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Biological Metal Recovery from Wastewaters [190]
 9783031737640

Table of contents :
Cover
Half Title
Advances in Biochemical Engineering/Biotechnology: Volume 190
Biological Metal Recovery from Wastewaters
Copyright
Preface
Contents
Sulphidogenic Bioprocesses for Acid Mine Water Treatment and Selective Recovery of Arsenic and Metals
1. Arsenic and Metals in Mine Waters
1.1 Needs for Metals and Arsenic
1.2 The Mining Heritage
1.3 Active Mining Industry
1.4 Current Arsenic Treatment Technologies
2. Principle of Metals and Arsenic Recovery as Bio-sulphides
3. Actors of the Bioreactions
3.1 Actors in Natural Environments at Low pH
3.2 Sulfate-reducing communities in AMD processing operations
4. Treatment Performances
4.1 Synthetic Waters
4.2 Real Acid Mine Water
4.3 Process Water
4.4 Integrated Processes
5. Perspectives
References
Biological Iron Removal and Recovery from Water and Wastewater
1. Introduction
2. Principles of Biological Iron Removal and Recovery from Water and Wastewater
2.1 Aerobic Biological Iron Rmoval and Recovery
2.2 Anaerobic Biological Iron Removal and Recovery
3. Bioreactor Types for Biological Iron Removal and Recovery
3.1 Suspended Cell Bioreactors
3.1.1 Attached Cell Bioreactors
3.2 Examples of Bioreactor Studies on Biological Iron Removal and Recovery
4. Pilot- and Full-Scale Applications of Biological Iron Removal and Recovery from Wastewater
4.1 Aerobic Processes
4.2 Anaerobic Processes
5. Patents for Biological Iron Removal and Recovery
6. Comparison of Biological Iron Removal and Recovery Methods
7. Beneficial Uses of Iron Products Removed from Wastewater
8. Conclusions
References
Aluminum Biorecovery from Wastewaters
1. Introduction
1.1 Global Importance of Aluminum
1.2 Aluminum Geochemistry
1.3 Aluminum Recovery
1.4 Aims and Scope of the Chapter
2. Aluminum Recovery from Different Solid Wastes
3. Aluminum Recovery from Domestic Wastewaters
4. Current Practices of Metal Removal from Acid Mine Waters
5. Aluminum Recovery from Acidic Waters by Sulfate-Reducing Bacteria
6. Limitations and Future Directions
References
Precious Metal Recovery from Wastewater Using Bio-Based Techniques
1. Introduction
2. Precious Metal-Containing Wastewater Streams
3. Technologies for Biorecovery of Precious Metals from Waste Solutions
3.1 Biosorption
3.2 Bioaccumulation, Biomineralization, and Bioreduction
3.3 Bioelectrochemical Recovery
3.4 Membrane Biofilm Reactors
4. Value-Added Products Recovery
5. Comparative Assessment of Precious Metal Biorecovery Technologies and Traditional Technologies
6. Concluding Remarks
References
Microalgae: A Biological Tool for Removal and Recovery of Potentially Toxic Elements in Wastewater Treatment Photobioreactors
1. Introduction
2. Source and Hazardous Effects of Potentially Toxic Elements
3. Drawback of Conventional Treatments for Potentially Toxic Elements Removal
4. Microalgae as Biosorbent
5. Mechanisms of Potentially Toxic Elements Removal by Microalgae
6. Factors Affecting Potentially Toxic Elements Removal by Microalgae
7. Potentially Toxic Elements Uptake by Microalgae
8. Combined Potentially Toxic Elements and Nutrient Removal from Real WW
9. Conclusions
References
Phytoextraction Options
1. Introduction
2. Sources of Metal(loid)s Across Different Types of Wastewaters
3. General Approaches in Plant-Based Wastewater Treatment
3.1 Free Water Surface CWs (FWS-CWs)
3.2 Subsurface Flow CWs (SSF-CWs)
3.2.1 Horizontal Subsurface Flow CWS
3.2.2 Vertical Subsurface Flow CWs
3.3 Intensified and Modified CWs
4. Phytoextraction of Metal(loid)s
4.1 Selection of Plant Species
4.2 Element Accumulation in Plants
4.3 Application of Phytoextraction in CWs
4.3.1 Rooted Emergent Hydrophytes
4.3.2 Floating Hydrophytes
4.3.3 Submerged Hydrophytes
4.3.4 Phycoextraction
4.3.5 Enhanced Phytoextraction
5. Element Recovery from Biomass
5.1 Element Enrichment Techniques
5.1.1 Compaction or Pressing
5.1.2 Biological Enrichment or Microbial Treatment
5.1.3 Thermal Enrichment Methods
5.2 Extraction of Elements from Enriched Plant Biomass
6. Conclusion
References

Citation preview

Advances in Biochemical Engineering/Biotechnology 190 Series Editor: Roland Ulber

Sabrina Hedrich Oliver Wiche Editors



Biological Metal Recovery from Wastewaters

190 Advances in Biochemical Engineering/Biotechnology Series Editor Roland Ulber, Kaiserslautern, Germany Editorial Board Members Thomas Scheper, Hannover, Germany Shimshon Belkin, Jerusalem, Israel Thomas Bley, Dresden, Germany Jörg Bohlmann, Vancouver, Canada Man Bock Gu, Seoul, Korea (Republic of) Wei Shou Hu, Minneapolis, USA Bo Mattiasson, Lund, Sweden Lisbeth Olsson, Göteborg, Sweden Harald Seitz, Potsdam, Germany Ana Catarina Silva, Porto, Portugal An-Ping Zeng, Hamburg, China Jian-Jiang Zhong, Shanghai, Minhang, China Weichang Zhou, Shanghai, China Katja Bühler, Leipzig, Germany Antonina Lavrentieva, Hannover, Germany

Aims and Scope This book series reviews current trends in modern biotechnology and biochemical engineering. Its aim is to cover all aspects of these interdisciplinary disciplines, where knowledge, methods and expertise are required from chemistry, biochemistry, microbiology, molecular biology, chemical engineering and computer science. This series upports and advances the UN’s sustainable development goals, mainly SDG7 (Affordable and Clean Energy), SDG3 (Good Health and Well Being), and SDG13 (Climate Change). Volumes are organized topically and provide a comprehensive discussion of developments in the field over the past 3–5 years. The series also discusses new discoveries and applications. Special volumes are dedicated to selected topics which focus on new biotechnological products and new processes for their synthesis and purification. In general, volumes are edited by well-known guest editors. The series editor and publisher will, however, always be pleased to receive suggestions and supplementary information. Manuscripts are accepted in English. In references, Advances in Biochemical Engineering/Biotechnology is abbreviated as Adv. Biochem. Engin./Biotechnol. and cited as a journal.

Sabrina Hedrich • Oliver Wiche Editors

Biological Metal Recovery from Wastewaters

With contributions by M. V. Alegre  B. Antolín Puebla  F. Battaglia-Brunet  S. Bolado Rodríguez  C. Falagán  P. A. García Encina  J. Jacob  E. Janneck  C. Joulian  A. H. Kaksonen  A. Kumar  J. Meier  I. Nancucheo  S. Ndlovu  A. Samarska  J. Sánchez-España  O. Wiche

Editors Sabrina Hedrich Institute of Biosciences TU Bergakademie Freiberg Freiberg, Germany

Oliver Wiche Applied Geoecology Group Zittau/Görlitz University of Applied Sciences Zittau, Germany

ISSN 0724-6145 ISSN 1616-8542 (electronic) Advances in Biochemical Engineering/Biotechnology ISBN 978-3-031-73764-0 ISBN 978-3-031-73765-7 (eBook) https://doi.org/10.1007/978-3-031-73765-7 © The Editor(s) (if applicable) and 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 If disposing of this product, please recycle the paper.

Preface

Water comprises the most essential resource on earth, which is indispensable for life in all ecosystems and drives economic development. With global population growth and industrial development, the demand for clean water rises, leading to the depletion of freshwater sources and causing environmental concerns. In the face of global water scarcity, efficient water recycling is crucial to ensure access to safe and clean water, protect public health, and preserve environmental ecosystems. Water recycling involves treating wastewater to remove contaminants and supporting sustainable water management by enabling the reuse and recycling of water resources for various purposes, such as agricultural irrigation, industrial processes, and even potable water supplies. Without effective treatment, waterborne diseases can proliferate, and pollutants can harm aquatic life and disrupt natural habitats. Industrial wastewaters also often contain economically valuable elements in concentrations relevant to recycling, some of which are considered critical raw materials. Therefore, recycling wastewaters can have several benefits, such as (1) providing clean water for various purposes, (2) protecting the environment, and simultaneously (3) recovering raw materials. While several physical and chemical wastewater treatment technologies, including filtration, coagulation, and sedimentation, have successfully been applied to the majority of wastewater streams, biological water treatment has gained significant attention. Especially for complex or dilute streams, biological methods have proven very efficient. Biological wastewater treatment harnesses natural processes and the metabolic activity of various organisms, including bacteria, fungi, algae, and plants. It uses natural processes to degrade or remove pollutants, often with minimal energy inputs and without harmful chemicals. Integrating biological treatment into conventional systems has proven to be an effective solution for managing wastewater, improving water quality, enhancing overall treatment efficiency and resource recovery, and ensuring compliance with environmental standards. Biological wastewater treatment processes offer a sustainable, efficient, and costeffective alternative to traditional chemical and physical methods. They are particularly effective in removing biodegradable organic matter, nutrients, and certain inorganic compounds, including metals and metalloids. Its relevance has grown with v

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the increasing emphasis on sustainable and environmentally responsible practices in wastewater management. One of the key advantages of biological treatment is its ability to handle a wide range of waste streams under varying environmental conditions. Despite these advantages, challenges remain in optimizing biological treatment processes for large-scale applications, particularly in dealing with non-biodegradable pollutants and ensuring consistent performance under fluctuating conditions. Ongoing research focuses on enhancing the efficiency, scalability, and economic viability of these systems to meet the growing demand for sustainable wastewater treatment solutions. This book provides an overview of various biological wastewater treatment methods, evaluating their advantages, limitations, and applicability in different settings. The various technologies focus on the various contaminants in industrial wastewaters with a special focus on the recovery of valuable raw materials using different setups. Chapter “Sulphidogenic Bioprocesses for Acid Mine Water Treatment and Selective Recovery of Arsenic and Metals” explores the recovery of iron from industrial wastewaters using iron-oxidizing bacteria in full-scale bioreactor operations. Chapter “Biological Iron Recovery from Waste Waters” focuses on aluminum biorecovery, addressing challenges such as process control and economic feasibility. Chapter “Aluminium Recovery from Waste Streams” examines the recovery of precious metals and other value-added products from wastewater. Chapter “Precious Metal Recovery from Wastewater Using Bio-Based Techniques” addresses microalgae-based technologies for the bioremediation of wastewaters containing copper, zinc, and arsenic. Chapter “Microalgae: A Biological Tool for Removal and Recovery of Potentially Toxic Elements in Wastewater Treatment Photobioreactors” introduces plant-based techniques for the phytoextraction of economically valuable elements from wastewater. The chapter discusses the dual benefits of pollutant removal and biomass utilization for bioenergy and raw material recovery. Chapter “Phytoextraction Options” focuses on the removal of arsenic from wastewaters by sulfate-reducing bioreactors via selective precipitation strategies. Freiberg, Germany Zittau, Germany

Sabrina Hedrich Oliver Wiche

Contents

Sulphidogenic Bioprocesses for Acid Mine Water Treatment and Selective Recovery of Arsenic and Metals . . . . . . . . . . . . . . . . . . . . Fabienne Battaglia-Brunet, Ivan Nancucheo, Jérôme Jacob, and Catherine Joulian

1

Biological Iron Removal and Recovery from Water and Wastewater . . . Anna Henriikka Kaksonen and Eberhard Janneck

31

Aluminum Biorecovery from Wastewaters . . . . . . . . . . . . . . . . . . . . . . . Javier Sánchez-España, Carmen Falagán, and Jutta Meier

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Precious Metal Recovery from Wastewater Using Bio-Based Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 119 Sehliselo Ndlovu and Anil Kumar Microalgae: A Biological Tool for Removal and Recovery of Potentially Toxic Elements in Wastewater Treatment Photobioreactors . . . . . . . . . . 147 Beatriz Antolín Puebla, Marisol Vega Alegre, Silvia Bolado Rodríguez, and Pedro A. García Encina Phytoextraction Options . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 181 Alla Samarska and Oliver Wiche

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Adv Biochem Eng Biotechnol (2024) 190: 1–30 https://doi.org/10.1007/10_2024_264 © The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 Published online: 28 August 2024

Sulphidogenic Bioprocesses for Acid Mine Water Treatment and Selective Recovery of Arsenic and Metals Fabienne Battaglia-Brunet, Ivan Nancucheo, Jérôme Jacob, and Catherine Joulian

Contents 1 Arsenic and Metals in Mine Waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.1 Needs for Metals and Arsenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2 The Mining Heritage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.3 Active Mining Industry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4 Current Arsenic Treatment Technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 Principle of Metals and Arsenic Recovery as Bio-sulphides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3 Actors of the Bioreactions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1 Actors in Natural Environments at Low pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2 Sulfate-reducing communities in AMD processing operations . . . . . . . . . . . . . . . . . . . . . . . 4 Treatment Performances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1 Synthetic Waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2 Real Acid Mine Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3 Process Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.4 Integrated Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 Perspectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

2 2 3 4 5 5 7 8 11 12 12 15 18 21 23 25

Abstract Human communities need water and mineral resources, the supply of which requires the implementation of recycling and saving strategies. Both closed and active mining sites could beneficiate of the implementation of nature-based F. Battaglia-Brunet (✉) CNRS, BRGM, ISTO, UMR 7327, Université d’Orléans, Orleans, France e-mail: [email protected] I. Nancucheo Facultad de Ingeniería, Arquitectura y Diseño, Universidad San Sebastián, Concepción, Chile J. Jacob and C. Joulian BRGM, Orleans, France

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solutions, including bioreactors involving sulphate-reducing prokaryotes (SRP), in order to separate and recover arsenic (As) and metals from aqueous stream while producing clean water. Selective precipitation strategies can be designed based on the selection of microbial communities adapted to the pH conditions, generally acidic, and to available low-cost electron donors. Laboratory batch and continuous experiments must be implemented for each type of mine water in order to determine the optimal flow-sheet in which As could be precipitated as sulphides (orpiment or realgar), inside the bioreactor or offline, through stripping of biologically produced hydrogen sulphides (H2S). The respective concentrations and proportions of As and metals and the initial acid mine drainage pH are key parameters that will influence the feasibility of efficient selective precipitation. SRP-based bioreactors could be combined with complementary treatment steps in optimised mine water management solutions that will minimise the production of As-contaminated end-solid waste. Keywords Acid mine water, Arsenic, Bioremediation, Metals, Precipitation, Sulphate-reducing procaryotes

1 Arsenic and Metals in Mine Waters 1.1

Needs for Metals and Arsenic

The global population growth combined with economic development and the technological innovation made necessary by the energy transition imply the rise and changes of the need in mineral resources. Such needs associated with decarbonised energy, in particular solar photovoltaic, wind turbines and electric vehicles, have been particularly documented in terms of resource constraints [1]. The criticality of metals and metalloids [2] is evaluated through a methodology that analyses simultaneously all influencing factors, including geological abundance, geopolitical concentration of ore deposits, mining and metallurgical technologies, economic policies, governmental instability risk, potential for substitution and environmental regulations. Arsenic (As), used in metallurgy and semi-conductors, appeared in the list of critical raw materials for the European Union due to increased economic importance caused by relatively higher increase in added value of application metals [3]. Arsenic is currently necessary for zinc (Zn) production (Zn electrowinning), glass making, manufacture of chemicals, alloys and electronics. This context highlights the need to develop and improve technologies supporting the sustainable exploitation of new geological resources, the minimisation of future wastes produced by ore processing and the valorisation of mine waste and the associated acid mine water discussed in this chapter.

Sulphidogenic Bioprocesses for Acid Mine Water Treatment and. . .

1.2

3

The Mining Heritage

Huge amounts of waste are produced when the geological materials are treated to extract the mining resources. Waste is produced at each stage of the mining process, but the most problematic and voluminous are coarse grain overburden and sub-economic ore produced during extraction and fine-grained tailings from mineral concentration processes such as flotation [4]. The extraction of metals like copper (Cu), nickel (Ni) or gold (Au) can produce up to 1,000 tonnes of waste for 1 kg of pure metal. These wastes often contain sulphide minerals that are vulnerable to oxidation when exposed to air and rainwater. These processes can lead to the generation of acid mine drainage (AMD), an acidic runoff containing elevated concentrations of metals (iron (Fe), aluminium, manganese, Cu, Ni, lead (Pb). . .) and metalloids such as As [5]. Their impacts can be felt near and far from their place of origin. When these effluents spill into the rivers, they degrade the habitat and water quality considerably, to the point of creating an environment incompatible with all forms of aquatic life. Mining has caused major environmental problems such as acid and metallic pollution of wetlands, aquatic environments and aquifers on a global scale. The magnitude of the impact was estimated in 1989, and reached 19,300 km of rivers and streams, and 72,000 ha of lakes and reservoirs [6] that had been significantly affected by mine water discharge. Downstream, these effluents can cause major disturbances for the use of the water. This is especially true for drinking water, leisure activities or irrigation, leading to loss of ecosystem services. Apart from the economic issues, these conflicting uses can cause serious problems to environmental and human health. Mine water streams are likely to contain As in various concentrations because of the solubility of arsenite (AsIII) and arsenate (AsV) in a large interval of pH. Arsenic is present in more than 156 mineral phases associated with sulphidic ores [7] found in closed mining sites. Both polymetallic and coal mine sites are potential sources of mine water containing As. Arsenic minerals are found in ore bodies historically exploited for many valuable elements including Au, Ag, Cu, Zn, tungsten [8], lead (Pb) [9], antimony (Sb) [10] but was also exploited for As itself [11, 12]. As an example of extremely high As concentration in AMD, Nordstrom and Alpers [13] mentioned the Iron Mountain mining district (California) that strongly impacted the Sacramento river, with about 300 tonnes of dissolved Cd, Cu and Zn discharged from the site. This site has seen successive historical mining of ores, from 1860s to 1982 for the recovery of exploitation of Ag, Au, cerium (Ce), Fe, Zn and pyrite for sulphuric acid production. One of the ancient mines of this district, Richmond Mine, hosted underground water presenting extremely high concentrations in As (850 mg/L), copper (9,800 mg/L), Zn (49,300 mg/L), Fe (68,100 mg/L). The occurrence of As with other valuable metals or metalloids in AMD is not anecdotal, as illustrated by many other examples, with Tharsis in the Spanish pyrite belt [14] containing 150 mg/L Cu, 450 mg/L Zn and 10 mg/L As; Carnoulès mine in France with close to 1,000 mg/L Fe and 100 mg/L As [15]; Komsomolsk gold mine in Siberia [16] with up to 4 g/L of As and 140 mg/L Zn, and 30 mg/L Cu; Morro Velho mine in Brazil (7.3 mg/L As,

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Table 1 Characteristics of a typical acid mine water containing arsenic (Carnoulès site, France, from Casiot et al. 2003 [15], named source of acid water (station 1), range of variation of parameters during 1 year of monitoring pH 2.7–3.45

SO4 mg/L 1,440–5,470

As mg/L 150–360

AsIII mg/L 48–280

AsV mg/L 0–92

Fe mg/L 950–2,110

FeII mg/L 560–2,110

1.8 mg/L Cu, [17]; Jugan mine in Malaysia (2.9 mg/L As, 1 mg/L Cu and 4.7 mg/L Zn [17]; Iron Duke mine in Zimbabwe (72 mg/L As, 20 mg/L Cu, 55 mg/L Zn and 24 mg/L Ni [17]. The characteristics of a typical acid mine water containing arsenic are given in Table 1. This worldwide historical mining heritage induces issues for human health, affecting water resources, environment and agricultural lands. Transforming this problematic consequence of past mining practices into secondary resources potential is a current challenge. Bioprocesses based on sulphate reduction belong to the pool of promising strategies currently explored, tested and optimised in order to selectively recover metals and metalloids from wastewater streams.

1.3

Active Mining Industry

Arsenic is a common impurity found in processed base metals ore such as copper, lead, zinc and is often associated with gold and silver-bearing ores. It is present in ores such as enargite (Cu3AsS4), luzonite (Cu3AsS4), tennantite (Cu12As4S13), tetrahedrite (Cu12Sb4S13) and realgar (As4S4) among others [18]. The depletion of large gold deposits has led to the processing of refractory ores associated with sulphides containing As which generates issues for mining companies since roasting, which is used for ~20% of major global gold producers, produces unfavourable toxic gas-containing As causing environmental and processing constraints [19]. On the other hand, the demand for base metals such as copper will keep growing, and consumption by emerging technologies such as wind generators and solar photovoltaics as well as the increasing world population have raised concerns regarding sustainable demand-supply balance in the long term [20]. Copper deposits in Chile have been identified as one of the important sources worldwide; however, Chilean high grades of copper oxides that are processed by hydrometallurgy are depleting and conversely, copper sulphides have increased as deposits are mined to a greater depth. Importantly, current and future copper concentrates face a critical challenge due to the presence of As which lowers the ore’s economic value [21–23]. This causes environmental and processing constraints, as most smelters will not process concentrates containing more than 0.5% arsenic, due to increasingly stringent worldwide environmental regulations [22, 24]. So far, 10% of the world supply of copper concentrates contains As above the penalty level (0.2%). In the case of Chile, the most common arsenic-bearing copper sulfosalt is enargite, which is difficult for concentrators to separate from only copper minerals due to similar floatability

Sulphidogenic Bioprocesses for Acid Mine Water Treatment and. . .

5

properties. Besides roasting and smelting, chalcopyrite-enargite concentrates release arsenic gases [18]. Therefore, the need for additional processing of these concentrates is a trending topic to dissolve Cu and As with the subsequent removal of arsenic from wastewater from the mining industries.

1.4

Current Arsenic Treatment Technologies

Currently, Ecometales which is a wholly-owned subsidiary of CODELCO, Chile, one of the world’s largest copper producers, is developing several technologies including leaching copper concentrates through high-pressure and high temperatures to obtain a stream of copper- and arsenic-rich solution that must be stabilised for long-term storage [23]. The removal of As from acidic aqueous conditions is usually achieved in mineral processing and metallurgical operations by lime neutralisation, ferrihydrite and scorodite formation, which are subsequently disposed by dumping over large areas where it presents an environmental risk due to possible leaching of arsenic from the sludge. The most common strategy used by mining companies for arsenic abatement focuses on the formation of AsV, a less toxic chemical species, providing improved environmental outcomes compared with AsIII. When AsIII is present, the first step to remove arsenic involves the oxidation of AsIII [25] and subsequent AsV is removed using chemical lime neutralisation or scorodite formation at temperatures below 100°C, though one commercial biotechnological company (Paques, the Netherlands) is promoting the formation of biogenic scorodite [26]. In this approach, biological oxidation of ferrous iron using thermoacidophiles controls arsenic precipitation in one step at pH 1.2 and 70°C. Alternatively, a chemical sulphidisation process has been implemented at the Saganoseki copper smelter (Japan) to remove AsIII from acidic liquors as As2S3. The process has advantages such as (i) smaller volume of sludge production compared with lime treatment, and (ii) higher water recovery, which allows for the treatment of arsenicfree acidic effluents to recover other valuable metals. A similar process was implemented by BQE water; however, there is a high cost of reagents, particularly sodium hydrosulfide (NaSH) [25].

2 Principle of Metals and Arsenic Recovery as Bio-sulphides Sulphate reduction is a biological reaction of sulphate respiration from the point of view of sulphate-reducing prokaryotes (SRP), and at the same time a “chemical” reaction of oxidation of an electron donor (molecule which is oxidised by the reaction) by an electron acceptor (molecule reduced during the reaction) which is the sulphate, SO42-. Electron donors can be either organic molecules (acids,

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alcohols, fatty acids, alkanes, etc.) or dihydrogen (H2) commonly called “hydrogen”. The following equations correspond to the result of the chain of bioreactions involved in the oxidation of acetate, glycerol or dihydrogen by SRP as electron donors resulting in the reduction of sulphate to produce hydrogen sulphide: Example 1: acetate as electron donor SO42- + 4 CH3COO- + 3 H+ = H2S + 2 H2O + 2CO2. Example 2: glycerol as electron donor C3H8O3 + 0.75 SO42- + 1.5 H+ = C2H4O2 + H2CO3 + 0.75 H2S + H2O. Example 3: dihydrogen as electron donor SO42- + 4 H2 + 2 H+ = H2S + 4 H2O. These processes consume H+ thus will promote the increase of pH in active sulphate-reducing systems. The produced hydrogen sulphide reacts with As and other metalloids and metals inducing the precipitation of sulphide bio-minerals, these reactions depend on pH [27]. Arsenic sulphide precipitation is favoured in acidic pH conditions [28], together with CuS, PbS, CdS and HgS, whereas NiS then ZnS precipitate in moderately acidic conditions. Iron sulphide FeS precipitation occurs at higher pH (pH > 5, [29], followed by MnS when pH increases above pH 8. As an example, the following successive reactions were observed while pH increased in batch experiments with SRP growing in a real AMD [29]: Precipitation of As: 2 H3AsO3 + 3 H2S = As2S3 (am) + 6 H2O Precipitation of Zn: Zn2+ + H2S = ZnS(am) + 2 H+ Precipitation of Fe: Fe2+ + H2S = FeS(am) + 2 H+ Arsenic sulphide (orpiment As2S3 or realgar AsS) precipitation occurs in a small window of geochemical and physico-chemical conditions, including low pH and small (limited) bio-production of H2S. High H2S production and high pH favour the formation of soluble thioarsenate complexes. In addition, the precipitation of As sulphide implies reaction between AsIII and H2S (and not AsV). Thus, when AsV is present, the reduction into AsIII is a first necessary step, that may be performed either chemically with H2S as electron donor or biologically in presence of active AsVreducing bacteria. Batch experiments of simple mixture of AsIII or AsV with H2S-containing SRP cultures were performed in order to identify the optimal conditions of As sulphide precipitation (Fig. 1, [30]). At low pH and in H2S limiting conditions, AsV reduction was the dominant reaction, consuming H2S more rapidly than As2S3 precipitation. When the pH increased, the two reactions occur simultaneously and As2S3 level reaches maximum values. At pH higher than 6, the two reactions were very slow. The optimum pH for As precipitation was in the range pH 2 to pH 4.5. The influence of pH on the precipitation of different types of sulphides can be exploited for the selective recovery of elements. Indeed, the control of pH conditions drives the types of sulphides that can be formed and recovered through a liquid/solid separation process.

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Fig. 1 Influence of pH on As sulphide precipitation in batch [30]. (a) AsIII precipitation and final pH according to initial pH, filtration a few minutes after mixing; (b) As precipitation starting with AsV according to initial pH at different times of reaction; (c) As precipitation according to final pH after different times of reaction. Conditions: initial AsIII or AsV concentrations 1.33 mM = 100 mg/ L; initial dissolved S-II 2.6 mM = 83 mg/L (brought as SRP culture supernatant). Reactions performed in glass tubes, 5 mL of mixture, with mineral oil at the interface to avoid H2S loss for the AsV experiment. The initial pH of As solutions was adjusted with NaOH or H2SO4 solutions

3 Actors of the Bioreactions SRP are well-known prokaryotes widespread in the environment including mine impacted ones. They are phylogenetically diverse and have been classified into five lineages within the Bacteria (Deltaproteobacteria, Clostridia, Nitrospirae, Thermodesulfobacteria and Thermodesulfobiaceae), and two phyla within the Archaea (Crenarchaeota and Euryarchaeota) [31]. Despite, they all share the ability to produce hydrogen sulphide from the dissimilatory reduction of sulphate during anaerobic respiration, preferably using hydrogen and simple organic substrates as electron acceptors, such as alcohols (ethanol, methanol), low molecular weight organic carbons (lactate, propionate, formate, pyruvate), but also hydrogen and some more complex organic materials as mentioned in Sect. 2 [31, 32]. The produced biogenic sulphides are used in treatments to precipitate metals and metalloids

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cations as metallic sulphides, and SRP are recognised actors of the bioremediation of acid mine waters [33, 34].

3.1

Actors in Natural Environments at Low pH

SRP have long been considered to be inhibited at acidic pH. Even though Tuttle et al. [35] reported in 1969 on dissimilatory sulphate reduction in acid mine water, for the next 20 years SRP were rather thought to occur in microniches with higher pH as there was no evidence of SRP growth at pH below 5 [36]. The first report of an SRP grown in acidic conditions was in 1993 by Johnson et al. [37] who enriched, from acid streamers of disused mines in North Wales, endospore-forming bacteria resembling to Desulfotomaculum, with a lower pH limit at 2.9. Since then, several SRP have been isolated from acidic mining effluents and sediments (Table 2). They all show optimum growth and activity in the pH range of 3 to 6 and are thus defined as moderate acidophiles, by contrast with extreme acidophiles growing optimally below pH 3 [38]. Most of the isolated strains are endospore-forming bacteria of the genus Desulfosporosinus, a close relative of the genus Desulfotomaculum. Three novel Desulfosporosinus species have been described as moderate acidophiles with a low pH limit below 4; namely D. acidiphilus isolated from an acid mining effluent decantation pond, D. acididurans isolated from acidic river sediments, D. metallidurans isolated from a microbial mat from a gold mine tailing dam [39–41]. Other Desulfosporosinus-related strains were isolated, from acidic sediments of the Tinto River [42] and from gold mine tailings in Kuzbass in Siberia [43]. The diversity of moderate acidophilic SRP is not restricted to Desulfosporosinus. Desulfovibrio sp. strain C.1 was retrieved from AMD sediments and showed same rate of sulphate reduction at pH 5.5 and 7 [44]. Two Desulfovibrio strains, isolated at pH 5.3 and pH 5.7 from a 1-L bioreactor inoculated with an SRP enrichment at pH 5.5 and operated with a pH gradient, were reported to have low pH limits at 2.8 (strain VK) and 3.5 (strain ED) for growth [45]. More recently, strain INE, isolated from acidic sediments from the Tinto River, has been proposed as a novel species of a new genus of moderately acidophilic SRP, Acididesulfobacillus acetoxydans, clustering separately from the Desulfosporosinus SRP genus of the Firmicutes phylum [46]. Strain CL4 (formerly proposed as “Desulfobacillus acidavidus”), [47] and strain CEB3 [48] should also be considered as strains of A. acetoxydans [46]. Strain INE has the capacity to completely oxidise organic substrates to CO2. This property distinguishes it from other moderate acidophilic strains known so far, since Desulfosporosinus and Desulfovibrio species are generally described as incomplete oxidisers producing acetate from organic substrates [31]. A limited growth on acetate has only been reported for D. metallidurans [41]. Recently, two strains affiliated to the genus Solidesulfovibrio were isolated from AMD sediments; however, a medium at neutral pH was used and there is no information about their ability to grow at acidic pHs [49].

Firmicutes

Firmicutes

Firmicutes

DeltaProteobacteria DeltaProteobacteria

Desulfosporosinus metallidurans

Desulfosporosinus sp. I2

Acididesulfobacillus acetoxydans

Desulfovibrio sp. C.1

Desulfovibrio sp. VK

Firmicutes

Phylum or class Firmicutes

Desulfosporosinus acididurans

Name of isolate Desulfosporosinus acidophilus

nd nd

2.8 – ca. 7 (5.7)

25–42°C (30°C)

4–32°C (22–28°C)

4–37°C (28°C)

15–40°C (30°C)

nd

3.8–6.5 (5.0)

1.7–7.0

4.0–7.0 (5.5)

3.8–7.0 (5.5)

pH range (optimum) 3.6–5.5 (5.2)

Temperature range (optimum) (°C) 25–40°C (30°C)

Table 2 Main properties of moderate acidophilic SRP isolates

nd

H2, formate, lactate, butyrate, fumarate, malate, pyruvate, glycerol, methanol, ethanol, yeast extract, xylose, glucose, fructose H2, lactate, pyruvate, malate, formate, propionate, ethanol, glycerol, glucose, fructose, sucrose, peptone, tryptone Lactate, pyruvate, malate, citrate, succinate, fumarate, butyrate, ethanol, glycerol, butanol, formate, palmitate, peptone H2, formate, acetate, glycolate, pyruvate, lactate, malate, fumarate, butyrate, succinate, L-cysteine, methanol, ethanol, glycerol, 1- and 2-propanol, yeast extract, xylose, glucose, fructose, maltose, sucrose, raffinose nd

Electron donors for sulphate reduction H2, lactate, pyruvate, glycerol, glucose, fructose

Acidic sediments from AMD (Brazil) Tailings of an abandoned Pb–Zn mine (Russia)

Tinto River sediments (Spain)

Tailing dam at an abandoned gold mine in Siberia (Russia) Oxidised layers of gold mine tailings in Kuzbass (SW Siberia)

Source of microorganism Acid mining effluent decantation pond (Chessyles-Mines, France) Sediment from White river and Tinto river (Spain)

(continued)

[45]

[44]

[46]

[43]

[41]

[40]

Reference [38]

Sulphidogenic Bioprocesses for Acid Mine Water Treatment and. . . 9

Thermodesulfobium narugense Thermodesulfobium acidiphilum Desulfothermobacter acidiphilus

Name of isolate Desulfovibrio sp. ED

Table 2 (continued)

Clostridia

Clostridia

Phylum or class DeltaProteobacteria Clostridia

pH range (optimum) 3.5 – ca. 7 (6.6) 4.0–6.5 (5.5–6.0) 3.7–6.5, (4.8–5.0) 2.9–6.5 (4.5) 37–65°C (50–55°C) 37–65°C (55°C) 42–70°C (55°C)

Temperature range (optimum) (°C) nd

H2, formate

H2, formate

H2, formate

Electron donors for sulphate reduction nd

Geothermally heated soil (Russia) Terrestrial hot spring (Russia)

Source of microorganism Tailings of an abandoned Pb–Zn mine (Russia) Naruto hot spring (Japan)

[52]

[51]

[50]

Reference [45]

10 F. Battaglia-Brunet et al.

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The moderate acidophilic SRP isolated from mining sites are mesophilic microorganisms thriving optimally at temperatures around 30°C, or just below for Desulfosporosinus sp. I2. However, isolates with a moderate thermophilic lifestyle have been isolated from geothermal sites, they belong to the genera Thermodesulfobium (T. narugense and T. acidophilum) [50, 51] and Desulfothermobacter (D. acidiphilus) [52].

3.2

Sulfate-reducing communities in AMD processing operations

In bioreactors set-up for sulphidogenic treatment of acidic waters rich in metals or even arsenic, the moderate acidophilic SRP are co-occurring with non-sulphidogenic bacteria. This was the case in the lab experiments described below (see Sect. 4), initially inoculated with sulphate-reducing consortia enriched from AMD-impacted sediments and conducted with glycerol as substrate. In continuous bioreactors fed with synthetic acid water, microbial community fingerprints showed that Desulfosporosinus spp. co-occurred with the fermentative Clostridium genus known to use glycerol and other organic substrates [23, 53, 54]. In a bioreactor continuously fed with real AMD, 16S rRNA gene metabarcoding was applied for a deeper characterisation of the microbial diversity and revealed a complex community of which main genera included the sulphate-reducing Desulfosporosinus, as the only SRP in the bioreactor, together with the acetate-utilising Acidocella and fermenters such as Cellulomonas, Microbacter and Luteococcus able to use glycerol but also sugars or complex polysaccharides [55]. Complex microbial communities have also been described in sulphidogenic bioreactors filled with complex organic substrates and treating artificial AMD. Lignocellulosic wastes promoted microbial diversity in laboratory wetland bioreactors, that contained several genera of both complete oxidisers (Desulfosarcinaceae, Desulfatirhabdium, Desulfobacter, Desulforhabdus, Desulfobacca) and incomplete oxidisers (Desulfovibrio, Desulfocapsa, Desulfomonile), together with lignocellulosic compounds degraders and sulphide-oxidising bacteria [56]. Aoyagi et al. [57] performed a diversity study in a vertical flow bioreactor using rice bran as organic substrate and treating real AMD (JOGMEC process) at an abandoned mine site and showed two main phyla to accumulate, Firmicutes with the fermentative Clostridium genus, and Bacteroidetes comprising complex carbohydrates degraders, and the main SRP to be Desulfatirhabdium, with Desulfovibrio and Desulfosporosinus in lower proportions. More recently, Lin et al. [58] described a high diversity in three vertical flow ponds treating high-risk AMD in Pennsylvania over a period of 643 days, with different accompanying genera but also distinct main head SRP (Desulfovibrio or Desulforhopalus) depending on the complex organic substrate used, crab shell or compost of spent mushroom. All ponds’ communities combined, a total of 19 SRP genera were retrieved, together with fermenters and polymer degraders, the latter believed to benefit to the fermenters that would provide the simple organic substrates to the SRP.

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In all these studies, SRP are members of complex communities composed of a variety of microorganisms with metabolisms complementary to the reduction of sulphate by SRP. On site, complex organic compounds are preferred substrates to provide the needed carbon sources to sustain an efficient and durable sulphatereducing activity, which is a key parameter for a successful sulphidogenic treatment of the AMD. The implication of several distinct metabolic groups is mandatory to degrade such complex substrates into more simple ones that are in fine usable by the SRP. Also, they will be able to use the acetate often produced during organic substrates degradation by SRP which are mostly incomplete oxidisers. Such metabolic interactions have been recently revealed by metagenomic analyses that explored the functional potential of microbial communities in laboratory-scale passive bioreactors operated with a mixture of complex organic compounds to treat artificial AMD, in which substrates for SRP were constantly provided by the cellulose degraders and sugar fermenters [59]. Furthermore, when focusing on the diversity of the SRP as key drivers of the sulphidogenic process, if they generally account for a lower proportion than other groups in the complex microbial communities, diversity studies have revealed that sulphidogenic bioreactors can be colonised by many distinct SRP genera, and that several genera co-exist in a same operation. This shows that other genera than those still isolated to date and listed above (see Sect. 3.1) are involved in the treatment of AMD by sulphidogenic processes or even are capable of reducing sulphates at acidic pHs. Moreover, when the AMD to be treated contains AsV, the dissimilatory reduction of AsV to AsIII can improve the efficiency of As sulphide precipitation. The arrA gene involved in AsV respiration, whose sequence was closely related with that of Desulfosporosinus Y5 [60], was detected in a biofilm performing As sulphide precipitation in low pH conditions [53]. This arrA sequence was also retrieved in enrichments precipitating arsenic sulphides [61].

4 Treatment Performances 4.1

Synthetic Waters

The mechanisms and efficiency of selective precipitation of metals and metalloids from acidic water have often been first studied with synthetic solutions of defined and reproducible composition. This can be justified by the need to control all experimental parameters and obtain results that could be easily published thus reproduced by the scientific community. The selective precipitation of Cu (50 mM) in presence of soluble Zn (50 mM) was obtained [62] with a sulphidogenic culture of Acidithiobacillus ferrooxidans at pH 2.6, using hydrogen derived from zero-valent iron corrosion as electron donor. In a small pozzolana-filled biofilter, Battaglia-Brunet et al. [53] entirely removed 1.33 mM AsV in 40–50 h of residence time, from a synthetic solution supplied with glycerol as electron donor. The reaction occurred at pH 5. High pH and high

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sulphate-reduction efficiency are detrimental to the bio-precipitation of arsenic sulphide. This phenomenon, linked to the formation of soluble thioarsenates, was observed when glycerol was replaced by H2, inducing a sharp increase of pH in the bioreactor [53]. In the same type of bioreactor fed with glycerol [53], the shift from AsV to AsIII feeding (Fig. 2) allowed to work with an inlet solution at pH 2 (average residence time 65 h) while keeping a very efficient As removal. The decrease of pH in the feed rather decreased the residual soluble As, in accordance with the indication of the preliminary batch tests (see previous section on involved reactions). Based on the observation that As removal was favoured by low pH and moderate sulphate-reduction activity, an attempt to design a system allowing the separation of As and Zn was elaborated. The device was composed of two column biofilters in series (Fig. 3). The idea was to obtain a limited sulphate reduction in the first bioreactor, and a higher sulphate-reduction activity in the second one. An increasing sulphate-reduction activity, controlled by different substrate availability, would allow a biologically-controlled gradient of pH and consequently different steps of sulphide precipitation. The two columns were filled with mixtures of pozzolana and agar enriched in nutrients (yeast extract 0.2 g/L in the first and 2 g/L in the second column, together with 0.78 g/L of KH2PO4). The feed solution (pH 3.6) was mimicking the composition of the real mine water of Carnoulès, with 100 mg/L total As (75% AsIII and 25% AsV), 20 mg/L Zn and 900 mg/L FeII. The feed solution contained just enough glycerol for the precipitation of As (245 mg/L), and a complementary amount of concentrated solution of glycerol was injected between the 2 column for the precipitation of Zn in the second one. The establishment of a pH gradient between the two columns was observed, pH 4.5 in the outlet of the first column, and pH 5.0 to 5.5 in the outlet of the second column. However, the separation of As and Zn was not very efficient: the average values of removal rates were 18% in column 1 and 22% in column 2 for As, and 4% in column 1 and 68% in column 2 for Zn, while Fe remained in solution. These narrow results were attributed to a too high initial flow rate, the application of a lower feed flow rate when continuous feed was started could have resulted in a more robust biofilm and better efficiency of As and Zn removal. Besides biofilters that may be up-scaled as semi-passive treatment processes, other types of bioreactors have been proposed and tested at laboratory scale with synthetic solutions for the selective recovery of metals. A pH-stat continuously fed bioreactor containing bacteria immobilised on glass beads was designed to study strategies to selectively precipitate metals from synthetic mine effluents [48]. In this system, the fixed-film layer is settled in the bottom zone and covered with an upper liquid layer mixed with an impeller. The pH is adjusted by controlling the feed flow rate of acidic solution to be treated, injected below the fixed-film layer. The working pH and selectivity of precipitation between Cu and Zn were controlled by adjusting the feed pH and the inlet concentration of glycerol (electron donor). These strategies were efficient to reach good selectivity between Cu and Zn with synthetic effluent whose composition was close to that of Mynydd Parys copper mine in north Wales [48]. The same type of bioreactor was operated with a synthetic mine water (pH 2.1)

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Fig. 2 Evolution of As concentration (a) in the outlet of a continuous bioreactor (described in [53]) fed with synthetic solutions, shift from AsV to AsIII in the feed and decrease of the feed pH and effect of the feed pH (b) on the average remaining As. Unpublished data from the BIOMINE project “Biotechnology for Metal bearing materials in Europe”, 2004–2008, https://cordis.europa.eu/ project/id/500329/reporting)

with a composition similar to the Azufre River [54] containing 0.5 mM Zn and 2 mM (112 mg/L) FeII as the major soluble metals. This bioreactor allowed the complete removal of Zn at pH 4.5, with 20% of Fe removal, and the rate of Fe precipitation

Sulphidogenic Bioprocesses for Acid Mine Water Treatment and. . .

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Fig. 3 Schematic representation of the 2-column system designed for the separation of As from Zn in mine water. Residence time was close to 40 h in each column

increased to 70% with a shift of pH from 4.5 to 6. Thus, the co-precipitation or selective precipitation of Zn and Fe could be managed by fixing the pH value through different strategies, including the feed flow rate and the electron donor concentration. This pH-stat bioreactor was also implemented with a synthetic mine water at pH 2.1 to precipitate AsIII and AsV (from 2 to 150 mg/L), the internal bioreactor pH being fixed at 4.5 [23].

4.2

Real Acid Mine Water

The phenomena described with synthetic solutions in controlled conditions could also be observed with real mine water. The AMD of Carnoulès mine, presenting high As and Fe concentrations, could be sampled and stored in anaerobic and cold conditions for a long time without major geochemical changes, allowing the implementation of several laboratory studies performed in batch [29] and continuous [55, 63] conditions, with glycerol as electron donor. The real AMD, whose pH

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Fig. 4 Quantities of As and Zn precipitated according to pH in the batch experiments with real mine water [28]. B1, B2 and B3 correspond to Batch 1, Batch 2 and Batch 3 (triplicates) [29]

had been adjusted to 4, was amended with glycerol and inoculated with SRP enrichment from the site [29]. A progressive increase of pH between pH 4.0 and pH 4.5 to 5.0 was accompanied by successive precipitation of As sulphides then Zn sulphides, while most of Fe remained in solution. A fixed-film bioreactor was filled in two different layers, with pozzolana mixed with bio-compounds and agar differentially enriched in nutrients [63] and was inoculated with the same SRP enrichment. This column bioreactor was fed with the real Carnoulès AMD without any pH adjustment (pH 3.3), only amended with 250 mg/L glycerol. The differential filling compositions did not allow the selective precipitation of As and Zn in different layers of the bioreactor; however, the 300 days experiment showed the feasibility to precipitate almost entirely As and Zn from the real AMD at pH 3, while Fe remained in solution, the outlet pH being always lower than pH 5, in a semi-passive type or bioreactor. The same bioreactor was efficient to remove together As and Zn but not Fe from another AMD from the same site, less acidic and less concentrated in As (18 mg/L), Fe (300 mg/L) and Zn (4.4 mg/L), by decreasing the glycerol dosing from 250 mg/L down to 50 mg/L [63]. Coming back to the fine analysis of the batch experiment results [29], the bio-precipitation of As and Zn was neither simultaneous nor completely selective, as illustrated by Fig. 4. This representation indicates that with the real mine water of Carnoulès, when pH is lower than a value close to 4.5, As is precipitated but not Zn. At higher pH, As and Zn are precipitated together. Up to 40 mg/L As could be

Sulphidogenic Bioprocesses for Acid Mine Water Treatment and. . .

17

precipitated before the pH increase linked to biochemical reactions induced Zn precipitation. According to these results, in a mine water at pH 4 with As concentration not higher than 40 mg/L, As could be efficiently separated from Zn in SRP-based bioreactors system supplied with glycerol, such as the two successive columns presented in Sect. 4.1. In the proposed system, arsenic could be exclusively precipitated in the first column while zinc could be precipitated in the second one. If the pH of the mine water was higher than 4.5, separation of As and Zn would not be possible. Moreover, the selective precipitation of Zn, excluding precipitation of Fe, implies that pH remains lower than 5 according to batch results. A methodology based on a good knowledge of the mine water and small experiment scale data could help to check the feasibility of selective recovery of As and metals. As an example, the columns 1 and 2 described in Sect. 4.1 were fed with real mine water (without pH adjustment) after the step of synthetic solution feeding previously reported. Based on all results of this experiment, a coarse correlation could be drawn between the outlet pH of columns and the total quantities of sulphur that precipitated with As and Zn (Fig. 5a). The pH corresponding to the beginning of Zn precipitation is close to 4.5, corresponding to around 1 mM of precipitated S thus 50 mg/L of As. It is slightly higher than the value obtained in batch (40 mg/L, Fig. 4), but this can be explained because in the batch experiment the initial pH had been adjusted to 4 whereas the columns were fed at the actual AMD pH of 2.85. In this system, pH 5 should be reached when around 1.7 mM of sulphur is precipitated as sulphides (Fig. 5). Thus, it is possible to roughly calculate the amount of Zn that could be selectively precipitated in presence of Fe, according to the concentration of As from 0 to 50 mg/L (Fig. 5b). The quantity of zinc that may be theoretically selectively precipitated from a mine water of type of Carnoulès (initial pH 3.8) would roughly vary from 110 to 45 mg/L. According to these coarse estimates, the selective recovery of zinc would be feasible from a mine water containing As and Fe. The Zn concentration in the particular mine water of Carnoulès site is too low, compared with As concentration, to consider a Zn recovery process based only on glycerol dosing in successive bioreactors; however, the results of batch and continuous experiment suggest that this recovery might be considered from a mine water presenting higher Zn and lower As concentrations. These theoretical speculations about the feasibility of selective recovery of valuable metals in As-containing mine water must be completed by a remark concerning the possibility to recover pure As sulphide. As a fact, many metal sulphides can precipitate at low pH (Cu, Cd, Zn. . .), together with metalloids. Indeed, the co-precipitation of As and antimony (Sb) was observed in a bioreactor fed with real Carnoulès mine water spiked with increasing concentrations of Sb [55]. As a consequence, the recovery of a pure biogenic As sulphide would imply either to use an AMD of suitable properties (low pH and As concentration strongly higher than Sb, Pb, Cd, Cu, Hg) or to find a way to separate As from other elements that precipitate at low pH.

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Fig. 5 Correlation between the outlet pH and the total amount of sulphur precipitated with the sum of As and Zn in columns 1 and 2 described in 4.1, fed with synthetic then real Carnoulès AMD (a) and theoretical quantity of Zn that may be selectively separated from Fe and As, by precipitation of ZnS in a mine water (initial pH 3.8) of Carnoulès AMD type (b), in presence of a range of As concentrations (from 0 to 50 mg/L)

4.3

Process Water

Sulphidogenic process based on microbial sulphate reduction has been demonstrated as an attractive approach to removing AsV from AMD, groundwater and metallurgy

Sulphidogenic Bioprocesses for Acid Mine Water Treatment and. . .

19

industry effluents. However, most of these studies have been carried out with neutrophilic SRP [64]. While some studies have shown the removal of AsV using acid-tolerant SRP [29, 53], to the best of our knowledge, little attention has been given to treating acidic mine process water containing high concentrations of AsV, particularly by using acidophilic SRP. Hernández et al. [23] have recently reported the removal of highly toxic AsV by using an acidophilic sulphate-reducing bioreactor with a mixed culture including an acidophilic sulphate reducer (Desulfosporosinus acididurans). In brief, the bioreactor was fed with an acidic synthetic water (pH 2.1) for up to 420 days, with varying concentrations of AsIII from 2 to 150 ppm. The results of XRD analysis showed the formation of arsenic sulphides, such as realgar which is described as a more stable sulphide rather than orpiment [29]. The removal of AsIII as sulphide is often unstable and unsatisfactory due to the formation of soluble formation of thioarsenite by-products, particularly at neutral pH [53, 65]. Under these circumstances, the sulphidogenic bioreactors would be disadvantageous since this process at low pH is a proton consumption reaction, increasing the pH of the effluent at values above 5 [66]. To minimise the formation of thioarsenic complexes, it is pertinent to use pH control of the bioreactor liquor to maintain the pH of effluent at values below 5. By using pH-stat continuous cultivation mode, the processed liquor can be fixed at pH values lower than 5.0; and the flow rates of influent liquors are dictated by the pH differential between the influent liquor and the set-operating pH of the bioreactor. In addition, under this configuration, any change in the hydraulic retention time (HRT) is modulated by the SRP activity, and therefore it is possible to determine the performance of the operation, particularly under long-term experimental conditions. Single-stage bioreactors allow lower investment and less working area, but they limit the recovery of “clean” metal sulphides from mine water since the precipitation occurs inside the bioreactor with the co-occurrence of bacterial cells [67]. The offline system avoids this complication and allows a more controlled mineralisation, facilitating the selective recovery of metal(oid)s sulphide particles due to the different solubility products of each sulphide [48, 68, 69]. Final products (H2S and CO2) with biosulphidogenesis under acidic conditions using acidophilic-acid tolerant SRP have the advantage that are readily transferred through gas streams from bioreactors to another reactor vessel for offline precipitation [70, 71]. By carrying out this strategy, the precipitation of AsIII as sulphide was evaluated by varying the pH of the AsIII solution in the offline vessel (Fig. 6). A continuous upflow biofilm bed sulphidogenic bioreactor (2 L working volume) based on similar modules described by González et al. [54] was commissioned for this experiment housing a mixed culture that was originally obtained from an anaerobic sediment of the Azufre River in the Andean Altiplano, Chile. The bioreactor was fed 5 mM glycerol, 1,000 mg/L sulphate with autotrophic basic salts [72], trace elements with a pH of 2.5. The pH of the bioreactor was set up at 4.5. The excess of H2S produced in the bioreactor by the SRP activity was delivered to the offline vessel with a stream of nitrogen with a flow rate 100 mL/min, containing 0.1 (v/v) of H2S. Between days 43 and 84, the HRT decreased steadily to reach an average HRT of ~34 h (Fig. 6b). The H2S produced during this period was used in the experiments to determine the influence of pH for the AsIII removal as sulphide

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Fig. 6 Sulphidogenic bioreactor (a) operating continuously under acid conditions for the offline precipitation of AsIII as sulphide, yellow-coloured vessel; changes in the HRT (open square) and glycerol in the effluent ( filled red triangle) during the operational phase (b) and rate of AsIII removal as sulphide at different initial pH with the produced biogenic H2S (c)

(Segura and Nancucheo, unpublished). For this, AsIII solutions with 100 mg/L were prepared at different initial pH (1.0; 3.0; 5.0; 7.0; 10). Each solution in the offline vessel was sparged with the H2S stream generated in the sulphidogenic bioreactor as a consequence of the consumption of glycerol supplied as electrons and carbon source (Fig. 6b). Between pH 5 and 10 (Fig. 6c), the removal of AsIII can be optimised obtaining a removal of 40 mg/L per hour using H2S off-gas produced with a continuous sulphidogenic bioreactor. Even though, the final pH of this reaction is determined by the protons from the H2S, the results obtained highlighted that sequential recovery of AsIII from acid process water might be done using pH control, particularly if the stream contains AsIII and Cu. Metal(oid)s have different solubility product (Ksp) and therefore if the concentration of soluble sulphide generated by biosulphidogenesis is limited (particularly at low pH), those with lower Ksp precipitate while those with larger Ksp remain in solution. Copper, for example, is far less soluble than AsIII (log Ksp values of -35.9 and -16.0, respectively) and Cu precipitates at pH 2 as sulphides, whereas AsIII requires a much higher pH to precipitate [73, 74].

Sulphidogenic Bioprocesses for Acid Mine Water Treatment and. . .

4.4

21

Integrated Processes

In the treatment of mine drainage, it is standard and necessary to have multiple complementary stages in both passive [75] and active treatment, even though there are treatment systems consisting solely of SRP-based bioreactor [76]. There are numerous comprehensive papers that describe the criteria to help choose the optimal combination of treatment stages [77]. Depending on mine water chemistry, flow rate and site space available, SRP reactor may face a number of limitations. Although SRP can generate alkalinity and thereby has the potential to neutralise acidic conditions, their activity is often more effective at near-neutral pH levels [39]. Challenges may emerge if the selected high-kinetics SRP consortium is inhibited by the acidic pH of the AMD, if the site’s available space limits the size of the SRP system that can be installed, or if the AMD’s acidity is beyond the tolerance of any SRP. Under such circumstances, pH adjustment or buffering strategies become necessary to ensure effective treatment. The oldest and most common solution to these issues are successive alkaline producing system (SAPS) [78] and reductive alkaline producing system (RAPS) [79]. SAPS are designed to treat AMD by utilising layers of organic and limestone material to create conditions favourable for SRP. The system typically consists of a limestone layer covered by a layer of organic substrate, such as compost or manure, which provides carbon sources for the SRP. As AMD percolates through the system, the SRP consume the organic matter and produce alkalinity, increasing the pH and facilitating the precipitation of metals. The limestone layer further adds alkalinity to the system, ensuring the efficient removal of acidity and metals. RAPS are similar to SAPS but are specifically designed to operate under conditions with reduced oxygen ingress. The reduced oxygen levels help to minimise the potential for metal hydroxide precipitation at the system’s inlet, promoting more efficient metal removal by reduction further within the system. But other solutions also exist, for example combinations of SRP bioreactor and anoxic limestone drain (ALD) have been used to assist SRP in pH elevation [80, 81]. ALD are passive treatment system developed in the 1980s and designed to treat AMD by neutralising its acidity without introducing oxygen to the water. Essentially, it is a buried, limestone-filled channel or trench that intercepts AMD before it becomes oxygenated. As the AMD flows through the limestone in an anoxic (oxygen-free) environment, the acidity in the water dissolves the limestone, releasing alkalinity and thereby increasing the pH of the water. This process effectively neutralises the acid without the limestone being coated by iron oxides, which can occur when there is too much oxygen present, and which would otherwise reduce the effectiveness of the limestone in neutralising acidity [82]. Lime pre-treatment has also been used for the same purpose [83]. Finally, we can assume that there are probably very interesting synergies between Dispersive Alkaline Substrate (DAS) and SRP bioreactors, although this has not yet been demonstrated on a full scale. DAS is a new material used in the remediation of AMD for the first time at full scale in 2011, which combines fine-grained limestone

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sand with a coarse inert matrix such as wood shavings, was created to address the shortcomings of conventional remediation systems intended for coal mines, which have proven inadequate for treating mine waters from the Iberian Pyrite Belt. The high acidity and metal content in these waters lead to rapid precipitation of solids and quick clogging or passivation (coating) of the reactive grains, challenges that DAS aims to mitigate. The small grains provide a large reactive surface and dissolve almost completely before the growing layer of precipitates passivates the substrate. The high porosity retards clogging [84]. DAS exhibits distinctive features, including its ability to remove Fe3+, consume dissolved oxygen and increase pH levels, creating an optimal environment for the subsequent integration of an SRP system downstream. Its capacity to remove divalent ions such as Co, Ni and Zn is somewhat limited, thereby underscoring the complementary nature of SRBs in the treatment process. In addition, spontaneous sulphate-reduction activity has been observed within the depths of DAS, highlighting the compatibility between the water treated by the DAS and the conditions favourable for SRP activity [85]. Although SRP can precipitate heavy metals as sulphides, high concentrations of certain metals can be toxic to the bacteria [86]. Metals such as copper, chromium and cadmium can inhibit SRP activity sometimes even at relatively low concentrations, affecting the overall treatment process. Under these circumstances, employing SRP processes that separate the precipitation of metals with H2S from sulphate reduction might be advantageous, as described in [87]. Alternatively, implementing a partial pre-treatment of the effluent to lower the concentrations of these inhibitory elements could be beneficial. Among the techniques available for this purpose, alkaline treatments, as those previously discussed, can be particularly effective. SRP bioreactors require a continuous supply of organic substrates as electron donors for the reduction of sulphate to sulphide. The cost and availability of suitable substrates can be a limiting factor, especially for large-scale or continuous operations [88]. To optimise the use of organic matter, it may be beneficial to specifically target the precipitation of sulphides for elements of interest, either for their economic value for recovery purposes such as Cu, Ni, Co, Zn or due to their environmental impact for waste management considerations such as As and Cd, in contrast to Fe and Al. Furthermore, the biological reduction of sulphate generates hydrogen sulphide (H2S), a toxic and corrosive gas. Managing H2S emissions is crucial to ensure safety and environmental compliance. In addition, excessive concentrations of H2S can inhibit SRB activity itself [89]. Intentionally limiting the input of organic matter can also be a way of ensuring that no H2S is released. In order to manage excess H2S production, frequently employed techniques alongside SRP bioreactors typically involve oxidation steps including aerobic wetlands and combinations of cascades and settling ponds for low-acid effluents [90], or oxidation bioreactors for acid effluents. Aerobic stages can be positioned either downstream the SRP treatment – and then SRP activity need to be constrained by residence time and/or a limited supply of organic matter – or upstream, provided that the residual presence of soluble Fe3+ can be managed before reaching the SRP bioreactor. For acid effluents with significant Fe content, a combination of oxidation and alkaline treatment methods might be necessary to effectively manage the treatment process.

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When dealing with limited space, it is crucial to meticulously weigh the pros and cons of each treatment technology. Although SRP systems necessitate extended residence times, leading to the need for larger reactor volumes compared to aerobic methods, the latter produces hydroxide sludge that is significantly less dense (and thus more voluminous [91] than the sulphides precipitated by SRP) [92]. This discrepancy in sludge density means that the space required to accommodate the voluminous aerobic sludge could negate the spatial benefits offered by the more compact aerobic stages, unless regular cleaning is feasible. Maintaining optimal conditions for SRP activity (such as temperature, pH and nutrient levels) requires careful monitoring and control, which can be complex and costly. Moreover, the accumulation of metal sulphides can lead to clogging in the reactor, necessitating regular maintenance and potential downtime [93]. Additionally, the decomposition of organic matter can lead to the accumulation of undesirable by-products such as smell and residual oxygen demand. Hence, it frequently becomes essential to implement stages downstream of the SRP bioreactors, occasionally incorporating more resilient phases, to compensate for the SRP bioreactor variable performance and more commonly to provide a final polish. For this purpose, wetlands [81, 94], terraced aerobic polishing cells [95], rock filter aeration cells paired with aeration polishing ponds [96] and horizontal-flow limestone beds [90] can be implemented.

5 Perspectives The recovery of metals and metalloids from mine water and process water from metallurgical industries would be beneficial for the protection of environment and water resources and could contribute to meet the growing needs for critical elements including As. Bioprocesses based on the activity of SRP offer promising perspectives for the treatment of acidic streams containing metals and As. According to the composition and flow rate of the stream to be treated, and the fixed objectives, different bioreactor options could be applied, from the most active flow-sheet to semi-passive configurations. The selection of SRP enrichments tolerant to acid conditions enlarged to potential of bio-process configurations. In the most active processes, the biogenic H2S can be stripped from bioreactors and injected in separate tanks, with the possibility to precisely control pH at each treatment step by injection of acid or basic solutions. Conversely, in semi-passive systems, selectivity could be approached by limiting the electron donor supply. The feasibility of As and metals separation from Fe in an SRP bioreactor fed with two types of real As-rich mine drainage waters was demonstrated at laboratory scale. Limitation of sulphate reduction by low supply of glycerol allowed the removal of most arsenic while iron remained in the water [63]. The selective separation of As from Fe in As-containing AMD can be beneficial in terms of hazardous waste production compared to the direct lime treatment. Downstream the SRP bioreactor removing As and possibly other toxic elements

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precipitating at low pH as sulphides (Sb, Cd, Pb, Hg), a lime (or NaOH) treatment could be applied to rise pH at an environmentally suitable value and precipitate “clean” Fe hydroxides. The different batch and continuous experiments performed with the model Carnoulès AMD underlines the potential and difficulties of selective precipitation of As, Sb and metals, in particular in passive or semi-passive conditions when pH evolution is only controlled by electron donors dosing. However, the implementation of tests in batch and continuous conditions, performed with real mine water, allowed to progress in the understanding of phenomena, suggest possible improved bioreactor configurations and should help to identify the most promising types of real effluents that could be candidate for recovery of elements in the form of bio-sulphides. For some metals such as Zn or Ni, a selective precipitation in successive semi-passive systems might be obtained through dosing electron donor supply, if the initial geochemistry (proportion between As and metals) and initial pH were favourable. The feasibility of selective precipitation could be evaluated through bio-geochemical modelling and simulations based on experimental results. The speciation of metals (FeII versus FeIII) and metalloids (AsIII, AsV, SbIII, SbV) will influence the H2S consumption and must be taken into account in the theoretical and simulation steps. The mineralogy, structure and specific surface of biogenic sulphides can strongly differ from precipitates produced chemically by dissolving H2S into metals and metalloids-containing solutions. The observation of biofilms sampled in SRP bioreactors revealed that some elements, and particularly As, can precipitate forming a crust around bacterial cells. In batch experiments with real AMD, nanowires of realgar were produced [29]. These biogenic particles might present interesting properties that could find applications. The characterisation of bio-precipitates of arsenic sulphides and other biogenic sulphides should be explored as indices of the suitable conditions for their formation thanks to SRP activity were acquired. Moreover, if the use of SRP for metal recovery is to be developed further, there is a need for specific characterisations aimed at developing processes to separate valuable metals from gangue and then concentrate these metals to create a marketable product. Given the diffuse and poorly crystallised nature of these sulphides, these recovery processes will likely rely on hydrometallurgy. However, incorporating mineral processing steps should not be ruled out such as the gypsum separation via hydrocyclone developed in the SAVMIN process [97]. Beyond technical aspects, a new conceptual approach could enable the development of valuable metal recovery using SRP systems. Due to low flow rates or concentrations, the streams of valuable metals in mine drainage are often too minimal to justify a continuous water treatment process for their recovery. However, mine drainages typically have a very long lifespan, and their treatment represents a net cost to society. Current SRP bioreactors, especially passive systems, accumulate metals within their filter mass over years. Where a continuous metal recovery process from mine drainage may not be economically viable, a short and targeted operation to recover the substantial mass of metals (accumulated over years) in an SRP bioreactor could be. Such an operation could be conducted by mobile units

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during the dredging and decennial refurbishment of SRP bioreactors and could offset the water treatment cost to some extent. For these purposes, SRP bioreactors would benefit from being designed from the outset with the dual aim of treating water and producing sludge materials that are more readily valorised. The separation of precipitation of different metal sulphides into separate cells is the first and most obvious example of this kind of design modification [98]. This proposed new conceptual approach is inspired by the “anthropogenic ore” concept developed by Sapsford et al. [99]. Acknowledgement The research reported by Ivan Nancucheo is supported by the project FONDECYT Chile 12211606.

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Adv Biochem Eng Biotechnol (2024) 190: 31–88 https://doi.org/10.1007/10_2024_255 © The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 Published online: 2 July 2024

Biological Iron Removal and Recovery from Water and Wastewater Anna Henriikka Kaksonen and Eberhard Janneck

Contents 1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2 Principles of Biological Iron Removal and Recovery from Water and Wastewater . . . . . . . . 2.1 Aerobic Biological Iron Rmoval and Recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2 Anaerobic Biological Iron Removal and Recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3 Bioreactor Types for Biological Iron Removal and Recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1 Suspended Cell Bioreactors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2 Examples of Bioreactor Studies on Biological Iron Removal and Recovery . . . . . . . . . 4 Pilot- and Full-Scale Applications of Biological Iron Removal and Recovery from Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1 Aerobic Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2 Anaerobic Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 Patents for Biological Iron Removal and Recovery . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 Comparison of Biological Iron Removal and Recovery Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . 7 Beneficial Uses of Iron Products Removed from Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

32 33 34 41 43 43 48 56 56 64 65 65 70 79 80

Abstract Iron is a common contaminant in source water and wastewater. The mining and metallurgical industries in particular can produce and discharge large

A. H. Kaksonen (✉) Commonwealth Scientific and industrial Research Organisation (CSIRO) Environment, Floreat, WA, Australia Western Australian School of Mines: Minerals, Energy and Chemical Engineering, Faculty of Science and Engineering, Curtin University, Bentley, WA, Australia School of Engineering, University of Western Australia, Crawley, WA, Australia e-mail: [email protected] E. Janneck G.E.O.S. Ingenieurgesellschaft mbH, Freiberg, Germany e-mail: [email protected]

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quantities of wastewater with high iron concentrations. Due to the harmful effects of iron on organisms and infrastructure, efficient technologies for iron removal from water and wastewater are needed. On the other hand, iron is a valuable commodity for a wide range of applications. Microorganisms can facilitate iron removal and recovery through aerobic and anaerobic processes. The most commonly utilized microbes include iron oxidizers that facilitate iron precipitation as jarosites, schwertmannite, ferrihydrite, goethite, and scorodite, and sulfate reducers which produce hydrogen sulfide that precipitates iron as sulfides. Biological iron removal has been explored in various suspended cell and biofilm-based bioreactors that can be configured in parallel or series and integrated with precipitation and settling units for an effective flow sheet. This chapter reviews principles for biological iron removal and recovery, the microorganisms involved, reactor types, patents and examples of laboratory- and pilot-scale studies, and full-scale implementations of the technology. Keywords Biooxidation, Bioprecipitation, Iron, Recovery, Wastewater, Water

1 Introduction Iron is the most abundant element on Earth and the fourth most abundant in Earth’s crust by mass-% [1]. Elemental iron is rarely found in nature, as iron ions Fe2+ and Fe3+ readily react with oxygen- and sulfur-containing compounds to form various mineral phases such as oxides, hydroxides, hydrous oxides, carbonates, silicates, and sulfides [2, 3]. Iron is an essential nutrient for most life forms [2], has a central role in electron transport processes [4], is part of hemoglobin, a protein responsible for transport of oxygen in the blood of mammals [5], and plays an important role in many microbial enzymes [6]. Iron is also present in many wastewaters from mining, metallurgical operations, electroplating [7], and tanneries [8]. In hydrometallurgical processes, iron may accumulate in process streams from gangue minerals such as pyrite and pyrrhotite, and the presence of excess iron can cause the formation of precipitates and affect downstream recovery of valuable metals. The presence of iron in wastewater effluents can have detrimental impacts to receiving environments. Excess ferrous iron can cause oxygen consumption in receiving water bodies as it is oxidized to ferric iron. Iron-rich precipitates hamper fish spawning [9] and smother river sediments, decreasing oxygen diffusion and killing benthic organisms [10, 11]. Iron precipitation also increases turbidity of water, and thus reduces light penetration and primary production [10–12]. This may have a notable effect on the food chains in impacted waters [10]. Iron also causes undesirable microbial growth within water works [3], discoloration and staining, odor, metallic taste, and scaling and clogging of pipes and other

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infrastructure [13, 14]. The World Health Organization (WHO) does not have a health-based guideline for iron in drinking water. However, iron at concentrations above 0.3 mg L-1 stains laundry and plumbing fixtures, and turbidity and color may develop in pipe systems at concentrations above 0.05–0.1 mg L-1 [15]. Ferrous iron concentration of 0.04 mg L-1 can be detected by taste in distilled water, whereas threshold iron concentrations were 0.12 mg L-1 for mineralized spring water with total dissolved solids content of 500 mg L-1, and 0.3 mg L-1 for well-water [3]. Long-term consumption of water with elevated iron concentrations can cause so-called “iron overload” [13, 16], which includes impairment of hematopoiesis (the formation of blood cellular components). Untreated iron overload may cause hemochromatosis which damages organs [13, 17–22] and leads to weight loss, joint pain, and fatigue. Other health impacts of excess iron intake may include eye disorders such as retinitis, conjunctivitis and choroiditis, heart disease, and cancer [13, 23]. Due to the potential detrimental impact of excess iron, efficient technologies are required for removing iron from mining and industrial effluents before release into the environment. The recovered iron can potentially also be directed for beneficial use. Traditionally iron can be removed from hydrometallurgical process waters and mine waters using physical-chemical methods, such as precipitation as iron hydroxides, with alkaline chemicals, such as limestone (CaCO3), hydrated lime (Ca(OH)2), quicklime (CaO), soda ash (Na2CO3), sodium hydroxide (NaOH), and ammonia (NH3) [24]. Most common chemical is hydrated lime due to cost considerations, high reactivity, and easy availability. Chemical treatment is costly and requires dispensing equipment and facilities [24]. Iron-hydroxide sludges are often disposed in mine residue areas, creating legacy issues for future generations. Biotechnical processes can offer sustainable alternatives that have the potential to reduce chemical consumption and waste generation. Biological iron removal can be based on bio-catalyzed oxidation and reduction reactions, bioprecipitation, and biosorption. Biosorption has not yet been implemented in industrial scale, whereas bioprecipitation has been widely applied [25]. This chapter provides an overview of biotechnical iron removal and recovery approaches that utilize bioprecipitation and discusses their benefits and drawbacks.

2 Principles of Biological Iron Removal and Recovery from Water and Wastewater Iron removal and recovery from wastewater can be facilitated by both aerobic and anaerobic bioprocesses and the composition of iron precipitates varies depending on solution composition, pH, and temperature. In nature, iron occurs mainly in two oxidation states, namely +2 and + 3. The prevailing form is dependent upon factors such as pH, oxygen concentration, and redox conditions [26, 27].

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Aerobic Biological Iron Rmoval and Recovery

Aerobic biological iron removal and recovery is based on the ability of ironoxidizing microorganisms to oxidize ferrous iron (Fe2+) to ferric iron (Fe3+) (Reaction 1) and precipitation of ferric iron [28]. 4 Fe2þ þ 4 Hþ þ O2 → 4 Fe3þ þ 2 H2 O

ð1Þ

Ferrous iron (Fe2+) is stable in anoxic environments, but is spontaneously chemically oxidized by oxygen in aerobic conditions. The rate of abiotic oxidation depends on the concentration of protons, dissolved oxygen, ferrous iron, and temperature [2, 27]. Spontaneous chemical oxidation of ferrous iron is very low at pH < 4 [2]. Biocatalyzed iron oxidation is of particular interest for the treatment of hydrometallurgical process waters because at low pH values (< 2) it occurs several orders of magnitude faster than abiotic oxidation at ambient pressure and temperature [29, 30]. At higher pH values abiotic Fe2+ oxidation starts to compete with biooxidation as shown in Fig. 1 [31]. At pH > 5, the rate of abiotic Fe2+ oxidation increases by a factor of 100 when the pH is raised by one unit [32, 33]. At pH > 8 and Fe2+ concentrations >50 mg L-1, the availability of dissolved O2 becomes the rate-determining step. This is the experience of the authors from the operation of a

Fig. 1 The effect of solution pH and redox potential (Eh, against standard hydrogen electrode) on the stability of Fe2+ and the competition between biological and abiotic Fe2+ oxidation (Adapted from [31])

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large industrial chemical water treatment plants for iron removal with high Fe2+ concentrations of 50 to 700 mg L-1 and it fits with the findings of Kirby et al. [34] and Gil et al. [35] concerning the influence of O2 concentration on Fe2+ oxidation rate. Biological iron oxidation is carried out by a range of bacteria and archaea. Iron oxidizers can be divided into acidophiles and neutrophiles. In terms of temperature, iron oxidizers include psychrotrophs which tolerate low temperatures, mesophiles, and thermophiles. Examples of iron-oxidizing bacteria and archaea are shown in Tables 1 and 2, respectively. The ability of microorganisms to gain energy from iron oxidation depends on the redox potential of the ferrous/ferric couple and the redox potential of potential electron acceptors (Fig. 2). The redox potential depends on solution pH (which affects the solubility of iron) and the presence of possible complexing agents [2]. The redox potential of the ferrous/ferric couple is most positive at low pH values where most iron species are soluble. With increasing pH values (from 2.5 upwards) the reducing power of ferrous iron increases enormously due to the formation of ferric hydroxy- and oxyhydroxy compounds [39] and the redox potential of Fe3+/ Fe2+ decreases continuously and can even turn to negative values. At circumneutral pH, the redox couples Fe(OH)3/Fe2+ and Fe(OH)3 + HCO3-/FeCO3 prevail with the solid phases having a fairly low solubility and redox potentials of E’0 (transformed standard redox potential for biochemical reactions at pH = 7) = -236 mV and E’0 = +200 mV, respectively (Fig. 2). The presence of sulfate and other complexing agents also lowers the redox potential of the ferrous/ferric couple as compared to the free ferrous/ferric couple [2, 36]. In very acidic environments, oxygen is the only usable electron acceptor that enables energy generation from iron oxidation. Perchlorate (ClO4-) could theoretically also act as an alternative electron acceptor under acidic conditions; however, its scarcity and toxicity to acidophiles limit its feasibility [2]. The redox potential of the oxygen/water couple depends on solution pH as protons are consumed in oxygen reduction according to Reaction 2 [2]: 0:5 O2 þ 2 Hþ þ 2 e - $ H2 O

ð2Þ

Oxygen is more energetically favorable as an electron acceptor for iron oxidation under acidic than neutral conditions, as the redox potentials of the oxygen/water couple are +1,120 mM and + 820 mV at pH 2 and pH 7, respectively [2, 40]. On the other hand, a neutral pH enables alternative electron acceptors to be used, as at pH 7 the redox potential of ferrous carbonate/ferric hydroxide is sufficiently low [2]. At neutral conditions, biological iron oxidation can also be coupled to photosynthesis by phototrophic purple bacteria, as the midpoint potential of photosystem I is approximately +450 mV [2, 39]. The biogenic ferric iron may precipitate as various mineral phases such as ferric hydroxide (Fe(OH)3), ferrihydrite (Fe2O3·0.5H2O), jarosite (AFe3(SO4)2(OH)6, where A is a monovalent cation), schwertmannite Fe16O16(OH)x(SO4)y·nH2O

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Table 1 Examples of iron-oxidizing bacteria with potential for biological iron removal and recovery (Adapted from [2, 31, 36–38]). NA = not available Species Acidiferrobacter thiooxydans Acidihalobacter prosperus Acidimicrobium ferrooxidans Acidithiobacillus ferrivorans Acidithiobacillus ferrooxidans Acidovorax ebreus Alicyclobacillus aeris Alicyclobacillus disulfidooxidans Alicyclobacillus ferrooxydans Alicyclobacillus pohliae Alicyclobacillus tolerans Aquabacterium sp. Azospira oryzae Chlorobium ferrooxidans Dechloromonas agitata “Ferrovum myxofaciens” Ferrimicrobium acidiphilum Ferrithrix thermotolerans Ferritrophicum radicola “Gallionella ferruginea” Leptospirillum ferrooxidans Leptospirillum ferriphilum “Leptospirillum ferrodiazotrophum” Leptothrix cholodnii Leptothrix discophora Marinobacter aquaeolei “Mariprofundus ferrooxydans” “Paracoccus ferrooxidans” “Pseudogulbenkiania ferrooxidans” Rhodobacter capsulatus Rhodomicrobium vannielii Rhodopseudomonas palustris Rhodovulum iodosum Rhodovulum robiginosum Sideroxydans paludicola Sulfobacillus acidophilus Sulfobacillus benefaciens Sulfobacillus sibiricus Sulfobacillus thermosulfidooxidans

Metabolisma Au A F Au Au ND, M Au F Ae NA F ND ND/Ae, H P F or M Au H H Ae Au; Ae Au Au Au

Temperature range (optimum) (°C) 21–47 (38) 23–41 (33–37) 4.5–5.0, depending on sulfate concentrations) as shown in reactions (1, 2, and 3) [17, 20, 21], changes of pH and sulfate concentration can contribute to the removal of this metal from mine waters. 4Al3þ þ SO4 2 - þ 14H2 O $ Al4 ðSO4 ÞðOHÞ10 · 4H2 O þ 10Hþ 4Al3þ þ SO4

2-

þ 22–46H2 O $ Al4 ðSO4 ÞðOHÞ10 · 12–36H2 O þ 10Hþ Al



þ 3H2 O $ AlðOHÞ3 þ 3H

þ

ð1Þ ð2Þ ð3Þ

This was demonstrated in the early 2000 by Tabak et al. [100], who designed a system of several reactors to recover metals from AMD from the Berkeley acid pit lake. The first process developed was a four-reactor system that involved a first reactor for precipitation of Cu and Zn as CuS and ZnS, respectively, at pH 3.0 using biogenic hydrogen sulfide followed by a second reactor to precipitate aluminum hydroxide at pH 4–5. However, the pH in the second step was chemically adjusted using KOH and the hydrogen sulfide had to be removed from the system before entering the second reactor to avoid precipitation of iron sulfides. The third reactor was kept at pH 5.0 and hydrogen sulfide was reintroduced into the reactor to precipitate ferrous sulfide. The last reactor set at pH 8.0 was used to precipitate the remaining metals as sulfides. The purity of ZnS and CuS was over 95% and the purity of aluminum hydroxide was over 90%. The systems showed the possibility of using sulfidogenesis for the recovery of metals, even though the recovery of aluminum was due to the chemical increase of pH. The system could be improved so the recovery is entirely driven by microbial activity. If the iron is removed previously to aluminum, it would be possible to combine sulfate reduction/sulfidogenesis for the recovery of the different metals. This was demonstrated by the same authors when

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they introduced a step for iron removal as ferric hydroxide between Cu and Zn precipitation, so that precipitation of aluminum could take place in the presence of hydrogen sulfide. A conclusion from the work of Tabak et al. [100] is that special attention needs to be paid on pH control of the reactors to obtain pure metal precipitates and on the metal chemistry to avoid impurities on the precipitates. More recently, Falagán et al. [98] used a 2.3-L bioreactor containing a sulfidogenic microbial consortium to remove aluminum from artificial acid mine water (pH 2.5–3.0) containing 41–122 mg/L (1.5–4.5 mM) Al3+ and 564–1,228 mg/ L (5.9–14.9 mM) SO42-. The design of the experiment considered the pH at which aluminum forms hydroxysulfates in AMD effluents (pH 4.5–5.0). The experiment was inoculated with an enrichment that contained several acidophilic sulfidogenic bacteria known to reduce sulfate (Desulfosporosinus acididurans, Peptococcaceae PL4, Peptococcaceae CEB3) and other accompanying microorganisms [98]. The reactor was continuously fed with the acid mine water amended with glycerol used by the microbes to reduce sulfate. The pH inside the reactor was set to 5.0, and when the pH inside the reactor increased, the acid feed liquor was pumped in, decreasing the pH back to pH ~5.0. The combined action of proton consumption by the sulfate reducers (reactions 4 and 5) and proton release by the precipitation of aluminum as hydroxysulfates or hydroxides (reactions 1, 2, and 3) resulted in a net pH increase. 4C3 H8 O3 þ 7SO4 2 - þ 14Hþ → 12CO2 þ 7H2 S þ 16H2 O 4C3 H8 O3 þ 3SO4 2 - þ 6Hþ → 4CH3 COOH þ 4CO2 þ 3H2 S þ 8H2 O

ð4Þ ð5Þ

Aluminum precipitation was mostly in the form of minerals such as felsöbányaite, hydroxybasaluminite, and alunite (reactions 1 and 2; Fig. 4). The removal of aluminum was correlated to the hydraulic retention time during the last 60 days of the experiment (r2 = 0.7), when the reactor showed a stable microbial community and relatively constant glycerol consumption (Fig. 5). This was supported by the mass balance for net proton generation and consumption. At given time, ~54% of the protons used for the microbial reduction of sulfate were derived from the formation of aluminum hydroxysulfates (assuming felsöbanyaite was formed by reaction 1; Fig. 3), while ~41.5% were derived from the free protons from the feed liquor and ~4.5% from the dissociation of the acetic acid and hydrogen sulfide. Thus, the possibility of using sulfate reducers for the recovery of aluminum from AMD was shown to be technically feasible. The precipitation of aluminum was also observed for the acidophilic sulfatereducing bacterium Thermodesulfobium sp. strain 3baa [101]. The strain was grown in batch assays (100 mL) using mineral media adapted to the composition of pore waters of an acidic pit lake in Lusatia, Germany, from which the strain had been isolated [102]. The medium contained 108 mg/L (4 mM) Al3+, 280 mg/L (5 mM) Fe2+, and 1,776 mg/L (18.5 mM) SO42-. The pH was set at 3 and the experiment was run at 25°C. In contrast to the experiment of Falagán et al. [98], hydrogen gas (H2) was used as electron donor, which is considered to be a suitable substrate especially under low-pH conditions and has been applied in the form of synthesis gas at larger

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Fig. 4 SEM images of precipitates and microbial cells collected in the aluminum-precipitation bioreactor designed by Falagán et al. [98]: (a) cells and felsöbanyaite precipitates; (b) sparse cells and aluminum oxyhydroxysulfate particles; (c) rods and aluminum precipitates; (d) microbial cell and aluminum oxyhydroxysulfate particles. The white circles with numbers indicate spots for mineralogical identification of the aluminum solids (as deduced from EDX analyses): 1–3, 8, felsöbanyaite; 4, 5, 10, hydrobasaluminite; 6, 7, 9, 11, 12, alunite

Fig. 5 Glycerol ( filled violet circle) and aluminum ( filled yellow square) consumption (%) and hydraulic retention time (in hours, filled green triangle) in the bioreactor experiments by Falagán et al. [98] (modified from [98], with kind permission from Elsevier)

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scale to treat acidic, metal-rich wastewaters (fixed-bed reactor with 30-L pore volume, [103]; 6-L bioreactor, [104]). As soon as sulfate reduction began and pH increased, aluminum started to precipitate which was shortly followed by iron sulfides. SEM combined with EDX revealed the formation of large amorphous spherical Al-rich precipitates in the bulk phase and smaller globules on cell surfaces [101]. The globules accumulated and became larger over time finally covering completely the bacterial cells. Similar observations were made earlier in enrichment cultures at pH 3 [102]. Large spherical aluminum precipitates were also reported for field samples from acidic Al-rich pit lakes ([24]; see also Fig. 2). The absence of the sulfur in the EDX spectra of precipitates forming in the cultures of Rüffel et al. [101], however, suggested the presence of gibbsite instead of hydroxysulfates. The TEM did not indicate accumulation of aluminum within the cells—neither in the cytoplasm nor in the periplasm [101]. For (enrichment) cultures set to higher pH values (pH 4, 5, or 6) considerable smaller aluminum particles were observed [102]. The dependence of aluminum particle size on neutralization rate and the possible influence of microorganisms were investigated in a follow-up experiment by Maar [99] using a mineral medium with 108 mg/L (4 mM) Al3+ and 1,440 mg/L (15 mM) SO42- posed to pH 3. Aluminum particles were produced by i) immediate titration with 1 M NaOH, ii) slow titration of the mineral medium with 0.1 M NaOH, and iii) growth of Thermodesulfobium sp. 3baa in the mineral medium. At pH 4.5 as growth or titration endpoint, the smallest particles (diameter average ± standard deviation; n = 15) were detected with immediate titration (48 ± 19 nm), followed by slow titration (131 ± 44 nm) and bacterial sulfate reduction (508 ± 221 nm) ([99]; Fig. 6). Although aluminum particles were also observed to be attached to surfaces (such as the polycarbonate carriers included in the batch assays) the particles formed mainly in the bulk phase. Aluminum particles adhering to bacterial cell surfaces were much smaller than those in the bulk phase and varied in numbers (Fig. 6c, d). Particle size and corresponding particle surface are of direct interest for the recovery of precipitated aluminum hydroxides/hydroxysulfates and their adsorption capacities toward other contaminants or metals. Finally, the occurrence of intracellular Al accumulation has also been accidentally observed in hyphae-like structure found in cultures of sulfate-reducing bacteria (Fig. 7). The identification and characterization of what it looks like fungi was not addressed at the time of the study, so that their identity and characteristics remain unknown. However, a detailed examination of the intracellular aluminum deposits by SEM showed a crystalline morphology with acicular crystals radially growing in the inside of the hyphae-like structure. The composition of these compounds (obtained by EDX) included aluminum, iron, sulfur, and silica in molar proportions which could not be clearly related to a specific mineral phase. Besides, these intracellular compounds contained significant chromium (up to 0.3–0.5 wt.%; Fig. 7) which indicates that bioaccumulated aluminum solids may also incorporate toxic metals in their composition. This fortuitous finding is coherent with recent research which has reported aluminum bioaccumulation by different species of acidophilic fungi isolated from different sites (e.g., [105, 106]) and indicates that

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Fig. 6 SEM images of Al-rich globules forming in the medium at the endpoint of pH 4.5 reached by immediate titration (5.4 g/L aluminum, 54.7 g/L zinc, 25.4 g/L copper; [107, 108]). Several authors, therefore, recommend to exploit the potential of indigenous microorganisms [58, 109, 110]. In case of solids, the pulp density must be considered in respect to contact area as well as the problem of shear stress [55]. Possible growth inhibition of the performing organisms, the problem of accumulating biomass as well as the need for direct contact between cell and substrate for efficient degradation are all aspects which need to be considered when deciding for a direct, one-step or two-step, or indirect leaching process [58]. Indirect leaching (using the biotically produced acids) circumvents the problems of growth inhibition, in addition, separates the leaching process from the biomass production step and prevents the biomass re-adsorption of dissolved Al3+ and aluminum (hydr)oxides as shown by some fungi. However, if cell contact enhances the overall degradation process, direct leaching would be preferable. Especially organisms suitable for a direct one-step leaching process (such as

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bacteria well adapted to extreme conditions and at the same time exhibiting low biomass production) are potential candidates for dump leaching on a larger scale [58]. The examples given for recovery of aluminum from wastewaters which have been treated with alumina involved the use of photolithoautotrophic organisms that do not require the addition of large amount of electron donors such as organic or reduced sulfur compounds. Roberts et al. [69] reported the growth of the algae directly in the to-be treated water without the need for further supplement of nutrients. The growth of the algae within the detention basin allows aluminum removal at scale. Overall, concentrations of aluminum must be relatively low (micromolar range), as photosynthetic organisms such as cyanobacteria (which may be used for the treatment of Al-rich waste) are considerably sensitive to Al [111]. Further studies looking at suitable microalgae should be conducted to determine the ability for aluminum removal and/or recovery. There are several microalgae (e.g., Chlamydomonas spp. or Coccomyxa spp.; [112–114]) detected in very acidic and metal-rich mine waters (including aluminum) that may prove suitable for this task also in respect of aluminum recovery from AMD. The removal of aluminum from the wastewater is based on the adsorption of the metal to the surfaces of cellular and extracellular biomass. In order to recover the aluminum from the biomass, the biomass has to be first separated from the water (e.g., by sedimentation and/or centrifugation) so it can be again (bio)leached (e.g., by fungi) after desiccation or even pyrolysis, as has been described for metal-rich sewage sludge [115]. Hence, the growth of cyanobacteria or microalgae may be a low-cost alternative as these microorganisms do not require the addition of nutrients and their biomass could be harvested for other uses such as biogas and synthesis gas production. Future research should also consider chemoorganoheterotrophic organisms with other optimum pH for growth close to the optimum pH of aluminum precipitation as well as a higher aluminum tolerance. However, our focus should not only be on microorganisms, the ability of plants to accumulate metals should be also considered. There are some plants that are known to accumulate aluminum, producing oxalic acid in response to aluminum stress (e.g., [116]). In Melastoma malabathricum (Indian rhododendron) and in Fagopyrum esculentum (buckwheat) aluminum binds to oxalic acid as a non-toxic Al-oxalate which accumulates in the different plant tissues [117, 118], or bind with citric acid as in Hydrangea macrophylla (hortensia) that accumulates Al-citrate in the cell sap [119]. The application of higher plants is a possible alternative to the use of microorganisms when it comes to the amelioration of soils [75, 111] and would allow an easier separation of the aluminum accumulating biomass from the soil. It is also worth to mention the formation of aluminum nanoparticles using Sargassum ilicifolium biomass extract [120]. The method proposed requires heating a pH 4 solution containing the algae extract to 80°C. When in contact with aluminum sulfate, the formation of a cloudy white precipitate indicates the formation of Al2O3 nanoparticles. To obtain the nanoparticles then it is necessary to calcine the precipitate. The method seems energy-demanding and it is not clear if the mechanism in which the Al-particles are formed is metal-specific. It also needs to be

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investigated if other metals may also co-precipitate with the Al-nanoparticles if using a multi-metal solution. Recovery of aluminum from acidic mine water can be achieved by microbial sulfate reduction as shown in Sect. 5, which increases the pH and leads to the precipitation of aluminum hydroxosulfates and/or (hydr)oxides. Sulfate-reducing bacteria require anoxic conditions and a suitable electron donor, such as H2 or low molecular weight organic acids or alcohols. The use of organic waste material which includes mainly polymers would need to be degraded first into soluble molecules by other organisms. As discussed in Sect. 5, the process of growth and production of metabolites can be separated from the actual recovery (in this case, precipitation) process. This avoids the problem of low acid and metal tolerance by SRB in case of highly fluctuating wastewater compositions. In addition, the produced hydrogen sulfide could be stripped from the spent medium and precipitation of sulfidic minerals could be separated from the precipitation of aluminum hydroxides. Currently, recovery of aluminum using sulfate reduction may not suppose a clear economic advantage since providing the optimum conditions for sulfate-reducing microorganisms to operate (e.g., suitable pH conditions, electron donors, and other nutrients) may surpass the benefits obtained from its recovery. Thus, further studies focusing on process optimization should be carried out to obtain reliable data to assess the economic cost and benefits. In addition, future studies to assess the efficiency of the process should focus on process optimization under natural conditions with real aluminum-rich acid waters. The process could be coupled to the selective recovery of other metals (e.g., Cu) using the hydrogen sulfide generated during sulfate reduction to produce metal sulfides (e.g., CuS). Further studies should focus on studying these processes in comparison to existing chemical processes to assess their feasibility or in combination with other chemical processes. Plants, algae, and/or other microorganisms like fungi and bacteria may serve as efficient drivers for achieving a circular economy and reducing environmental pollution.

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Adv Biochem Eng Biotechnol (2024) 190: 119–146 https://doi.org/10.1007/10_2024_257 © The Author(s), under exclusive license to Springer Nature Switzerland AG 2024 Published online: 15 June 2024

Precious Metal Recovery from Wastewater Using Bio-Based Techniques Sehliselo Ndlovu and Anil Kumar

Contents 1 Introduction ........................................................................................................................ 2 Precious Metal-Containing Wastewater Streams ............................................................... 3 Technologies for Biorecovery of Precious Metals from Waste Solutions . ........................ 3.1 Biosorption ................................................................................................................ 3.2 Bioaccumulation, Biomineralization, and Bioreduction ........................................... 3.3 Bioelectrochemical Recovery ................................................................................... 3.4 Membrane Biofilm Reactors ..................................................................................... 4 Value-Added Products Recovery ....................................................................................... 5 Comparative Assessment of Precious Metal Biorecovery Technologies and Traditional Technologies ...................................................................................................................... 6 Concluding Remarks .......................................................................................................... References ................................................................................................................................

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Abstract The recovery of metals from waste material has been on the increase in the past few years due to a number of reasons such as supporting the diversification of metal supply resources. In addition, the alternative use of the waste material for metal recovery can add to the main production line, boosting production throughput and profitability thus, allowing companies to sustain their activities during times of low commodity prices. While there has been a lot of research and interest in the recovery of precious metals such as platinum group metals (PGMs), Au, and Ag from solid waste material, there has been limited focus on the recovery of these value metals from wastewater. This is mostly related to challenges associated with finding cost-effective technologies that can recover these metals from solutions of low metal concentrations. In recent years, bio-based technologies have, however, become established as potential alternatives to traditional techniques in the treatment of S. Ndlovu (✉) and A. Kumar School of Chemical and Metallurgical Engineering, University of the Witwatersrand, Johannesburg, South Africa e-mail: [email protected]

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wastewater due to their ability to recover metals from solutions of low concentrations. While wastewater might be characterized by some significant value metal content, it also contains other components that have potential economic value if recovered or converted to by-products. Such an approach may not only provide an opportunity for extraction of metal resources from wastewater but also contributes toward the circular economy. This chapter presents insights into precious metal recovery from wastewater using bio-based technologies, compares such an approach to the traditional techniques, explores the recovery of other value-added products and finally considers some of the challenges associated with the large-scale application of the bio-based technologies. Keywords Precious metals, Bioprocessing, Wastewater, Recycling, Circular economy

1 Introduction The unique physical and chemical characteristics of “platinum group metals (PGMs),” i.e., “platinum, palladium, rhodium, ruthenium, iridium, and osmium” – imply that they are extremely versatile and can be used in numerous applications, spanning multiple industries. In addition to their appeal in jewelry, PGMs have long since played a significant role in the automotive industry through their application in manufacturing autocatalytic converters necessary to mitigate or neutralize hazardous pollutants from internal combustion engine exhaust emissions. The autocatalytic converter containing different amounts of platinum (Pt), palladium (Pd), and rhodium (Rh) acts as an active catalyst to convert toxic emissions to less toxic gases. The manufacturing of autocatalytic converters has consistently contributed over 70% to global demand for PGMs. On the other hand, the market demand for PGMs has consistently been satisfied by a supply from the mining processes to a large extent and recycling to a lesser extent. In response to the growing PGM demand, industry and academic research have attempted to explore the development of alternative material substitutes that can perform the same catalysis reactions as those of PGMs in autocatalytic converters. However, very little progress has been registered on this front and as such, PGM application in the automotive industry has continued to dominate [1]. With the world currently undergoing a transition from fossil fuels toward greener energy technologies driven by the global imperative to decarbonize, the internal combustion engine and the catalytic converter may, however, be consigned to the scrap heap of history. This, however, does not necessarily spell the end of a powerful demand for PGMs. Firstly, the unparalleled efficiency of the PGMs as catalysts for reducing emissions makes it highly unlikely that they will be replaced as the automotive catalysts of choice very abruptly. As such, PGMs are likely to continue their role in emissions control of the internal combustion engine “tail” that will exist

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as the world transitions away from fossil fuels. Secondly, the exceptional catalytic, physical, and chemical properties of PGMs will continue to underpin, strongly, both the innovative applications of existing uses and new ones. PGMs have been designated as crucial for facilitating the development of pioneering technologies underlying significant global shifts such as the clean energy transition. In the move toward carbon neutrality, these metals are considered critical in the hydrogen economy, where platinum catalysts in proton exchange membranes are instrumental to both the generation of hydrogen as a clean energy source through electrolysis of water and their application in fuel cells used for powering electric vehicles. While it is estimated that green hydrogen will portray a major role within the world’s energy mix and could supply up to 25% of the global energy demands by 2050, platinumcatalyzed hydrogen-fuelled electric vehicles are also being repositioned to the front of the class because they create zero emissions, which is in high demand around the globe. Thus, the potential for hydrogen to decarbonize not only the transportation sector but a larger portion of the energy sector is a good indicator of sustained demand for PGMs in the future. PGMs are rated as “critical” by the leading nations of the world. The term critical describes both the high supply risks and the high susceptibility of a system to a possible supply outage associated with a particular mineral. The global PGM supply from mining activities is mainly controlled by South Africa, Russia, and Zimbabwe [2]. South Africa has an estimated 80% of the global platinum reserves and contributed 70% of the platinum supply, 35% of the palladium supply, and 75% of the rhodium supply in 2014 [2]. However, mining activity in South Africa has recently been faced with numerous challenges such as erratic energy supply and labor-related disputes. On the other hand, Kolesnikova [3] reported that Russia’s palladium base was also dwindling, meaning that supply from palladium mining alone can no longer be depended on. Furthermore, the ongoing geopolitical crisis in Russia has not only created an energy shock but greatly emphasized the potential for rapid onset of uncertainty in the metal supply chain. For example, PGM prices increased in the first quarter of 2022 on fears of interruption to Russian deliveries emanated. However, while the latter half of the same year saw a constant improvement in the availability, and a general decline in prices, amidst weakness in primary and secondary supplies, the uncertainty of the war makes it difficult to predict the long-term stability in the availability of metals from this region. On the other hand, the uncertain social, economic, and political environment in Zimbabwe has hampered comprehensive and transparent investment into metal resource exploitation. Finally, mining companies also face increasing stakeholder demands and regulatory pressures for better environmental, social, and governance (ESG) principles. It is adequate to state that such above highlighted issues have the potential to bring production disruptions and significant uncertainty to the global PGM supply chain, as well as trade, diplomatic, and political issues. This has led to governments globally placing more emphasis on diversifying their metal supply chains and announcing initiatives to secure metals critical to their economic growth and development, particularly for meeting global climate goals.

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The recycling of spent autocatalytic converters is one of the most recognized contributors to an alternative supply chain of PGMs being second only to traditional mining sources. Another secondary source of PGMs that has, however, not received much attention is process effluent from refinery plants. The consistent increase in PGMs demand has led to the expansion of their extraction and refining operations around the world [4]. While most processing plants have optimized the refinery process as effectively as possible, there is, however, some discharge of PGMs wastewater to the environment, resulting in loss of valuable metals and unreclaimed water which is an environmental concern. In addition to the effluent from refinery streams, other potential sources of PGMs include municipal and medical wastewater streams. Besides PGMs, gold (Au) and silver (Ag) are also considered as the other components of the precious metals category. These metals are commonly used in coinage, jewelry, electronics, dentistry, and medicine. Gold, in particular, is considered as one of the most important sources of currency entry in the economics of any state and is used in finances and investing. Throughout history, this metal has been seen as a symbol of wealth and used for financial transactions. The first purely gold coins are believed to have been manufactured in 560 BC and this tradition continues today, with gold still considered the most popular precious metal for investments and financing. The extraction of gold and silver from ores is one of the oldest and well established activities in the history of mining. The early approach to gold extraction involved rudimentary artisanal mining techniques centered on placer mining. However, with the increasing complexities in the gold ore deposits, modern advanced technologies tend to consist of a much more comprehensive process flow sheet involving underground mining, cyanide leaching, carbon adsorption techniques. In addition, there has been a drive toward the adoption of much more environmentally friendly reagents as alternatives to the use of cyanide in response to increasing environmental concerns. At the same time, the ever increasing demand for gold coupled with the predominance of lower grade, more refractory, and complex ores has resulted in increasing capital investments and subsequent operational costs. While most companies make a reasonably good profit when the market is upside, it is the downside that can be threatening to the viability of most operations as a result of the high mining costs. Because of this, mining companies have also been exploring the diversification of their resource portfolios by considering not only the primary resources but also the secondary resource of gold and silver. The interest in gold extraction from secondary sources such as mine tailings, jewelry, dental applications, and electronic waste has been increasing. For example, 4,415 t of gold were supplied in 2017 worldwide mainly from primary resources but also, a significant and increasing part from recycled materials and tailings [5]. The discussion above shows that there is a growing recognition not only by mining companies but also by global governments that recycling underpinned by the circular economy model will be key to moderating the long-term need for primary resources and securing metal supplies. This phenomenon is not only driven by economic and social factors but also related to environmental aspects that are often

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neglected. Recycling offers the opportunity for companies to reduce their dependency of metal production from a single source, thus boosting the production throughput and profitability and allowing companies to sustain their activities during times of low commodity prices. There is also the advantage of reduction on ecological impact as the processing of such material offers a second opportunity to process them in a more holistic and responsible way with regard to leaving behind residues of less toxic nature to the environment. However, it is imperative to understand that the processing of secondary resources offers challenges that are slightly distinct from those posed by the processing of primary resources. As such, technologies that have been used in traditional metal processing industries might not give comparatively high yields and process efficiencies. Technologies such as adsorption, chemical precipitation, solvent extraction, ion and resin exchange, and metal scavengers such as Smopex® have been traditionally employed to remove minimal concentrations of precious metallic ions from wastewater. Despite the wide commercial application and acceptance of such technologies, they do, however, possess several drawbacks such as inadequate removal of metals, excessive reagent usage and/or energy demands, high capital costs, and the production of large amounts of waste products that need disposal [6]. Due to the type of reagents applied in the processes, most of the waste streams that are generated are not biodegradable and thus, tend to pose a long-term negative impact on the environment. Such issues are a major concern for the mineral and metal processing sector which is facing increasingly restrictive environmental laws and policies and, thus, driving the companies to place more emphasis on the development of environmentally friendly processes that would be more compliant to environmental laws. Satisfactory processes should be those that accomplish the target of metal recovery and wastewater treatment at a reasonable cost with minimum generation of environmentally hazardous residual solids. Some of the technologies that have been deemed to align with the abovementioned requirements are those that relate to the use of biological processes such as bioaccumulation, biosorption, and bioelectrochemical recovery. Biosorption and bioaccumulation processes differ from each other in a way that in the first process, the metallic contaminants get attached to the cell wall surface, whereas in the latter process, they are accumulated within the cell. Biosorption, in particular, has received significant attention due to the similarities in the principles of operational properties that it shares with the traditionally and widely understood absorption and ion exchange processes. Bioelectrochemical recovery systems are a growing technology with potential to simultaneously treat wastewater and recover other value resources. The system encompasses electrochemical recovery techniques such as “microbial fuel cells (MFCs), microbial electrolysis cells (MEC), microbial desalination cells (MDC), and microbial recycling cells (MRC)” [7]. This chapter will first look at the characteristic composition of wastewater streams containing precious metals. The chapter will then explore the different bio-based approaches that have been applied for precious metals recovery from aqueous wastewater streams. The chapter will also look at other by-products besides the metal components, which can be biorecovered through the treatment of these wastewaters. Finally, the chapter will

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conclude with a brief look at some of the challenges associated with the use of the bio-based technologies in the recovery of precious metals from waste solutions and offer suggestions on the way forward.

2 Precious Metal-Containing Wastewater Streams The continuously increasing industrialization and urbanization, population growth, and agricultural practices are significantly affecting the usage of available water resources and the quality of water as well. While the mining and metal-producing sector generally consumes a relatively small amount of water at the national level, it can, however, have a major impact on water use at the local and regional level if the water used requires a large proportion of the freshwater [8]. Thus, globally, mining companies are facing increased international pressure to minimize their freshwater consumption. At the same time, the metal extraction and manufacturing processes tend to result in the generation of an enormous volume of effluent streams characterized by metal and organic contaminants such as residual metal ions and chemicals associated with the metal extraction, purification, or recovery processes. As such, the constituents of the waste streams can range from metallic salts, organic solvents, cleaning aids, and greases [9]. Besides the mining and metal extraction waste solutions, other potential sources of precious metals include urban wastewater, where these metals are possibly transferred from urban mines (municipal solid waste, mining waste, end of life vehicles, and electronic waste) by runoff and surface water streams [10]. The presence of PGMs has long been recognized in municipal wastewater, which arise from road dust present in stormwater runoff [11]. The PGM content in these wastewaters typically arises from the deterioration of catalytic converters leading to the emission of PGM-containing particles. In addition, wastewater from hospitals, dental clinics, and pharmaceutical and fine chemicals industries has been noted to harbor ample concentrations of precious metals as well [12, 13]. For instance, Pt can be present in hospital wastewater, and this is largely related to the administration of anti-cancerous drugs such as cisplatin and carboplatin. Wastewater from dental hospital can contain Au, Ag, Pt, and Pd, because of the use precious metals alloys in indirect dental restorations (bridges, inlays, onlays, and crowns) [10]. Other streams that can contain significant precious metals like Au, Ag, and Pt are wastewater from jewelry and plating industry. Table 1 provides an overview of precious metals concentrations in the wastewater originating from the above-mentioned sources. Recovery of these precious metals can be economic at large process volumes, and, further due to the accumulation of metal values in the wastewater streams over time. The application of recycling, regeneration, and reclamation systems are some of the common strategies that are usually adopted by companies to minimize the draw on freshwater resources. Recycling and regeneration require that the treatment eliminate any contaminants that may have accumulated, as the treated water in

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Table 1 Precious metal concentrations in the wastewater of different origins

Wastewater types Dilute aqua regia wastewater

Plating industry wastewater Metallurgical effluents Hospital wastewater

Effluents from the oncology ward, in Iran Oncological unit in Vienna, Austria Precious metal refinery effluents Acid mine drainage

Gold mine wastewater

Cyanidation tank After INCO process Sewage sludge from German municipal wastewater treatment plant

Industrial wastewater

Type of precious metals Pt4+; Au3+; Pd2+ Au Au; Pt; Pd

Concentrations 4 mg/L; 53 mg/L; 46 mg/L

Reference [14]

[15] [16]

Pt

2,828 mg/L 4.2–5.8 mg/L; 12.9–15.3 mg/L; 131– 137 mg/L 3–762 μg/L

Pt

3–250 μg/L

[13]

Pd; Pt; Rh

[18]

Au; Ag; pd.; Pt; Rh Au; Ag

28–357 mg/L; 10–185 mg/L; 36– 179 mg/L 1.12 mg/L; 1.03 mg/L; 2.3 mg/L; 0.554 mg/L; 1.265 mg/L 0.561 mg/L; 0.664 mg/L

[20]

Au; Ag

0.260 mg/L; 1.5 kg dw of roots per month and m2) nitrophilic and metal-accumulating herbaceous or woody species that can be

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used for the removal of metal(loid)s in hydroponics (biofilters) in a stationary or moving aqueous environment (rhizofiltration) [126, 127]. Of these terrestrial plants, Vetiveria zizanioides, Phalaris arundinacea, and Festuca arundinacea are perennial and able to adapt and grow in water and acid environments. Vetiveria zizanioides produces extensive root systems that provide an enormous surface area and exhibit great potential for phytoremediation of wastewater contaminated with Zn, Fe, Cu, Cd, and Pb [120]. It can resist acidity, alkalinity, salinity, and heavy metals. Although it is a tropical grass, it can also tolerate extreme temperatures from -15 °C to 60 °C. Other terrestrial plant species commonly used in rhizofiltration belong to the Brassicaceae (e.g., B. juncea) and Asteraceae (e.g., H. annuus) but have the major shortcoming that these species are annual species, which complicates their handling in continuous hydroponic wastewater treatments. Although the emphasis in this chapter is on plant species suitable for phytoextraction of metalloids, it should be noted that for the removal of pollutants, other mechanisms, such as physical absorption of elements at the root surface or element exclusion during root–shoot translocation, are of great importance [128]. In Sect. 4.2, we will deepen the discussion on species-specific element accumulation among different species. Finally, in Sect. 4.3.5, we will focus the discussion on possible mechanisms to enhance the phytoextraction processes, including the synergetic effects of the selection of diverse species composition (mixed cultures) on the phytoextraction efficiency and spectrum of the target elements removed.

4.2

Element Accumulation in Plants

Element uptake and accumulation by plants is a complex process involving an array of plant-associated factors and factors associated with the physicochemical properties of the growth environment influencing the element’s mobility, chemical speciation, and, consequently, element availability for plant uptake [129–131]. Given that plants evolved acquisition strategies and uptake systems to utilize essential elements in their metabolism, it is not surprising that the accumulation observed for essential plant nutrients such as N, P, Fe, Mn, Cu, and Zn is often far higher compared to non-essential elements (e.g., Cd, Pb, As). In particular, plants require relatively high cellular concentrations of the essential macronutrients N, P, K, Mg, Ca, and S. On average, plants require 15 g/kg N, 10 g/kg K, 2 g/kg P, 2 g/kg Mg, 5 g/kg Ca, and 1 g/kg S in shoot dry matter for adequate growth [132]. Thus, phytoextraction of nutrients, above all the most abundant nutrients N and P, largely depends on growth rates and biomass development, which explains why large growing macrophytes like Typha, Scirpus (Schoenoplectus), Phragmites, and Juncus are most commonly used in constructed wetlands [86] for the removal of nutrients. Compared to the macronutrients, the concentrations of metallic micronutrients (e.g., Fe, Mn, Cu, Zn) and non-essential metal(loid)s are typically orders of magnitude lower. However, some plant species are able to accumulate more than 1000 mg/kg of essential and non-essential metals without signs of toxicity. These species evolved under

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conditions of high element mineralization in their growth environment and adapted to situations of high element availability by alteration of their functional architecture. Generally, the functional adaptations of plants include two basic physiological strategies: accumulators and excluders [26, 27]. The majority of plant species tolerate metal(loid)s in the substrate by physiologically restricting the element transport or excluding the elements at the sites of uptake through extracellular complexation with organic ligands [10]. Accumulation and detoxification may also occur by precipitation of hardly soluble compounds inside root cells, e.g., at walls or in mitochondria (e.g., Tl2O3). In contrast, accumulators efficiently acquire/utilize elements from the growth environment and avoid element toxicity in the roots by rapidly transporting the elements to the shoots, where they are sequestrated in the leaf tissue (petioles, sheathes), in vacuoles, and in trichomes [9, 10]. Some elements can be excreted via methylation and volatilization, like Se and possibly Hg (see Fig. 1). In the terrestrial environment, various plant species have been reported to be suitable for the phytoextraction of essential and non-essential metalloids, and these plant species are also potentially interesting for the rhizofiltration of wastewater. The majority of specialized hyperaccumulators belong to the Brassicaceae, Fabaceae, Euphorbiaceae, Asteraceae, Lamiaceae, and Scrophulariaceae (Ni, Cd, Pb, Zn). In addition, in some rather unspecialized hyperaccumulators, metal accumulation occurs as a consequence of their nutritional strategy in the rhizosphere (e.g., Mn accumulation in Proteaceae), where they unintentionally mobilize potentially toxic elements together with nutrients. More information on hyperaccumulation and metal (loid) phytoextraction in terrestrial environments is given by Brooks [128] and van der Ent et al. [133]. Similar to terrestrial metalliferous environments, in the aquatic environment, element availability is readily given because the elements are present in dissolved forms already (although the availability of dissolved complexes is still questionable), and their accumulation into plants is limited mainly but not solely by species-specific capacity for uptake and root–shoot translocation rather than element solubility in the rhizosphere [10, 23, 134]. Therefore, it is not surprising that some quite spectacular (hyper)-accumulations can be achieved in aqueous systems and plants belonging to the ecological group of macrophytes or hydrophytes that have already been an essential biocomponent of artificial wetland systems worldwide for more than 70 years (Table 3). It is generally assumed that there are similar pathways of element accumulation in vascular freshwater plants and vascular plants of the terrestrial environment, usually involving the absorption of metal(loid)s at the root surface, radial transport, and root–shoot translocation through active and passive membrane transport [10]. Striking differences in physiology among these functional groups, however, may arise from differences in adaptations to water availability, especially differences in root structure and transpiration rates [135] influencing element uptake and movement of elements within the plants by mass flow. With regard to non-essential elements, some plant species have a significant promise for phytoremediation of polluted waters by phytoextraction. Although metal(loid) concentrations in natural waters are typically low and often range between a few 100 ng/ L and μg/L [6, 136], the concentrations reported in naturally growing macrophytes are often orders of magnitude higher [29, 31, 115, 137] (Table 2). Under these

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Table 2 Compiled data on concentrations (average and ranges) of trace elements in soil, terrestrial plants, river water, and aquatic plants Element As Pb Co Cd Cr Cu Zn Ga Ge ∑REE

Soil averagea,b (μg/g) 6.83 27 11.3 0.41 59.5 38.9 70 15.2 2.0 182

Terrestrial plantsa (μg/g) 0.05–1.7 0.02–3.6 0.02–1.0 0.02–0.4 0.1–0.5 5–30 27–150 0.02–5.5 0.02–1.1 0.03–1.13

River waterb,c,d (μg/L) 0.1–2.7 0.01–3.8 0.02–0.4 0.001–0.4 0.3–11.5 0.2–2.6 0.3–27 0.005–0.01 0.004–1.1 0.05–1.1

Aquatic plantsb,c,d,e,f (μg/g) 2.7–8.2 6.1 0.3–2.5 1.0–1.4 2.2–4.0 7.9–12 37–52 0.5 0.6–2.8 9.8–171

a

Kabata-Pendias [55], bBrooks [29], cWiche et al. [111, 138], dRai [101], eValitutto et al. [31], Rawlence and Witton [139]

f

conditions, many macrophytes appear to be invariably hyperaccumulators, showing element accumulation factors (concentration in plants/concentrations in water) >> 1. Thus, according to Valitutto et al. [31] the plant–water concentration ratios varied from about 1000 to 200,000. In contrast, the bioconcentration factors are typically observed in most terrestrial plant species < 1 (Table 2). The literature indicates that the element concentrations of Co, Cu, Mn, and V in aquatic plants are very similar to terrestrial plants, whereas freshwater vascular plants may contain higher concentrations of Ag, As, Cd, Cr, Ni, Pb, Se, and U, most likely due to higher element mobility in the aquatic environment [137, 140]. Although plant nutrient uptake systems predominantly shuttle nutrient ions through membranes during radial transport and long-distance transport, there is broad evidence that at least some uptake systems cannot distinguish between essential nutrients and non-essential chemical cognates when the elements share chemical similarities with respect to charge and ionic radii. This has been demonstrated for Cd2+ and Zn2+ [9], Ca2+ and REE3+ [11, 134], SiOH4 and GeOH4 [111, 138], PO43- and AsO43- [29], and K+ and Rb+ [141]. Consequently, when phytoextraction of Ge is the desired goal, Si-accumulating grasses such as monocotyledonous macrophytes should preferably be selected. At the same time, dicots that typically contain more Ca in their plant tissues [142] are the better accumulators for REE [11, 112]. Element accumulation among different freshwater vascular plants may enormously differ between emergent and submerged life forms as well as free-floating species and life forms rooted in the sediment. Outridge and Noller [137] demonstrated that the concentrations of non-essential metalloids such as As, Cd Co, Cr, and U are usually higher in submerged plants than in other life forms. Unlike floating hydrophytes, submerged species use their leaves to uptake (heavy) metals [1]. They are able to absorb heavy metals by passive transport through the cuticle directly in

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the water column. During the passive absorption process, the attraction and movement of positively charged heavy metal ions could be achieved by both the negatively charged cutin (a waxy polymer, the main component of the plant cuticle) and pectic acid contained in the plant cell wall [23]. However, similar to most terrestrial plants, freshwater vascular plants rooted to the sediment show higher concentrations of non-essential metalloids in roots than in the shoots, indicating that the elements are mainly absorbed from the sediments rather than the waters [55, 115, 116, 137, 143]. Indeed, radio-tracer studies gave evidence that root uptake from sediments with subsequent translocation to aboveground tissues is the principal pathway for metal movement [144–146]. Hence, the element concentration of rooted macrophytes is generally proportional to the metal’s concentration in the underlying sediments when they are present in potential plant-available element pools [143], regardless of whether the plant develops emergent or submerged life forms. Nonetheless, for aquatic macrophytes that possess roots but do not have a close physical association with sediments (e.g., Ceratophyllum demersum, Lemna minor), the water is undoubtedly the principal source of elements [143]. The element concentrations in the sediments, in most cases, are derived from the settling of particulates and element absorption from the waters, and there must be, to some degree, a constant proportionality between these two phases that depends on the chemical properties of the element physicochemical water properties and the substrate. However, there is not usually a good correlation between element abundances in sediments and those in rooted aquatic plants [29], which suggests that elements are immobilized in the substrate and the soil–plant transfer depends on chemical and physical processes in the rhizosphere. Above all, sediment pH, redox potential, and organic content are three critical sediment variables that affect the phase partitioning of metals and their bioavailability [143]. Not all forms of an element are equally available for uptake by the biota (Kabata-Pendias 2001). Metals associated with various sediment fractions (e.g., adsorbed on particle surfaces, occluded in iron and/or manganese oxyhydroxides, bound to organic matter) may not be readily available for rooted macrophytes. A low pH and reducing conditions enhance metal release from Fe and Mn oxides [70, 88], making them potentially available for plants [144]. With regard to the redox conditions, the most significant metal mobility likely exists within an intermediate zone (-150 > Eh > +200 mV) between strongly reduced and strongly oxidized sediments which also influences metal accumulation in macrophytes [143, 144]. Notably, all macrophytes are known for their ability to oxygenate their substratum and can maintain gradients of several hundred mV as far as 30 cm deep [147, 148], which enhances the sequestration of metals. Sediment organic content is another factor that may directly affect metal phase partitioning away from ionic forms, making elements bound to organic matter less available to rooted aquatic macrophytes. Metals that preferentially bind to organic ligands, like Cu and Zn, should be less available to rooted macrophytes when such ligands are available for binding [144, 145]. Practically, this has substantial consequences for metal phytoextraction in CWs because in the early stage of CW operation, metal removal is mediated mainly by metal sorption to the substrate [68, 69] where Al, Fe, and Mn precipitate as insoluble compounds such as oxides,

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oxyhydroxides, and hydroxides [69], which, in turn, act as scavenger for elements including Zn, Cu, Cd, Ni, and Pb [70, 76, 87]. However, plants actively determine the chemical and physical properties of the root zone by the release and consumption of oxygen, release of protons, and various metal-chelating carboxylates such as citrate, malate, and oxalate that can mobilize elements through ligand exchange and complexation reactions [134]. There are significant differences among plant species and genotypes in their ability to release compounds via their roots into the soil, and this depends on the nutritional status, especially the form of N, Fe supply, and P nutrition status [8, 130, 131, 134, 149, 150]. In CWs, growth and biomass production are significantly determined by the availability of phosphorus [150]. It is reasonable that at least some macrophytes deploy carboxylate-based P-mining strategies [149]. Besides direct root-substrate/water interactions, the rhizosphere also provides a suitable environment to various soil microbes that directly or indirectly affect the metal removal process in CWs via biosorption, oxidationreduction of metals [151], and the release of element solubilizing secondary metabolites [68, 92]. There is evidence that microbial interactions in the rhizosphere play a fundamental role in hyperaccumulation processes [152], and anaerobic metalreducing bacteria can also govern the reduction of toxic metals to insoluble and less toxic form (Se, Cr) [151]. Rhizosphere bacteria in laboratory microcosms facilitated the uptake of Se and Hg in salt marsh bulrush (Scirpus robustus) and Polypogon monspeliensis [91]. Ha et al. [116] and Sakakibara et al. [117] reported that Eleocharis acicularis accumulates Zn, Cu, Cd, Pb, In, and Ag more than 100-fold in the shoots relative to the element concentrations in the water without markedly discriminating essential (Cu, Zn) and non-essential, economically valuable elements (Cd, In, Ag). Compared to Eleocharis acicularis, Phalaris arundinacea (reed canary grass) produces much higher biomass and is able to accumulate high concentrations of Ge, Zn, and REE in the shoots [109]. It is a perennial wetland grass native to Europe, Asia, northern Africa, and North America with a high capacity for rapid vegetative spread, deep root systems, clonal reproduction, and wide ecological amplitude regarding soil moisture and mineralization. Typically, P. arundinacea can be harvested two times per year (spring and autumn), allowing the phytoextraction of more than 100 g Cd, 50 g Ge, 80 g REE, and 1.2 kg Zn per hectare when the plants are exposed to waters with 0.7 mg/L Ge and 1.8 mg/L REE [109]. Other plant species have been most profoundly studied with regard to the accumulation of typical heavy metals, and information on the accumulation of other strategic elements is very scarce in the literature. Khellaf and Zerdaoui [104] demonstrated Zn accumulation by Lemna gibba. Abhilash et al. [153] found a high Cd accumulation by Ipomea aquatica and Limnocharis flava, and Zhang et al. [102] reported Wolffia globosa as a suitable candidate for the phytoextraction of As, which is able to accumulate more than 1000 mg As per kg. Table 3 presents an overview of suitable plant species for phytoextraction focusing on the element spectrum given in the literature. Unfortunately, most of these studies reported a very narrow range of elements in the plants without considering other elements beyond the target elements studied. Also, knowledge of the dynamics

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Table 3 Element accumulation in shoots of plant species suitable for phytoextraction. Numbers in superscript next to a species-specific pH optimum Plant species Eichhornia crassipes

Life form Free-floating

pHoptimum 5.8–6.0

Pistia stratiotes

Free-floating

6.5–7.2

Wolffia globosa

Free-floating

5.0–7.0

Salvinia sp., Salvinia molesta

Free-floating

5.0–8.0

Lemna gibba, Lemna minor

Free-floating

5.0–7.0

Myriophyllum spicatum

Perennial, submerged, rooted Perennial, submerged, rooted Perennial, submerged, rooted Submerged 1 completely rootless 2,3 rooted or freefloating Perennial, emergent

7.0–8.5

Target elements REE (La, Ce) Al, As, Ag, Ba, Cd, Cu, Cr, Hg, Mg, Mo, Mn, Ni, Pb, Zn Fe, Mn, Cr, Pb, Cu, Zn, Ni, Co, Cd As (>1000 mg/kg DW) Fe, Cd, Ni, Mn, Zn, Cu, Cr and Pb Pb, Cr, Zn; Hg, Cu, Cd, Ni, U, As, Mg, Fe Co, Cu, Ni, Zn

6.0–8.0

Cr

[106]

6.5–8.0

Pb, Cu, Cd, Zn

[107]

6.0–8.0

As

[29]

1

3.9–8.6 6.0–7.5 3 4.5–7.5 4 6.0–7.0 1 3.9–8.6 2 5.6–7.8 3 5.0–7.5

Mn, Cr, Ni, Fe, Pb, Co, Cu, Zn, Cd

[80]

Mn 1 Al, Ba, Sr, Pb 2 Fe, V, Zn, Cd, 3 Cu and Ni

[108]

1

4.5–7.5 4.5–8.0 3 6.0–7.5 4 5.0–7.0

Cd, Cr, Fe, Pb, Cu, Ni

[82]

1

Zn, Hg, Cr, As, Se

[87]

Ge, REE, Zn

[24, 109– 112]

Vallisneria spiralis Myriophyllum verticillatum Ceratophyllum demersum,1 Egeria densa,2 Lagarosiphon major3

Phragmites australis, Canna indica,2 Typha latifolia,3 Hydrocotyle umbellata4 Phragmites australis,1 Salix viminalis,2 Populus canadensis3

Typha latifolia,1 Phragmites australis,2 Ceratophyllum demersum,3 Alisma plantagoaquatica4 Typha angustifolia,1 Schoenoplectus californicus2 Phalaris arundinacea

1 Perennial, emergent 2 Phanerophyte, Shrub 3 Phanerophyte, Tree 1,2,4 Perennial, emergent 3 Perennial, submerged freefloating Perennial, emergent Perennial

2

2

6.1–7.8 6.5–7.0 4.5–8.2

Reference [1, 31, 96–101]

[95]

[102] [1, 103] [1, 103, 104] [105]

2

(continued)

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Table 3 (continued) Plant species Phragmites australis,1 Bambusa vulgaris,2 Chenopodium album,3 Amaranthus cruentus4 Phragmites australis, Chrysopogon zizanioides Cynodon dactylon Colocasia esculenta Amaranthus spinosus Cyanthillium cinereum Fimbristylis bisumbellata Mikania micrantha Diplazium esculentum Alternanthera sessilis,1 Alternanthera philoxeroides2 Alisma lanceolatum,1 Carex cuprina,2 Epilobium hirsutum,3 Juncus inflexus4

Life form Perennial and annual

Phragmites australis

Perennial, emergent Perennial, hemicryptophyte

Eleocharis acicularis

Typha angustifolia Ludwigia stolonifera Colocasia esculenta,1 Typha latifolia2 Typha latifolia,1 Phragmites australis, and Colocasia esculenta3 Vetiveria zizanioides

Perennial Perennial Perennial Annual Annual Annual Perennial Perennial Emergent, perennial Perennial hemicryptophytes

Perennial, emergent Perennial, floating, rooted Perennial, emergent Perennial, emergent Perennial, semiaquatic emergent

pHoptimum 1 4.5–8.0 2 5.0–6.5 3 7.0–8.0 4 5.5–7.0 4.5–8.0

Target elements Pb, Cs

Reference [113]

Zn, Cr, Ni, Pb

[78]

6.0–7.0 5.6–7.0 5.5–7.0 4.0–6.0 n. d. 3.6–6.5 5–6 1 6.0–7.5 2 4.8–7.7 1 6.5–7.5 2 n.d. 3 6.5–8.5 4 6.0–8.0 4.5–8.0

Pb Mg, Fe, Zn Zn Mg, Fe, Zn Mg, Pb Pb Pb, Zn Mg, Fe, Pb, Zn

[114] [114] [114] [114] [114] [114] [114] [114]

Al, As, Cd, Cu, Cr, Fe, Mn, Ni, Pb, Sn, Zn

[67]

Ge, Pd, Rh

[112, 115]

6.0–7.5

[116, 117]

6.1–7.8

In, Ag Cu, Zn, As, Cd, Pb Zn and Cd

[118]

5.5–7.0

Pb, Cr and Cd

[119]

1

Mn and Zn (TF > 1) Pb, Cu, Zn, As, Cr, Co, Ni, Mn

[120]

Zn, Fe, Cu, Cd and Pb

[122]

5.6–7.0 4.5–7.5 1 4.5–7.5 2 3.9–8.6 3 4.5–7.5 4.5–10.5 2

[121]

of element accumulation over time is a field of future research. Element accumulation depends on plant age and shows a seasonal variation [88, 121]. Young, fastgrowing roots typically have higher element uptake rates than older roots, and young plants exhibit higher translocation rates than older plants [72]. Further, element uptake rates are influenced by seasonal changes in plant metabolism and transpiration rates. Rai et al. [121] found that the macrophytes (T. latifolia, P. australis, and C. esculenta) demonstrate higher bioconcentration factor (BCF) and translocation

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factor (TF) in summer than in winter. Wetlands generally experience the most significant seasonal variability in terms of photosynthesis, respiration, and growth [72]. A typical seasonal pattern for the concentration of many elements within macrophyte tissues is an early and late season maximum surrounding a mid-season minimum. The mid-summer minimum has been attributed to the dilution of elements by carbon incorporated into structural components of rapidly growing plants [143]. However, the concentration minimum might not automatically imply a low phytoextraction efficiency when the low concentrations are compensated by a high biomass development [108].

4.3 4.3.1

Application of Phytoextraction in CWs Rooted Emergent Hydrophytes

Despite the large number of aquatic plant species that are theoretically suitable for metal(loid) phytoextraction due to their theoretical accumulation behavior observed in lab- and pilot-scale experiments (Table 3), in reality, the application of many species is limited by the abiotic growth conditions above all climate and hydrology of different CW types. CWs planted with woody plants (trees and shrubs, especially Salix spec.) are pretty seldom utilized to treat wastewater. A few functional examples are reported by the US EPA [71] and Vymazal [62]. Although some tree species from the genera Salix, Betula, and Populus efficiently accumulate a variety of metals [29, 108] and represent interesting species for zero-discharge wastewater treatment plants due to their high transpiration [124], their applicability suffers from their slow growth and long life cycle. Thus, in the first few years of life, trees may not demonstrate significant effectiveness. However, in the long term, woody metal accumulators may provide an environmentally sustainable solution to improve the long-term performance of treatment systems [29]. It has been found that Salix viminalis is highly effective in removing macroelements (phosphate, nitrate, nitrite, ammonia, chloride, sulfate, Ca, Mg, K, and Fe) and potentially toxic metals such as Cd (58–71%), V (100%), and Zn (84–92%), while Populus canadensis most efficiently assimilates Cu (49–60%) and Ni (55–67%) [108]. For saline wastewater treatment in tropical and subtropical climates, Taxodium distichum (bald cypress), Melaleuca quinquenervia (paper bark tea tree), and Kandelia candel (mangroves) can be ideal species [62]. Greenway et al. [65] used Melaleuca quinquenervia together with the herbaceous plant species Cymbopogon citratus (lemongrass) to treat highly concentrated sewage in CW systems with biochar as an amendment to sand media. Submerged life forms and floating non-rooted plants (Eichhornia crassipes, Lemna spp.) cannot be used in SSF-CWs. In contrast, emergent substrate-rooted macrophytes or terrestrial plants can still be used in any CW type, including hydroponic systems. Given that the species from the genera Phragmites, Typha, Scirpus, Juncus, and Phalaris form by far the highest biomass and show strong metal

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tolerance, they are most profoundly used for phytoextraction applications. Of these genera, Phragmites australis and Typha latifolia are among the most used plant species for the removal of metals from wastewater in CWs [70]. These species are perennial, which is an essential factor for their metal uptake rate per unit area of the wetland [70]. Even when the concentrations of non-essential metalloids in aboveground biomass are relatively low and often do not reach the concentration thresholds for hyperaccumulators [20, 21], the plants still achieve high phytoextraction rates due to their high biomass, which justifies their widespread application in many free surface flow and subsurface flow CWs for Fe, Cd, Zn, Ni, Pb, and Cr removal [80, 82, 121]. These species were successfully tested for the remediation of landfill leachates [154] and wastewater from mining industries and municipal wastewater [155], where they achieved removal efficiencies of 96%, 96%, 95%, 85%, and 80% for Al, Cd, Cr, Zn, and Pb. There is evidence that Typha angustifolia and Typha domingensis preferentially accumulate non-essential elements in the roots/rhizomes, and only a minor fraction of metals is translocated to the shoots [118, 156]. Notably, results on root–shoot translocation may be strongly biased by overestimated root concentrations relative to shoot concentration because metal ions and substrate particles contained in the wastewater may be absorbed onto the root surface and apoplast (Han et al. [157]). With regard to the economically valuable, strategic elements, Bonanno [115] showed that Phragmites australis accumulates rhodium, a platinum group metal with high economic relevance. Plant samples collected along a riverside affected by massive urbanization and intensive agriculture contained concentrations of 1.11–1.13 mg/kg. The subsurface flow phytoremediation systems plated with Phragmites australis demonstrated a high efficiency of removing Al (up to 97%), Ba (95%), Pb (81%), and Sr (51%) from sewage [108]. Also, Phragmites australis and other species from the functional group of grasses, such as Phalaris arundinacea, are considered accumulators of Ge, representing meaningful agents for raw material recovery from wastewater streams [110]. When grown in dilute solutions containing 0.7 mg/L Ge, Phalaris arundinacea accumulated 37 mg/kg Ge [109]. In tropical climates, Cyperus alternifolius has been reported as a very promising species for metal phytoextraction. Due to the high nutrient demands to achieve their fast growth rates, these plants accumulate high amounts of N, P, Fe, and Mn and, consequently, elimination rates of 47%–99% considering many different wastewater types (textile, pharmaceutical, agricultural, and organic discharges). A comparison of three plant species (Cyperus alternifolius, Cynodon dactylon, and Typha latifolia) in a pilotscale VFCW for the removal of potentially toxic metal(loid)s from the wastewater revealed the highest phytoextraction potential for T. latifolia, followed by C. alternifolius and C. dactylon [158]. Mal and Rangabhashiyam [123] successfully employed Iris pseudacorus, Scirpus americanus, Canna indica, and Stenotaphrum secundatum in CWs for metal removal. More recently, Moogouei and Chen [113] demonstrated that combinations of aquatic and terrestrial plants (Amaranthus cruentus, Bambusa vulgaris, Chenopodium album, and P. australis) in hydroponic CWs can be a promising approach for the phytoextraction of Pb and Cs. Saleh et al. [119] concluded that L. stolonifera can be considered a phytoremediation agent with

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high efficiency in the remediation of wastewater containing toxic heavy metals like Pb, Cd, and Cr. Yasar et al. [159] compared the potential of Phragmites karka and P. stratiotes plants in vertical and horizontally constructed wetlands systems for wastewater treatment. They found a higher P and Cl removal efficiency of P. stratiotes than for Phragmites karka. The vertical system cultivated with P. stratiotes exhibited a removal percentage efficiency of 95% for phosphate and 51% for chloride, highlighting the high potential of non-rooted floating species in element phytoextraction, which is a consequence of their physiological capacity for element uptake and the direct contact of the roots with the elements dissolved in the water. Ali et al. [80] have demonstrated a sufficient efficiency of HCWs (VSSF-CW and serially connected phyto-treatment ponds) in municipal wastewater purification to reuse the purified water for crop irrigation with a sustainable level of environmental security in water deficiency regions. This research used Phragmites australis, Canna indica, Typha latifolia, and Hydrocotyle umbellata), which are relatively effective in metal extraction from wastewater. Rana and Maiti [120] investigated the potential of Colocasia esculenta and Typha latifolia for removing heavy metals from municipal wastewater. The genotypes used in this study exhibited translocation factors 1 for Mn and Zn.

4.3.2

Floating Hydrophytes

Although many floating non-rooted macrophytes have been extensively described as the most effective element accumulators [1], CWs planted with free-floating hydrophytes are not widely used in practice. One of the main reasons is high maintenance and operating expenses related to constant plant removal, as well as subsequent proper disposal [62]. The sizes of free-floating hydrophytes vary from large plants with massive leaves and roots, such as Eichhornia crassipes (water hyacinth) and Pistia stratiotes (water lettuce), to tiny plants, such as Lemnaceae (duckweeds, e.g., Lemna spp., Spirodela polyrhiza, or Wolffia spp.) [62]. High productivity and growth rates are distinctive features of free-floating hydrophytes. However, Eichhornia crassipes and Pistia stratiotes are invasive species that are considered some of the most prolific free-floating macrophytes in the world [29, 150] and are frost-sensitive, so they cannot survive in temperate and cold climates [62]. Conversely, Lemnaceae species are light-frost tolerant and consequently dwell in a much wider geographic area. Duckweeds can be found in any free surface flow CWs because they are naturally and easily transported by wind and birds [62]. Aside from these ecological considerations, Gemeda et al. [125] investigated the phytoremediation capability of three floating hydrophytes – water lilies, duckweed, and water hyacinth (E. crassipes) – for the removal of trivalent and hexavalent chromium in aqueous solutions and demonstrated the highest Cr accumulation in E. crassipes (82 mg CrVI and 323 mg CrIII per kg plant biomass) when the plants were grown in artificial aqueous solutions containing 10 mg/L Cr(III) and Cr(IV). Rezania et al. [98] assessed the performance of free-floating macrophytes in different

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CW types (P. stratiotes and E. crassipes). They revealed that E. crassipes is suitable for treating industrial wastewater. A comparison in the accumulation of Fe, Cu, Cd, Cr, Zn, Ni, and As in Pistia stratiotes, Spirodela polyrhiza, and Eichhornia crassipes identified E. crassipes as the most efficient candidate for the accumulation of selected heavy metals followed by P. stratiotes and S. polyrhiza [101]. Here, E. crassipes removes more than 79% of the metals from wastewater through phytoextraction. In addition, E. crassipes was found to accumulate REE, especially La and Ce. The concentration of Ce in E. crassipes was 74,000 times higher than its content in the water of electric power plant reservoirs [31]. In a pilot-scale CW, Akowanou et al. [160] assessed the efficiency of the combination of three floating hydrophytes – E. crassipes, P. stratiotes, and Lemna minor – for domestic wastewater. The study revealed that such a combination of free-floating aquatic plants resulted in a significant reduction in sewage contamination levels (based on five indicators). The pond with water hyacinth showed the highest elimination rate of organic carbon pollution (COD, BOD5), while the pond with P. stratiotes was most effective in removing nitrogen (TKN), and the duckweed pond significantly removed phosphorus. Hence, the combination of different floating wetland plants for domestic wastewater purification can enhance the overall biological treatment performance. Abhayawardhana et al. [103] investigated the phytoextraction of metals (Cr, Cu, Fe, Ni, and Pb) and nutrients (TN and TP) from municipal wastewater using S. molesta and L. gibba plants. These free-floating macrophytes showed quite similar high removal and accumulation abilities along with significant BFs ranging between 618 and 870, which proves that they are suitable candidates for the posttreatment of sewage. Thus, S. molesta demonstrated the average uptake efficiency of TN – 73.3%, TP – 72.6%, Cr – 81.6%, Cu – 69.8%, Fe – 65.2%, Ni – 66.3%, and Pb – 74.8%. The average extraction efficiency of TN, TP, Cr, Cu, Fe, Ni, and Pb shown by L. gibba was 62.1%, 77.2%, 86.9%, 69.7%, 73.1%, 61.8%, and 85.7%, respectively. Unlike free-floating wetland plants, floating-leaved hydrophytes are able to absorb various elements from bottom sediments and substrates where they are rooted. As a rule, these aquatic plants have large and branched rhizomes, which are connected to floating leaves via long peduncles. However, very few CWs and ponds have been designed with floating-leaved hydrophytes. At least one CW reported in the literature has been planted with spatterdock (Nuphar lutea) and designed for the tertiary treatment of sewage from the city of Orlando, Florida, with a population equivalent (PE) of 800,000. This floating-leaved plant CW is located in Iron Bridge, Florida. The most widespread species of this type of hydrophyte are water lilies (Nymphaea spp.), spatterdock (Nuphar lutea), and Indian lotus (Nelumbo nucifera) [62]. Watercress, a rooted floating-leaved plant (Rorippa nasturtium subsp. aquaticum), was found to accumulate arsenic at a concentration of more than 400 mg/kg (dry weight), especially when As is present As(V), whereas As(III) is less available for plant uptake [29].

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4.3.3

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Submerged Hydrophytes

Due to their life form, the application of submerged hydrophytes is restricted to free water surface CWs or for the treatment of wastewater with relatively low turbidity. Submerged hydrophytes are sensitive to a high concentration of suspended solids (SS), limiting the penetration of sun rays (photosynthetically active radiation) needed for photosynthesis. In this regard, they are recommended for use in tertiary wastewater treatment. Another limiting factor of the use of CWs with submerged hydrophytes is the need for sufficient dissolved oxygen content, as this life form prefers well-oxygenated waters [62]. The most commonly used submerged hydrophyte in full-scale artificial wetlands is watermilfoil (Myriophyllum spicatum). It has been used in a free surface flow CW in Montréal, Canada. There are also known examples of the use of naturally occurring (local) species of hydrophytes, such as southern naiad (Najas guadalupensis) and coontail (Ceratophyllum demersum) in the CWs of the Florida Everglades Stormwater Area [62]. Submerged hydrophytes are naturally covered with periphyton. Periphyton is an assemblage of algae that has many beneficial effects on organic contamination elimination. Autotrophic algae produce the oxygen needed for the oxidation of different pollutants and nutrient uptake. Nonetheless, the excessive periphyton growth may considerably hinder the photosynthesis of submerged macrophytes, blocking the penetration of photosynthetically active radiation [62]. Myriophyllum verticillatum tolerates high concentrations of Pb, Cu, Cd, and Zn and probably hyperaccumulates Pb [107]. Metal uptake tests with Vallisneria spiralis revealed a high Cr absorption, although Cr toxicity eventually negatively impacted photosynthesis [106]. Lesage et al. [105] demonstrated a high accumulation potential of Co, Cu, Ni, and Zn in Myriophyllum spicatum and successfully employed the plant for the remediation of industrial wastewater. Ceratophyllum demersum (hornwort), Egeria densa (Brazilian waterweed), and Lagarosiphon major (curly waterweed) are known for their strong capability for extraordinary accumulation of arsenic originating from geothermal activity in New Zealand [29].

4.3.4

Phycoextraction

Algae is an informal term used to refer to an ancient large and diverse group that includes both photosynthetic prokaryotic cyanobacteria (blue-green algae) and eukaryotic organisms of different sizes: microalgae (unicellular) and macroalgae or seaweed (multicellular). Algae are ubiquitous and unique bionts that can live in both aquatic and terrestrial ecosystems, in both fresh and sea waters, and they are also widespread in plant tissues as a member of beneficial microbiome or phytobiome [161]. These diverse lower plants have existed for more than 2 billion years. That is why some algae are extremophile organisms that are able to survive and even thrive in extreme conditions (e.g., extremely acidic or alkaline media, high temperature, and metal concentration). These properties of extremophile algae make them suitable

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for the treatment of industrial and mining-related acid effluents (AMD) containing high concentrations of non-essential and potentially toxic elements, such as REEs. Because of the low ratio between carbon and nitrogen (C/N < 0.2), acid drainage from REE tailings is highly resistant to biodegradation [48]. This noteworthy fact has significantly limited the use of heterotrophic organisms such as bacteria (e.g., Bacillus subtilis, Leisingera methylohalidivorans, and Phaeobacter inhibens), which were initially used for REE biosorption in acidic environments. Thus, heterotrophic bacteria-based REE recovery technologies are energy-intensive methods because of extensive organic carbon consumption [52]. In this regard, photoautotrophic and acidophilic algae for metal recovery (phycoremediation) are considered to be reasonably promising, flexible, and affordable [46, 52]. General element phytoextraction in algae is mediated by element absorption onto cell walls and cellular uptake [162]. The algal cell walls comprise functional hydroxyl (–OH-), carboxyl (–COOH), amino (–NH2), oxygen-containing, and sulfated groups that act as active sites for metal binding [46, 162]. Thus, algae represent highly efficient biosorbents for the removal of a variety of metals [123]. In addition, algae excrete various peptides that are involved in the extracellular complexation of metals and protect the algae from metal toxicity [163]. Saavedra et al. [164] demonstrated efficient removal of As, Cu, Mn, and Zn using Scenedesmus almeriensis and Chlorella vulgaris and Mal and Rangabhashiyam [123] used Scenedesmus obliquus for phycoextraction of Cd. Aleissa et al. [165] demonstrated the uptake and removal of U by filamentous green algae such as Spirogyra and Cladophora spp., and the two acid-tolerant microalgae Desmodesmus sp. and Heterochlorella sp. are able to accumulate Fe, Zn, and Mn at pH 3.5 [166]. For phytoextraction of REEs from mining effluents, algae-based technologies must be effective not only in recovering REEs but also in coping with high concentrations of ammonium nitrogen (NH4+-N), which derives from the chemical leaching of REE ores with ammonium salts in situ and causes unavoidably contamination of water bodies [73]. Some species of acidophilic microalgae have been successfully used for deammonification [52]. Zhang et al. [53] isolated the multi-tolerant co-flocculating microalgae Scenedesmus sp. and Parachlorella sp. from harsh REE tailings wastewater that are able to remove more than 90% NH4+-N. However, the role of Scenedesmus sp. and Parachlorella sp. in the biosorption and bioaccumulation of REEs remains unclear. Despite the superior properties of extremophilic microalgae, there is a significant hindrance to the use of acidophilic microalgae technology for efficient REE removal and recovery. This limitation is related to the positive charge on their surface (2–3 mV zeta potential), which definitely hinders the adsorption of positively charged REEs [167]. The trivalent state (+3) is typical of most REEs, except for Ce (+4) and Eu (+2) [11, 50]. REE organic complexes dominate when pH is 4 ~ 8 [46]. Thus, the synchronous recovery of a large amount of positively charged REEs and ammonium removal from AMD is a topical challenge [52]. One of the most effective solutions is calcium alginate-based immobilization of microalga or other microorganisms. This strategy has been successfully used to functionalize the surface of microorganisms possessing specific functional groups or surface charges.

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The cell immobilization potentially provides a practical path for the simultaneous biosorption of REEs and bio-removal of ammonium [52]. Sun et al. [52] demonstrated the high efficiency of Galdieria sulphuraria (acidophilic red alga) immobilized by calcium alginate (C12H14CaO12)n) in simultaneous removal of REEs and ammonium nitrogen from synthetic REE tailings drainage. The recovery percentages of lanthanum, yttrium, and scandium were 97%, 96%, and 99%, respectively. The removal of REEs took place via both physical and chemical adsorption. Physical adsorption was caused by the affinity of algal functional hydroxyl and carboxyl groups; meanwhile, chemical adsorption was based on the ion exchange of calcium cations (Ca2+) with La, Y, and Sm. It is important to note that the positively charged G. sulphuraria is ideal for the bio-removal of negatively charged platinum coordination compound – [PtCl6]2-. Moreover, hexachloroplatinate (2-) adsorbed by extracellular polymeric substances (EPS) and accumulated in algae cells is possibly transformed to platinum chloride (PtCl4). This assumption is based on an increase in the zeta potentials of EPS and algae cells [167]. Possibly, the phycoremediation potential of algal-based techniques could be improved by the addition of carbonate, simulating an increase in carbonic anhydrase (CA) activity that catalyzes carbon monoxide conversion or reversible hydration of CO2, leading to the formation of bicarbonate. Bicarbonate (HCO3-) is known to enter the cells of autotrophic microalgae through active transport, while CO2 enters the cells through diffusion. Consequently, the carbonate contributes to rapid biomass growth, higher production of extracellular polymeric substances, and element removal [48]. Jacinto et al. [168] used the living macroalgae Gracilaria gracilis (red algae) for the bioaccumulation of Y, Ce, Nd, Eu, and La. They demonstrated a high metal tolerance to both mono- and multi-element saline solutions of 500 μg/L of Y, Ce, Nd, Eu, and La, of which virtually 100% of all elements have been recovered. Similarly, Pinto et al. [51] studied REE accumulation in a mixture of living green, brown, and red macroalgae using Ulva lactuca, Ulva intestinalis, Fucus spiralis, Fucus vesiculosus, Osmundea pinnatifida, and Gracilaria sp. They found more than 1295 mg/kg REE in the mixed biomass (bioconcentration factor: 2500) without visible signs of stress, highlighting the high potential of living macroalgae for REE phytoextraction from strongly saline and acidic industrial effluents, including AMD.

4.3.5

Enhanced Phytoextraction

Besides the appropriate selection of plant species, phytoextraction efficiency is governed by the availability of metal(loid)s in water and growth substrate. Plant availability of any element depends on substrate/water-associated factors (pH, organic matter content, sorption capacity), element-associated factors (chemical forms, solubility, toxicity), plant-associated factors (root exudates, root morphology, capacity for element uptake, and root–shoot translocation), and microbial associated factors (microbial growth factors and metabolites involved in pathogen defense and element acquisition). Plant availability initially depends on solubility and chemical

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speciation, as plants can only take up elements dissolved in ionic forms. Indeed, the uptake of complexes has been scarcely reported in the literature, except for Fe. In wastewater, a large fraction of elements may appear as colloids and complexes, or absorbed onto the substrate, which ultimately limits the solubility and availability to the plant. Consequently, all chemical and physical changes in water parameters affecting element solubility (e.g., pH, Eh) and the presence of free ionic forms support element availability and uptake. When the water and rhizosphere properties principally allow radial transport and uptake for some non-essential elements/ions with high charge density (e.g., Pb, REE), root–shoot translocation can be restricted by interactions with cellular constituents lowering the element concentrations in aboveground plant parts, which is especially problematic when perennial rooted macrophytes are used, and certain parts of the shoots solely form harvestable plant parts. During the treatment of waters rich in non-essential metals but with a deficiency in nutrients, the application of appropriate doses of fertilizers can enhance phytoextraction through enhanced plant metabolism, growth, and biomass formation. The approaches related to enhanced metal(loid) mobility may target the optimization of water pH [70, 88] and Eh to reduce oxide formation and increase solubility. Other approaches for enhanced phytoextraction include the application of various complexation agents, substrate inoculation with microbes, using genetically engineered species, and co-cropping of species. The literature indicates that chelators such as EDTA, HEDP, NTPM, DTPA, EDTMP, DFOB, AGTA, HEDTA, and EDDNA and organic acid anions effectively mobilize a variety of metals, including Cd, Zn, Pb, Ge, and REE, through the formation of organometallic complexes ([169]). Despite its effectiveness, chelant-enhanced phytoextraction did not receive much attention in practice because of the compounds’ high prices, potential environmental hazards, and negative effects on plant growth [22]. The application of salicylic acid (SA) has been found effective in alleviating metal stress in the plant, resulting in enhanced phytoremediation potential of plants [9], but the effectiveness in CWs remains a field of future research. The inoculation with specific isolates of microbes can improve nutrient supply [10], plant growth, and element accumulation [85, 91, 151] through the production and release of growth-promoting hormones, alleviation of ROS stress, pathogen defense, and other secondary metabolites involved in element acquisition (metallophores) [92]. Microbial metallophores form very stable chelate complexes with a variety of elements, including Fe, Mn, Zn, As, Cd, Pb, Ge, and REE, thereby increasing their mobility in the growth substrate. Given that these bacteria preferentially appear in the root zone of plants where they benefit from rhizodeposition, the mobilization of metalloids by PGPR is restricted to the sites of uptake and thus more efficient than the artificial addition of chelates [109]. There is evidence that many hyperaccumulation phenomena across different taxa appear to be highly influenced by microbial interactions in the rhizosphere [30]. For example, the inoculation of Bacillus subtilis has improved the Cd phytoremediation efficiency of Medicago sativa by 139%, while the plant biomass has increased by 29% [85]. Other approaches use selected genotypes of specific species that possess certain functional traits involved in metal(oid) uptake, translocation, and tolerance. Nicotiana tabacum and Arabidopsis thaliana are examples of

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transgenic plants that have been designed for the enhanced phytoextraction of cadmium and mercury [1]. Wiche et al. [138] demonstrated that the accumulation of REEs, Fe, and Zn in Phalaris arundinacea depends on genotypic variation in element acquisition and uptake, which form the basis for the breeding of novel strains for phytoextraction purposes. Finally, enhanced phytoextraction can also be achieved by the application of mixed cultures in CWs. Given that different plant species accumulate a certain spectrum of metals (Table 3), the application of mixed cultures not only improves the ecological value and overall stress resilience of the vegetation [130, 131] but also improves phytoextraction efficiency through a broader spectrum of elements, a higher diversity in growth, and element uptake across the changing environmental conditions throughout the growth season and enhanced element availability through interspecific root interactions. The literature indicates that interspecific root interactions of grain legumes grown in mixed culture with grasses increase the mobility and plant availability of essential (Fe, Mn, P, N) and non-essential elements (Ge, REE, Cd, Th, U, Pb, As). Here, the legumes released large quantities of carboxylates and acidified the rhizosphere, which led to a higher metalloid accumulation in grasses that lack this ability [130, 131]. It is reasonable that hydrophytes with contrasting below-ground functional traits related to element acquisition also contribute to enhanced phytoextraction in CWs. Besides the fundamental role of root exudates, there is evidence that in hydrophytes, the element availability can be limited by the formation of iron plaques at the root epidermis, representing a barrier for element uptake, especially of non-essential elements with a high charge density [14, 170]. However, compared to terrestrial model plants, information on the physiological root traits of hydrophytes involved in element acquisition is very scarce in the literature. Therefore, the elucidation of processes in the rhizosphere of hydrophytes represents an important field of research in the future, enabling an appropriate selection of species and genotypes to achieve optimal phytoextraction rates through rhizosphere engineering.

5 Element Recovery from Biomass The harvest of biomass from CWs is an essential step in metal(loid) phytoextraction to avoid the liberation and recycling of elements with microbial litter degradation. At the same time, biomass represents an economically valuable secondary raw material for the production of energy and the recovery of elements as raw materials in the spirit of a phytomining process chain. This concept requests advanced and environmentally friendly biomass processing technologies, extensive laboratory and fullscale experiments to compare the technical efficiency, applicability and economic feasibility of recycling methods, extra capital and operating costs, additional time, and the development of post-phytoremediation biomass management policy [171]. Nevertheless, with the development and improvement of phytomining technologies [128, 172], the need for the recycling of post-phytoremediation metal-rich biomass can be turned into an economic benefit, especially in the case of highly

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commercially valuable elements such as Ni, REEs, Ge, Ga, Au, and PGM [20, 21, 28, 128, 129]. Given that hydrophyte biomass does not significantly differ from those of terrestrial plants in terms of the element extraction methods needed for the recovery of raw materials, all well-established (pyrometallurgy, hydrometallurgy, biohydrometallurgy) and innovative techniques (ionic liquid, mechanochemical, supercritical fluid, electrochemical) of extracting target element from conventional bio-ores are generally applicable [20, 21]. Along this process chain, energy production by biomass incineration or anaerobic fermentation is a critical step to minimize the volume of material and concentrate the elements in the residues (digestate, biochar, and ashes) [173]. Concomitantly, enrichment processes such as incineration or digestion facilitate the cracking of cell structures involved in element sequestration within the plant tissues, simplify the extraction process and reduce transportation and equipment costs. At the same time, the energy produced is an important economical factor to balance the expenses of the element extraction process. There are several enrichment techniques, from simple and low-cost to complex and expensive. The desired results of the enrichment process depend not only on the applied technique and its technical parameters but on other factors: feedstock quality, plant species, co-feedstock (manure, food and crop waste), element properties (oxidation state and metal/metalloid speciation features), and local condition (need for bioenergy, heating, bio-fuel, biofertilizer). Nevertheless, in the context of phytomining, it is important to choose a method with the highest biomass enrichment (rather than the biomass reduction factor) and the most negligible impact on the environment. These processing techniques can be divided into mechanical, biological (predominantly microbiological), and thermochemical or heat processing (Sect. 5.1).

5.1 5.1.1

Element Enrichment Techniques Compaction or Pressing

Compaction can be defined as a mechanical and intermediate method of rapidly diminishing phytomass volume and increasing metal concentrations in compacted products. The compaction system must be equipped with a robust press and a sealed filtrate collector. The pressing and leachate collecting systems must be hermetically connected to prevent the leakage of filtrate since it can contain toxic water-soluble forms of metals/metalloids. The principle is quite simple – the biomass is compressed to reduce water content and produce the enriched compacted product, and the resulting leachate is collected separately in a collection container. However, there is no understanding of how effective (percentage of biomass reduction, metal enrichment level in compacted products, the volume of filtrate, etc.) and expedient (applicability, price, payback period) this system is [20, 21, 23]. Obviously, the reduction in biomass volume is not as significant as in the case of using thermal methods, and the resulting compressed products need to be reprocessed. It is possible

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that compaction could be useful as a mobile system for rapid pre-reduction and pre-enrichment of biomass on a commercial scale.

5.1.2

Biological Enrichment or Microbial Treatment

Microbiological treatment includes two traditional techniques of organic waste material degradation: composting and anaerobic digestion [174], which can also be suitable for phytomining as a plant biomass enrichment stage [20, 21, 173]. These traditional techniques are based on the ability of microorganisms (bacteria, fungi, archaea) to decompose organic matter under both aerobic and anaerobic conditions [171]. The use of macro-organisms (earthworms) is also possible. In this case, the process and end-product are called vermicomposting and vermicompost, respectively. Composting is an aerobic, self-heating, solid-phase process of decomposing biodegradable organic materials. The primary aim of composting is the stabilization and humification of organic matter. The primary end-product is compost – a beneficial humus-like substance. It is usually used as a nutrient-rich soil amendment or fertilizer for food production. Besides, compost is an ecologically sustainable alternative to synthesized fertilizer since composting saves water and energy used to produce modern commercial fertilizer [174]. However, in the light of phytomining, metal-rich compost may be considered a “bio-ore” and used for the extraction of commercially valuable elements. The application of compost as a fertilizer is possible if CWs perform a polishing wastewater treatment step and metal concentrations in hydrophytes are insignificant. Generally, composting can be identified as a sustainable organic waste management technology in the food–energy–water nexus [174]. Nevertheless, currently, there is a lack of information on the feasibility, applicability, and efficiency of composting as a microbiological enrichment technique for “exotic” element (noble metals, REEs) phytomining purposes in order to draw firm conclusions about its feasibility or unsuitability [20, 21]. For instance, composting has been investigated as a preprocessing stage before the disposal of metalloid-rich biomass in landfills. Cao et al. [175] studied the reduction in the biomass of the As hyperaccumulating fern (Pteris vittata, 4,600 mg/kg As) and arsenic biochemical transformation during composting. Composting reduced the fern biomass by 38%. Arsenic losses from biomass occurred primarily through the compost leachate formation and, to a lesser extent, through arsenic volatilization [175]. Thus, composting demonstrates relatively low biomass reduction. Other disadvantages of the composting technique include a long composting period (2–4 months), expensive equipment, the need for strict monitoring, unpleasant odor, greenhouse gas emissions, leachate formation, and the leaching of water-soluble metals/metalloids [20, 21, 171, 174]. On the other hand, in the context of phytomining, the metal-rich leachate could be a raw liquid material for element extraction (precipitation, filtration). Anaerobic digestion is an efficient four-stage microbiological process of anaerobic decomposing organic matter into biogas (biomethane) and digestate (residue rich

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in nitrogen, phosphorous, and micronutrients) under anoxic conditions. Digestate can also be used as a highly beneficial soil amendment [174]. Furthermore, in the context of world population growth, fossil fuel depletion, and renewable bioenergy priority, this process has a significant advantage over aerobic composting. Zaffar et al. [173] investigated the influence of mesophilic and thermophilic digestion on the enrichment and fractionation of typical plant nutrients (P, Mn, Fe), potentially toxic metals/metalloids (Cd, Cr, Pb, As, Cu, Co, Zn) and commercially valuable elements (Ni, REE, and Ge). They demonstrated that all trace elements contained in the biomass of Phalaris arundinacea were significantly enriched during the fermentation process (Ge: 193% and REEs: 90%). Disadvantages of this biotechnology include the complexity of the processes underlying anaerobic digestion, as well as gaps in knowledge and understanding of optimal conditions. Generally, microbiological treatment is not yet as widespread as thermal methods for biomass enrichment since the reduction in harvested biomass volume is much less (approximately 38% versus 90%). Nonetheless, the combination of phytoextraction, bioenergy generation, and element recovery from digestate is a promising environmentally sustainable technology.

5.1.3

Thermal Enrichment Methods

Heat processing is most frequently used to decompose phytoremediation biomass, as these methods demonstrate the most significant reduction in biomass (more than 90%). Another advantage of heat enrichment is a much shorter residence (processing) time, from a few seconds to several hours. Thermochemical techniques are most frequently used due to various operating temperatures and regimes as well as working environments (media). Thermal methods include pyrolysis, gasification, combustion, and ashing. Also, there is a very similar group – hydrothermal techniques, including carbonization, liquefaction and gasification. The main differences between standard thermal treatment methods are lower temperature, higher pressure, and unique working media (subcritical water – carbonization, liquefaction, supercritical water – gasification) [171]. A significant drawback of thermal methods may arise from element losses through evaporation, which can be reduced through reduced temperatures. Pyrolysis is the thermochemical decomposition of biomass feedstock at moderate (300–800 °C) to high temperature (800–1300 °C) in an inert atmosphere (N2, Ar) or an aerobic environment [20, 21, 23]. The pyrolysis products are liquid (bio-oil), syngas (e.g., H2O, H2, CO, CO2, CH4, and other light hydrocarbons), and solid residual (biochar). Traditional electric pyrolysis is classified into three groups: (1) slow, (2) fast, and (3) flash or ultra-fast. This classification is based on pyrolysis technological parameters: temperature, heating rate, and residence time. The yield of the desired product can be adjusted using the main operating parameters. In the case of the subsequent recovery of valuable metals and metalloids, biochar is the most desirable and “convenient” product, as it is produced at the lowest operating

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temperatures and contributes to the highest yield of solid residual. Here, the main aims of pyrolysis are to maximize valuable element concentration in biochar and to minimize contaminant concentration in liquid and gas fractions [20, 21]. However, pyrolysis is a complex and costly technology that requires sophisticated equipment, which must provide and maintain an inert atmosphere, significant capital investment, and operation costs. Besides, there is a special requirement for initial biomass feedstock, that is, low moisture content of less than 10% [23], which can be challenging for hydrophytes with an average water content of 84% [176], requiring dewatering as a preprocessing step. Unlike pyrolysis, incineration is carried out in excess of oxygen or air at average temperatures from 350 to 950 °C. Currently, combustion is considered to be the most reasonable, competitive, and effective method of biomass enrichment for phytomining purposes since biomass reduction is more than 90%, and the metals are mainly concentrated in incineration residue or bottom ash [171]. However, the thermoscopic techniques may result in greenhouse gas, carbon monoxide (CO), and nitrogen oxide (NOx) emissions [20, 21]. In addition, high chloride content can result in element losses through the formation of anhydrous volatile chlorides, highlighting the need for future research. Ashing is very similar to the incineration process. However, the ashing of biomass is carried out at lower temperatures (300–550 °C), which saves the energy required to heat the green feedstock. Several studies proved the efficiency of the ashing technique in enriching solid residue with REEs [177, 178]. Due to ashing at the temperature of 500 °C for 2 h, the content of REEs increased by 7.9 times from 2032 mg/kg in the fern biomass to 15,956 mg/kg in the ash. According to the research results, 93% of REEs contained in the fern Dicranopteris linearis were transformed into ash [178]. Dicranopteris linearis is deemed to be a terrestrial hyperaccumulator of REEs as their content in the fern aerial parts can be more than that in common low-grade ores. The potential of Old World forked fern for phytomining was investigated by Jally et al. [177]. Based on their results, the concentration of REEs in the bio-ore (the ash after combustion at 550 °C for 3 h) was 30,000 mg/kg, which was about 11 times greater than that in the woody biomass (2600 mg/kg). Gasification is another thermal decomposition process used for heating and converting metal-rich biomass into solid residue and syngas, which is a mixture of mainly carbon monoxide and hydrogen gas. Gasification is carried out under oxidative conditions. Air, steam, or oxygen is used for partial oxidation of organic compounds. The working temperature of gasification is usually between 550 and 900 °C. After the introduction of gasifying agents, e.g., air or carbon dioxide, the temperature rises from 1000 to 1600 °C. It has been determined that the biomass reduction level decreases in the order of incineration > gasification > pyrolysis, showing a relatively high efficiency of gasification compared to pyrolysis [23]. Syngas, as a functional end-product of gasification, is commonly used to generate electricity through gas turbines or fuel cells. Also, the main syngas components, CO and H2, can be converted into valuable chemicals or fuels. To date, information

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on the fate of economically valuable metal(loids) such as Ge, REE, and PGM during thermochemical transformation is very scarce. Duan et al. [179] investigated the transformation of As in the ladder brake fern (Pteris vittata) biomass during the gasification process. They found that about 60% of As was released at 600 °C, which can pose a severe ecological threat.

5.2

Extraction of Elements from Enriched Plant Biomass

Compared to enrichment technologies, the sequential or direct extraction of elements is much less investigated. As a rule, research on the extraction of valuable elements is focused on various low-grade minerals, wastes, or secondary resources. Dinh et al. [21] found out that REEs have been bioextracted from waste electrical and electronic equipment (WEEE) shredding dust (La, Nd, Y, Ce, Eu) using the bacterium Acidithiobacillus thiooxidans, spent fluid catalytic cracking catalyst with the bacterium Gluconobacter oxydans, fluorescent phosphor powder (Ce, Eu, Gd, La, Tb, Y) using the bacterium Komagataeibacter xylinus, red mud or bauxite residue with the fungus Penicillium tricolor, coal by-products (All REEs except Sc, Pm) using the Escherichia coli strain and Arthrobacter nicotianae bacteria, and aqueous solution (Nd, La) with the algae Sargassum sp. All these studies primarily focused on biomaterials for leaching and REE extraction and could be easily expanded to plant biomass from terrestrial and aquatic environments. Indeed, there are a few good examples of a full-cycle phytomining of REEs [180– 182]. Jally et al. [177] successfully separated and extracted REEs and Al from the combustion ash from fern biomass. Chour et al. [180] used the ion exchange leaching technique to recover REEs from plant biomass by treating biomass of Dicranopteris linearis with a 0.5 M nitric acid solution in the presence of exchange resin for the sorption of REEs. The REEs adsorped were eluted with 3 M nitric acid. The overall efficiency of REE extraction from the fern biomass was 78%. The comprehensive hydrometallurgical method for extracting REEs from the same REE hyperaccumulator fern was applied two years later [181]. The developed technology was based on leaching, precipitation, and calcination. Initially, the harvested and dried fern biomass was treated with a strong acid (0.25 M H2SO4) or a potent chelating agent (EDTA). After the extraction-leaching step, REEs were precipitated by oxalic acid (pH 2.6). Under optimal conditions, the REE precipitation was more than 90%. At the final stage, the residue was converted into rare earth oxide as a valuable end-product via calcination at 700 °C for 2 h. Eventually, after calcination, the end solid REE oxide contained about 69% of the REEs present in the original fern biomass. The hydrometallurgical approach to recover REEs from ferns naturally growing on former mine tailings was first applied by Laubie et al. [182]. Direct leaching was achieved using EDTA, followed by REE precipitation with oxalic acid. The overall REE recovery yield was around 70%.

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The recovery of germanium from combustion ash can be carried out through acidification of Ge-rich ash and distillation with hydrochloric acid (HCl). As a result, GeCl4 is captured in an aqueous phase and then the changes in the pH value result in the yields of the target product – germanium dioxide (GeO2), which is a highly commercially valuable material on the modern market [57]. Krisnayanti et al. [172] proposed to extract silver (Ag) and gold (Au) from dried tobacco biomass. The concentrations of Ag and Au in the dried plant material were 54.3 mg/kg and 1.2 mg/ kg, respectively. Ashing at approximately 300 °C was used as the biomass enrichment step. The ash was then mixed with borax (Na2B4O7 · 10H2O) and silver as a collector metal. To extract metals, a mixture of the ash and reagents was smelted at high temperatures, and the noble metals contained in the enriched ash were converted into bullion.

6 Conclusion Phytoremediation in CWs enhances water security level and biodiversity and provides sustainable environmental protection and high efficiency in the elimination of nutrients and potentially toxic metal(loid)s. Depending on the composition of wastewater, the biomass used for the removal of elements contains high concentrations of nutrients (N, P) and an array of economically valuable elements, including REE, Ge, Ni, Au, Ag, and PGM. Thus, biomass from CWs must be considered a valuable raw material for bioenergy production (biomethane, bio-fuel), nutrient-rich bio-fertilizers, and precious elements recovery. Nowadays, there is rapid progress in improving the phytoextraction efficiency of CWs because of the high relevance in efficient pollutant removal and reuse of wastewater. However, many studies did not further consider the utilization of biomass for later element enrichment and raw material extraction. Due to the complexity of the phytomining process chain, research requires close collaboration between ecologists, engineers, microbiologists, and chemists to cover the entire process. At present, phytomining in CWs is still a nascent technology requiring more fundamental and applied research before it can be commercially applied. Nevertheless, combining the bioremediation of wastewater with raw material production clearly offers an environmentally friendly contribution to cope with increasing wastewater discharge, increasing demands for fertilizers and high-tech metals, and closing of element cycles in the face of climate change and finite resources.

References 1. Ali S, Abbas Z, Rizwan M, Zaheer IE, Yavas I, Ünay A, Abdel-Daim MM, Jumah M, Hasanuzzaman M, Kalderis D (2020) Application of floating aquatic plants in phytoremediation of heavy metals polluted water: a review. Sustainability 12(5):1927. https://doi.org/10.3390/su12051927

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