive metals (e.g., Fe, Al, Cu, Pb), biological solutions can treat metals sensitive to sulfide precipitation (e.g., Zn, Cd, Cu, Pb) or that susceptible to bioreduction reactions (e.g., Se, Cr, U). High alkaline and adsorptive mechanisms can treat Co and Ni since they require high pH levels (> 8–9) to undergo removal and are not sensitive to metal sulfide precipitation. Although not mentioned previously, settling basins and in-pit settling are in the recommendation flow sheet. Like that seen for water treatment facilities, settling basins and in-pit settling involves the direction of a contaminated flow into a high surface area structure. Once infiltrated into a high surface area structure, suspended solids separate from the water by particulate settling. Induced particle settling occurs from a velocity reduction upon entering a large surface area. Specifically, the velocity reduction of the particles reduces drag forces and allows particles to settle within the basin or pit by dominant gravitational forces. Overall, settling basins are effective when metals can precipitate or co-precipitate. Therefore, it is suitable for pH-sensitive metals Fe and Al at near circumneutral pH levels (pH ≥ 4) and metals that can be co-precipitated under low to moderate concentrations (e.g., Cu, Pb) or minimal concentrations (e.g., Co, Ni, Zn, Cd). The exact values for these metal concentration ranges depend on the mine drainage’s chemistry and the design of the basin or pit. Nonetheless, hydraulic surface loading generally ranges between 0.02–0.04 L/s/m2 [1, 2].

Fig. 5.1 Flow chart of recommended passive treatments for different target metals, pH levels, and flow regimes in mine drainage. Color gradients depict the metal concentration the technology can be applied for. Exact numerical ranges depend on the drainages’ chemistry. 1 Biological treatments should be used with caution when used to treat divalent metals and Se/Cr simultaneously. 2 SSCW is only effective against Ni and cannot be used for Co. + Permeable reactive barriers (PRB) have the potential for higher flows (≤ 120 L/s)

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• • • • • • • • • •

Permeable reactive barriers (PRB)

Low energy requirements and operating costs Reduces treatment times Reduces contaminant exposure via treatment subsurface Occurs subsurface, more robust under colder climates Potential for higher flow rates (120 L/s) Containment of entire plume Use of local media reduces transportation costs Typically have long service lives (~ 10–30 years) Can remove diverse contaminants using multiple barriers Reduced environmental impact compared to conventional pump and treat options (continuous wall with waste reactive material) • ZVI can remove H2 S produced from SRB

Advantages

Technology • Susceptible to passivation • Regeneration of spent material produces secondary pollution • Loss of hydraulic conductivities, reactivity of material, and retention time of contaminants • Gas production (CH4 , H2 S) reduces hydraulic conductivity • Replacement of exhausted materials incurs high operational costs • Breakthroughs difficult to predict • Limited to shallow depths (< 50 ft) due to construction limitations • Requires temporal and depth characterization of site’s hydraulics • Impermeable barrier must intercept plume and control deeper flows • Freeze/thawing causes particulate agglomeration, loss of reactivity, and retention times • ZVI and caissons reduce the environmentally “friendliness”

Challenges and limitations

Table 5.1 Passive treatment technologies with their respective advantages, challenges and limitations, and effective targets

(continued)

Biological: Zn, Cd, Cu, Pb, Fe, Se Alkaline: Co, Ni, Fe, Zn, Cd Adsorptive: Cu, Cd, Zn, Pb, Se, Co, Ni Reactive: Cu, Cd, Ni, Zn, Pb, Se

Effective targets

122 5 Recommendations and Challenges

In-Pit treatments

No pumping requirements Low cost associated with waste materials Eliminate trucking cost with on-site amendments Adsorptive amendments mitigate sludge production Contains contamination at source; mitigates further contamination in connecting streams • MgCl2 amendments have the potential to produce ore-grade hydrotalcite • In-situ bioreactors proliferate diverse microbes due to O2 gradient (capable of treating more diverse contaminants)

Advantages

• • • • •

Technology

Table 5.1 (continued) Challenges and limitations • Alkaline amendments create secondary pollution via sludge production • Susceptible to metal leaching upon drying/rewetting events and freeze/thawing events • Winter climates reduce biological treatments’ efficacy • Secondary pollution upon adsorbent regeneration • Requires continual dosing of amendments • Incurs higher operating costs • Nutrient amendments can cause eutrophication • Difficult to predict amendment lifetimes • Requires large depth to surface area ratios

Effective targets

(continued)

Biological: Zn, Cd, Cu, Pb, Fe, Se Alkaline: Co, Ni, Fe, Zn, Cd Adsorptive: Cu, Cd, Zn, Pb, Se, Co, Ni (low concentrations) Mixed: all targets

5.2 On-Site Treatment of Mine Drainage 123

Advantages

• Low cost • Shown to be cheaper than activated treatments under low flows (< 0.02 m3 s−1 ) and low acidities (< 300 mg L−1 ) • Environmentally friendly • Macrophytes create protective microenvironments for SRB; making CW more robust under harsher conditions

Technology

Constructed wetlands

Table 5.1 (continued) Challenges and limitations • Uncertainty in the lifetime of complex substrates • Complex substrates yield low hydraulic conductivities • Unpredictable metal seepage/removal under freeze/thawing cycles • Colder climates reduce microbial activity • Winter months reduce residence times; reduce metal removal via sorption, precipitation, and biodegradation • Metal accumulation in plant shoots hazardous for surrounding wildlife • Metal precipitation can cause matrix clogging • Lose economic advantage under high flow rates and acidity levels • Higher removals require larger surface areas • Heavy rainfall events cause metal leaching and reduce metal removal

Effective targets Cu, Cd, Zn, Ni, Pb, Fe

(continued)

124 5 Recommendations and Challenges

Advantages

Saturated rock fills/gravel bed • Easier to construct in comparison to conventional treatment • Easy operation reactors • Capacity to treat large volumes of water • Low energy requirements • Small environmental footprint • Lower capital and operating costs compared to conventional treatments • Biofilms formed on waste rock protect microbes from harsh environmental conditions

Technology

Table 5.1 (continued) Challenges and limitations

Effective targets

• Requires long hydraulic retention Se, NO3 , Cr times • Requires hydraulic control • Requires large surface areas • Requires full understanding of the systems’ stoichiometry, kinetics, and performance with different organic amendments • Lack of data available for metal sulfide removal • Hydraulic conductivity affected by waste rock porosity • Presence of selenium and nitrate can reduce metal sulfide removal • Requires a polishing step to remove H2 S • Difficult to maintain anaerobic conditions • Aerobic conditions create thick biofilms and reduce the systems’ hydraulic conductivity • Requires backwashing if biofilms get too thick (continued)

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Advantages

• Prevents contamination of surrounding environment • Only method to prevent mine drainage production • Self-healing coatings may provide long-term suppression of acid mine drainage production

Technology

Source control

Table 5.1 (continued) Challenges and limitations

Effective targets

• Co-disposal with alkaline material N/A can increase metals’ solubilities • Coatings often destabilize under acidic conditions • Organosilane coatings and dry covers are susceptible to microcracks and defects • Organic covers can dissolve iron precipitates via reduction pathways and susceptible to biodegradation • Water covers unsuitable in dry regions and only slows down oxidation • Design of cover systems must be site-specific and must withstand changing climate conditions • Silicate/PO4 coatings require H2 O2 , unpractical for large-scale applications • Inorganic coatings demonstrated little selectivity for sulfide mineral • Phosphate-based passivation layers may cause eutrophication • Silane-based coatings exhibit low selectivity for sulfidic minerals • Long-term stability of carrier-microencapsulation is unknown • Catechol’s cost hinders large-scale carrier microencapsulation use

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5.2.1 Saturated Rock Fill (SRF) and Gravel Bed Reactor (GBR) Recommendations SRFs are ideal in-situ bioreactors that reduce costs while treating a range of contaminants. Specifically, SRFs create saturation conditions within inter-pore structures of waste rock to limit oxygen diffusion. Saturation conditions reduce carbon amendment requirements and suppression of AMD generation while creating ideal environments for SRB and SeRB. It further reduces costs by utilizing mining pit sediment as a source of indigenous microbes and using waste rock as a substrate to support biofilm. Although minimal reports explore SRB efficacy in the SRF design, in-pit treatments have demonstrated the principle of organic amendments applied in situ for SRB-driven treatments. Therefore, SRFs should further elevate biological in-pit treatments by establishing saturation conditions and further supports biofilm formation. SRFs should replace biological in-pit treatments when structurally possible but are limited to mine drainages that are slightly acidic (pH ≥ 4) and are unsuitable when Co and Ni are contaminants of concern. SRFs construction further require the strategical placement of waste rock to capture rainfall and snow; hydraulic connections to SRF downstream; long HRTs that are sufficient under changing seasons and flow regimes; and an understanding of present reactions and their stoichiometry. Understanding the dynamic behavior of waste rock porosity from carbon amendments and biofilm formation is imperative for predicting and forecasting hydraulics and HRTs in SRFs. There are also several limitations to using SRFs. First, there are no studies that quantify sulfate removal. It is unclear whether the developed ORP in SRF enables the simultaneous removal of sulfates, Se, and nitrates. Understanding the potential for simultaneous sulfate removal is crucial since Se and nitrates can potentially inhibit sulfate reduction and limit the ability of SRFs to treat divalent metals (e.g., Zn). Thus, research should assess the feasibility of SRFs to remove all three contaminants using an ORP gradient. While SRFs should not treat sulfates in the presence of nitrates and Se until this is understood. Second, SRF designs are not well suited to treat highly acidic mine drainage as this adversely affects environmental conditions for SRB and SeRB. In these instances, alkaline in-pit treatments are viable options or pH buffers may be added to SRFs if economically feasible. There is minimal published literature on GBRs due to their proprietary nature. However, their removal mechanism and design emulate that of SRFs. GBRs, therefore, can remediate AMD at the same capacity as SRFs while providing a pump and treatment method. They become an attractive solution if SRFs are ill-suited to the geometry of the contaminated pit or if plumes develop. GBRs will require a buffer under acidic conditions (pH < 4) and are not suitable for metals insensitive to metal sulfide precipitation (e.g., Co, Ni). Furthermore, the utilization of SRFs and GBRs for mine drainage treatment is limited to flow rates ≤ 150 L/s since higher flow rates do not allow for sufficient retention times for the biological removal of metals and alkalinity generation.

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5.2.2 In-Pit Treatment Recommendations In-pit treatments are attractive remediation strategies when drainage is inaccessible or the surrounding area is non-excavatable. It further simplifies treatments since amendments can be tilled directly in the pit, its design does not require hydraulic control and is less sensitive to environmental conditions (e.g., heavy rainfall). The range of contaminants it can treat is broad since there are no limitations on the number of amendments supplied. Alkaline in-pit treatments are viable when treating contaminated drainage with high acidity and become particularly effective with pH-sensitive metals (e.g., Fe and Al). Under low metal loadings, the precipitation of pH-sensitive metals can remove neighboring, soluble metals by co-precipitating at near-neutral pH levels. However, under higher loading of divalent metals, pH levels must exceed 8 to effectively remove pH-insensitive metals Zn, Ni, Cd, Co, and Cu. If desired, pH-selective precipitation can also aid in metal recovery. The benchmark alkaline amendments are limestone and quicklime. Limestone provides a longer-term and more stable pH increase, while quicklime can induce high pH levels quickly to precipitate hard-to-treat metals such as Co and Ni. Nonetheless, these amendments are costly and damaging to the environment. Thus, industrial and agricultural waste materials are preferred as cheap, environmentally friendly amendments and should be selected based on their proximity to the mining site. Examples include recycled concrete aggregates, cement kiln dust, oxygen furnace slags, eggshells, biochar, coal refuse, and earth minerals. Their desired characteristics include high carbonate content, high dissolution rates, and high porosity. However, industrial materials can leach metals into the mine drainage and should be pre-tested. Alternatively, immobilization in-pit treatments are an attractive solution for pH-insensitive metals and are well suited for drainage containing low concentrations of metals and moderate acidity. Drainages with high acidity may have worsened treatment due to H+ competition with metal cations unless the adsorbents are alkaline with neutralization abilities. Desired adsorbent attributes include diverse functional sites; small pore diameters; high exposed surface area; and regeneration capacity. Regeneration capabilities without requiring expensive or environmentally damaging chemicals are paramount in preventing secondary pollution and remain a critical challenge in their use. Promising waste-derived adsorbents include fly ashderived zeolites; Ca-alginate; biochar; dried sludge; and green waste. A limitation of adsorbents is their inefficiency in sulfate removal. Although adsorbents with high point-zero-charge have the potential for sulfate remediation, minimal sulfate removal occurs from competing ions in complex mine drainage. Biological in-pit treatments can remediate drainage with high sulfate concentrations when SRF designs are infeasible. Maintaining optimal growth conditions for SRB is critical in the biological remediation of mine drainage. Thus, amendments dosages should ensure pH levels and COD:S ratio are at 5–8 and 1.5, respectively. If there are native SRBs capable of effectively precipitating metals, then biostimulation should be used with waste carbon-rich materials. The carbon-rich materials must be a blend of labile and recalcitrant organic fractions. Simple, labile carbon

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sources include sewage, and molasses, while recalcitrant organics include green waste, compost, and sugar bagasse. Carbon amendment selection should be according to cost, availability, and proximity to the affected mine site. Additionally, algal growth via N/P additions can create additional metal scavengers and carbon sources for SRB growth. Considerations include if there are N/P sources near the site, confidence that primary production will support the SRB communities, and long-term effectiveness. Mixed in-pit treatments are desirable when remediating hard-to-treat mine drainage. As such, alkaline amendments can remove metals as precipitates and achieve ideal conditions for SRB growth. Carbon sources can act both as adsorbents and as an energy source for SRB. If alkalinity and SRB cannot remove certain metals, one can add adsorbents with a high affinity for that metal. Additionally, adsorbents work synergistically with microbes by adsorbing metal contaminants nearby and providing surfaces for microbial attachment and biofilm formation. Overall, considerations for in-pit treatments are their economics, the optimal ratios of materials, and their longevity. In-lab technical analysis and economic assessments for potential waste materials can yield information on their feasibility before application. Resource recovery, such as ore-grade metals, is possible with alkaline amendments and can offset treatment costs if feasible.

5.2.3 Constructed Wetland Recommendations CWs are viable in environments with large surrounding areas that do not exhibit heavy rainfall and cold climates. Their implementation in such conditions is undesirable since they are susceptible to wash-out events, sensitive toward preferential flows and reduced biological treatments upon freezing, and require large land capita. Furthermore, CWs are economically advantageous under low flow (< 20 L/s) and low acid loadings (< 300 mg/L) and has experienced good removal (> 95%) for hard-totreat metals: Cd, Cu, Zn, Cr, Ni, and Cu. Its remediation of Se has yet to be elucidated and thus is not recommended. Nonetheless, CWs offer a long-term treatment solution capable of removing contaminants via biological reactions, phytoextraction, adsorption, cation exchange, precipitation, and co-precipitation. Its design should consider the extremes of the surrounding area’s climate and potential droughts and floods [3]. The implemented design should not be over-engineered and instead should emulate a natural wetland [3]. Although mentioned as an on-site treatment, CWs can be used on-site as a pump and treat method, or to be used to intercept developed plumes. Vertical subsurface flow regimes are optimal since they reduce land requirements and improve the contact time between the mine drainage and the CW’s constituents. Contact time is crucial in ensuring the dissolution of alkaline media, such as seashell grit and limestone, which neutralizes mine drainage and creates suitable SRB growth conditions. The CW should promote saturation conditions by strategically placing macrophytes and creating tight hydraulic control. Identified suitable macrophytes for AMD are Typha domingensis, Phragmite austalis, Typha angustifolia, Desmostachya binnata, Saccharum bengalense, Typha latifolia, Iris

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pseudoacorus L, Juncus effusus, Schoenoplectus acutus, and Phalaris arunginacea. Generally, persistent emergent macrophytes are suitable for CWs, where the particular macrophyte selected should be according to their metal affinity and BCF and TF values [3]. If economically feasible, citric acid can reduce Fe- and Mn-plaques around macrophytes. Furthermore, successful CWs use blended organic and cellulosic waste materials to proliferate and sustain SRBs and maintain sufficient hydraulic conductivities. Successfully tested organic waste materials are cow manure, domestic wastewater, plant litter broth, and activated sludge; successfully tested cellulosic waste materials are bamboo chips, walnut shells, and woodchips. Mixing different carbon sources increase the diversity and richness of microbial communities in CWs, which may improve mine drainage treatment. Depending on cost, citric acid amendments proliferate acidophilic heterotrophs and reduce Fe-oxidizing microbes. These organics should maintain a COD:S ration of 1.5 to prevent the growth of methanogens. Lastly, there will be start-up time associated with CWs due to their dependency on SRB remediation and their associated slow growth (≥ 20 d). Start-ups should consider alkaline amendments to buffer treatment or to introduce a recycle stream during these initial times.

5.3 Capturing and Treating Mine Drainage Seepage If on-site treatments fail and mine drainage seepage occurs, PRBs, GBRs, and CWs can intercept plumes for treatment. PRBs will be the main focus since the previous sections discussed CWs and GBRs. PRBs require an excavatable capture zone with surrounding material with relatively low hydraulic conductivities and an impermeable lower bound. If satisfied, the selection of the PRB design will depend on the mine drainage composition. For instance, if the mine drainage exhibits low acidity and metal concentrations, adsorptive PRBs are an optimal choice to ease operation and reduce costs. Adsorbent PRBs require low HRTs while promising adsorbent amendments are Apatite and thermally activated wood ash which are effective for Cd, Cu, Co, Ni, and Hg. Other promising materials are bentonite, volcanic ash soil, red soil, and waste foundry sand. Nonetheless, only premature findings are available for the latter amendments, and it is essential to conduct long-term column tests before implementation. If mine drainage has low-to-moderate acidity and low-to-high metal concentrations, biological PRBs with an SDES design become suitable. SDES designs protect microbes from high metal concentrations without exhibiting sulfate diffusion limitations. Nonetheless, gypsum supplementation to the reactive layer can prevent sulfate limitations. Since SDES treatment is SRB-driven, it will be more effective than alkaline treatments in treating pH-insensitive metals (e.g., Cu) and create more dense precipitates that ease their separation and dewatering characteristics. SRB can also remediate Se by using it as their terminal electron acceptor. Small dosages of sulfated or Fe-dosed ZVI can accentuate removal mechanisms required for Se remediation.

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However, biological/SDES PRBs are unsuitable for Co and Ni since they are insensitive to metal sulfide precipitation. Biological and SDES PRBs should have long HRTs to sustain SRB growth. The HRT should also gradually increase until alkalinity generation is sufficient. Organics that have shown promise are whole-cell algae, lipid-extracted algae, sugar cane bagasse, and sludge. SDES designs have successfully used blends of pinus radiata compost, biodigested sludge from local wastewater treatment plants, and calcium sulfate. Nonetheless, this design does produce H2 S, a compound that is both toxic and corrosive. ZVI can be dosed into the PRBs modestly to scrub residual H2 S before effluent release. If the drainage instead experiences high acidity and low metal concentrations, alkaline materials with strong sorption capacities are recommended in a DAS configuration using sand or cellulosic waste (e.g., wood chips) as the dispersant material. It should also operate under a low HRT to reduce desorption potential. Promising alkaline materials are cement kiln dust (CKD) and calcite-enriched clam and mussel shells since they both exhibit neutralization capabilities and high surface areas that ease metal adsorption. Nonetheless, these materials cannot remove sulfates. If this is a treatment goal, a subsequent biological PRB or a stratified layer comprised of organics and SRB may be implemented to remove sulfates. A hybrid model with a low HRT is proposed if mine drainage is highly acidic and contains high metal concentrations. Specifically, a DAS model that incorporates additional organics to promote SRB proliferation and activity can be viable in such a case. In the latter scenario, low concentrations are advantageous as it allows adsorptive removal to dominate. However, once high concentrations persist, precipitation will then begin to dominate. Hydroxide precipitation is subpar at removing these persistent divalent metals and at such high concentrations since co-precipitation will likely not be sufficient. Thus, inducing sulfide metal precipitation with hydroxide precipitation may allow for the co-removal of both pH-sensitive and insensitive metals. This hybrid model is also promising for waters with high sulfate concentration. However, if Co and Ni persist in the drainage, SRB will not be sufficient. Instead, high alkaline levels of pH > 8 can remove these metals from the solution. Nonetheless, in all scenarios, material selection is imperative. For PRB designs to succeed, materials need to be selected by their cost, availability, lifetime (≥ five years), and hydraulic conductivity. PRBs can offer a long-term solution (~ 30 years) to mine drainage treatment if hydraulic conductivities can be maintained, thereby emphasizing the importance of implementing DAS or SDES design.

References 1. Sludge Dewatering and Disposal. https://ebrary.net/183832/environment/sludge_dewatering_d isposal (Online) 2. M. Guillou, Water storage and sedimentation basins: concept and sizing (2013) 3. U. Epa, O. Wetlands, A handbook of constructed wetlands a guide to creating wetlands for

Chapter 6

Outlook

Abstract This chapter delves into each passive treatment (i.e., source control, inpit treatments, saturated rock fills, permeable reactive barrier, gravel bed reactors, and constructed wetlands) future outlook based on recent research trends and their identified learning gaps. Keywords Source control · Reactive barrier · Constructed wetlands · Saturated rock fills · Gravel bed reactor · Passive

6.1 Source Control Source control requires ongoing research since it is the only method that suppresses and eliminates AMD production. For instance, studies on construction techniques have yet to demonstrate the efficacy of compacting saturation layers and its effect on particle binding. Another strategy to improve particle-binding may involve the integration of cheap, polymeric adhesives with tailings. These approaches may reduce void volumes, and in turn, oxidant exposure in tailings. Additionally, there has been a systematic lack of blending and co-disposal amendments’ characterization regarding their hydrophobicity and micro properties (e.g., porosity, contact angle), and potential chemical modifications. Understanding macro and micro properties will better enable researchers to predict performance and synergism among materials. Additionally, chemical modifications via polymeric additions may create non-wetting surfaces and allow for the fine-tuning of blending material properties. These research directions will enable the optimization of blending and co-disposal remedies. Another promising avenue in microencapsulation includes nanoparticle embedment to reduce void volumes and oxidant exposure. However, there has been a lack of understanding nanoparticle effects regarding their size, shape, and degree of crosslinking, therefore warranting further investigation and optimization.

© The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 C. Chidiac et al., Passive Treatments for Mine Drainage, SpringerBriefs in Applied Sciences and Technology, https://doi.org/10.1007/978-3-031-32049-1_6

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Another exciting research direction is the idea of self-healing passivation layers. Although it was recently explored by Li et al. by using HNT and BTA to improve PropS-SH silane coatings, there are quite a few limitations. These limitations are the complex chemistry that requires unfeasible conditions (i.e., high temperatures, solvents), its lack of cost competitiveness, and its unknown long-term suppression. To try and reduce the complexity and potentially cost, a recommended direction is to emulate the idea of self-healing concrete, a current research interest for building science [1]. Self-healing concrete could be a promising direction for AMD since concrete is comprised of CaCO3 is the main component in concrete and has already demonstrated its effectiveness in treating mine drainage. Cement, or possibly its waste derivatives, can therefore be blended in a capping mechanism with low permeable amendments, such as sludge, to create a surface cap that resists exposure to oxidants. Within these concrete materials, there can be a loading of pH-sensitive capsules containing the same healing agents proven for self-healing concrete (e.g., epoxy resin, acrylics) or bacterial agents (i.e., urea-bacteria) [1, 2]. This approach uses cheaper materials with simple chemistry that can continuously generate alkalinity and induce microencapsulation. Microencapsulation via alkalinity generation and precipitation induction has already proven to work in the long-term (> 2 y). Perhaps this can push the technology further and create an even more attractive solution to AMD production. Another important research direction is the identification of catechol replacements in CME techniques. As mentioned, catechol improves the selectivity of CME techniques, thereby reducing dosage requirements and improving coverage. Although catechol holds its advantages, its cost remains the major limitation. There is a lack of research on finding alternatives to catechol. It is recommended that future research look at cheap materials which exhibit a catechol structure. Examples include poly(dopamine) and its derivatives, as each of these contain catechol moieties and are highly available. Lastly, there are inconsistencies with testing passivation layer’s efficacies. Specifically, it is difficult to objectively compare the efficiencies among different source control tactics because of the diversity in the waste rock reactivity. To truly assess the capacity of these passivation layers, researchers should systematically test their amending layer with different waste rock reactivity to assess their functionality without creating biases. Indeed, one study found that passivation strategies targeting pyrite oxidation rates are insufficient in treated the acid released from secondary mineralization [3]. While pyrite is the most abundantly studied mineral for passivation techniques, this highlights the importance of understanding their performance limitations under broader implications.

6.2 Permeable Reactive Barriers (PRBs) PRBs have been thoroughly investigated and have shown great promise in their treatment abilities. Nonetheless, research should pivot from the conventional amendment ZVI to more sustainable materials to improve this technology’s sustainability and economics. Additionally, future research should focus on diffusion active PRB

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designs which use intercalating layers of conductive and reactive layers. This design allows the usage of sustainable waste materials locally available to mining sites. The waste materials can include industrial and organic wastes since both have shown promise in the DAPRB configuration. However, biological methods are preferred unless the contaminated plume exhibits high acidity, in which case a blend of alkaline and organics are beneficial. Furthermore, research in this direction needs to look at the efficacy of the layers’ thicknesses and the number of layers utilized. Once optimized, the designs require testing under long durations (≥ 2 years). If deemed suitable for larger scales, it would be beneficial to conduct life cycle assessments on the new design and compare against ZVI PRBs. Furthermore, although there has been extensive research on the efficacy of biological PRBs, no studies have examined the associated microbial framework. Specifically, there have been no studies on the metagenomics of biological PRBs or phylogenetics and how they may shift from different designs and amendments. Understanding relative shifts in gene expression and microbial proliferation will yield information necessary for their optimization.

6.3 Constructed Wetlands CW research has thoroughly investigated the efficacy of different waste organics, flow regimes, and the efficacy of different macrophytes. Future research should implement organics as blends to increase the richness and diversity of the microbial community. There is a further lack of understanding of the lifetime of these organics, which is a notable cause for CW failure. Thus, research should better estimate these lifetimes. Since SRB sustaining CW remediation exhibits slow growth rates, the start-up of CW often takes several weeks (≥ 20 d). As demonstrated in another study, alkaline media can sustain treatments during start-up times. Although seashell grits are suitable, research should investigate alternative porous, alkaline waste materials. Finding alternative alkaline media could aid in technology sustainability as different areas will have varying waste material availability. For instance, biochar is a promising amendment used to treat wastewater [4]. Granular biochar derived from alkaline materials offers supporting media with high specific surface area and ion-exchange capacity with diverse and abundant functional groups. Thus, biochar can remove metal(loid)s via electrostatic attraction, ion/ligand exchange, complexation, precipitation, and pi-binding interactions. It further has been found capable of reducing Cr and Hg from solution. Additionally, granular biochar enables sufficient hydraulic flows and creates the potential for recycling waste biomass and regenerating valuable resources (e.g., biodiesel, biohydrogen). Thus, future research should investigate granular biochar derived from different waste materials as CW media for AMD treatment. Finally, the performance of CWs has not been optimized due to the lack of adequate monitoring and long testing times [5]. Thus, future studies should run longer with tight monitoring to decipher optimal conditions for CW.

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6.4 Saturated Rock Fills/Gravel Bed Reactor SRFs and GBRs are promising passive technologies that can remediate a range of mine drainage contaminants. However, to further improve their sustainability and cost, future research should pivot away from employing soluble organics and toward locally available complex waste organics, while focusing on their lifetime, impacts on waste rock porosity, and biofilm formation. Furthermore, waste materials with high amounts of nitrogen and phosphorus are of interest to induce primary algal production. Primary algal production can broaden removal routes using algae as metal scavengers. Moreover, future research should assess the efficacy of simultaneous sulfate-, nitrate-, and Se-reduction to evaluate their simultaneous removal with ORP gradients. Lastly, it is unclear where reduced metals deposit in SRFs and how this evolves with time. It is imperative to understand the stability of reduced metals within these SRFs and if there’s a potential for wash-out under cell lytic events.

6.5 In-Pit Treatments As in-pit treatments are well developed, there is a need to focus on technical and economic evaluations of waste materials. Economic comparison of different materials, waste or otherwise, relative to mining site location and AMD discharge regulations would assist industrial efforts to incorporate passive mine drainage treatment. This work should include optimization of material ratios needed for ideal treatment, dependent on water chemistry considerations. Moreover, modifying alkaline and adsorbent materials is a topic of growing interest. For instance, work has investigated the ability to adjust divalent:trivalent metal ratio in the aqueous solution to form hydrotalcite or LDH structures, which have a capacity for metal removal. Understanding these ratios in industrial mine drainage with process control considerations has been negated and is critical for their success. Similarly, fine-tuning functional groups of materials can enable higher affinities for specific contaminants. For example, modified waste-derived biochars can selectively bind to ions and provide a tool for mineral recovery in AMD. Additionally, since biochar is electrically conductive, low-energy currents may enable the desorption of metal cations for their recovery and biochar regeneration. Research should investigate the potential of electrically regenerating biochar and its ability to recover metals and other valuable products. For biological treatment, ongoing research should continue to quantify the ideal ratios of labile to long-term refractory carbon sources, taking water chemistry and native microbes into account. Long-term recalcitrant carbon sources, particularly waste materials local to mining sites, are needed to make bioremediation a longerterm solution. Additionally, identifying promising native species, alongside their

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growth requirements for biostimulation, can remove the necessity for bioaugmentation and should be investigated at local sites. Lastly, alternative passive processes for in-pit treatment, such as coagulation with chitosan, should be further developed to extend the breadth of feasible treatment options [6]. The future of this field is utilizing these treatment tools, ideally with low-impact materials such as industrial waste, to recover valuable metals through adsorption, co-precipitation, pH-selective precipitation, biological treatment, and other developed processes.

References 1. S. Guo, S. Chidiac, Self-healing concrete: a critical review 2. N. de Belie, E. Gruyaert, A. Al-Tabbaa et al., A review of self-healing concrete for damage management of structures. Adv. Mater. Interfaces 5, 1–28 (2018). https://doi.org/10.1002/admi. 201800074 3. R. Fan, G. Qian, Y. Li et al., Evolution of pyrite oxidation from a 10-year kinetic leach study: implications for secondary mineralisation in acid mine drainage control. Chem. Geol. 588, 120653 (2022). https://doi.org/10.1016/J.CHEMGEO.2021.120653 4. Application of biochar as an innovative substrate in constructed wetlands/biofilters for wastewater treatment: performance and ecological benefits. Elsevier Enhanced Reader. https://reader. elsevier.com/reader/sd/pii/S0959652621003760?token=FD9B371D7A7AD32EC74E7FA334F 7EA14979FAD504E839169B119D7EE58F93D15B07A9CD7EF71B673E3203C6541F29 9A9&originRegion=us-east-1&originCreation=20211228214045. Accessed 27 Dec 2021 5. U. Epa, Wetlands O, A handbook of constructed wetlands a guide to creating wetlands for 6. C.O.A. Turingan, N.E.S. Lintag, C.I.L. Ngan et al., Post-Treatment of acid mine drainage using laboratory-extracted common mushroom chitosan. Chem. Eng. Technol. 44, 1962–1969 (2021). https://doi.org/10.1002/CEAT.202100220