Elsevier

Hydrometallurgy

Volume 197, November 2020, 105444
Hydrometallurgy

CEReS – co-processing of coal mine & electronic wastes: Novel resources for a sustainable future

https://doi.org/10.1016/j.hydromet.2020.105444Get rights and content

Highlights

  • Successful demonstration of a novel co-processing concept for electronic- and sulfidic coal production wastes.

  • Technical feasibility of individual units at bench and pilot scale.

  • Integration of the process flow-sheet and simulation in silico.

  • Economic and Life Cycle Analysis of the flow-sheet.

  • Recovers value and reduces environmental impact.

Abstract

Most coal mines produce waste which has the potential to generate acid mine drainage (AMD). If not properly managed, this can cause environmental damage through contamination of ground and surface waters and soils for hundreds of years. At the same time, the pace of technological development means that most electrical and electronic equipment becomes obsolete within a matter of years, resulting in the generation of vast quantities of electronic waste (e-waste). Where this cannot be recycled, it must be discarded. The CEReS concept is a co-processing approach for both waste streams to produce metals and other valuable products, and to reduce or eliminate the their environmental impact. This brings together two waste streams from opposite ends of the supply chain; turning each into a novel resource in a single, coherent ‘grave-to-cradle’ process. This industrial ecology approach is key to supporting a circular economy whilst securing the sustainable supply of critical raw materials. The project successfully elaborated a novel co-processing flow-sheet comprising: (i) the accelerated bioweathering of AMD-generating coal production wastes to generate a biolixiviant; (ii) the pyrolysis and catalytic cracking of low-grade PCBs to produce hydrocarbon fuel, a halogen brine and a Cu-rich char; (iii) the leaching of base metals from the char using the biolixiviant; (iv) the reuse of stabilised coal wastes; and (v) the full or partial (as enriched substrates) recovery of valuable metals. These process units were demonstrated individually at lab-pilot scale. The data were then used to validate the entire flow-sheet in an integrated process simulator and determine the economic balance. Finally, an LCA approach was used to demonstrate the environmental benefits of the CEReS process over the status quo.

Introduction

The European Union's urgent need for increased job creation, economic growth and resource independence is seemingly juxtaposed with its commitments to environmental rehabilitation and protection. Environmentally sustainable economic growth as a reality requires the development and implementation of innovative solutions to primary resource shortages and waste management.

CEReS was a project funded by the Research Fund for Coal and Steel (RFCS) to develop a co-processing solution for the treatment of acid-generating coal mine wastes and the recovery of (critical) raw materials from electronic wastes. Poland was chosen as a case study region, largely due to its substantial coal mining industry, challenges with subsequent wastes and relatively underdeveloped e-waste recycling sector. The project set out to demonstrate the technical feasibility of the individual unit processes at laboratory/pilot scale. These were then integrated in silico, and an economic assessment made. A life cycle assessment (LCA) approach was used to evaluate the environmental benefits of the CEReS process compared with the business-as-usual, “do-nothing” scenario.

Europe hosts large and growing volumes of wastes from past and present mining activities, of which coal production is the greatest single contributor. At the turn of the millennium, coal wastes accounted for some 2.4 Gt of the 5.9 Gt total mine wastes and tailings stockpiled within the EU (Charbonnier, 2001; Lottermoser, 2010). Since 2008, Europe has produced on average over 500 Mt. solid fossil fuels pa, of which approximately 20% was hard coal (Eurostat, 2020). Poland is Europe's largest producer of hard coal with 31 active coal mines, producing an average 74 Mt. hard coal (declining from nearly 84 Mt. pa in 2008 to just over 63 Mt. pa in 2018; Eurostat, 2020) and ~30 Mt. waste pa (Szczepañska-Plewa et al., 2010).

Coal mining waste is produced at all stages of mine development and deposit exploitation, from shaft sinking to making rock drifts in deposits, but the vast majority (almost 95%) of the waste produced is mineral processing waste (coarse-grained, fine-grained and flotation waste as well as sludge from sludge-water circulation systems). Hard coal mining is one of the biggest producers of industrial waste in Poland, accounting for 25–30% of annual industrial waste production (GUS, 2018). Currently about 90% is used in geoengineering while the rest is deposited in the environment, with well over 600 Mt. currently stockpiled (Szczepañska-Plewa et al., 2010). This has mainly been stored in heaps, which has resulted in increased costs of hard coal production and permanent adverse changes to the natural environment. The Polish Act on Mining Waste1 states that this waste should be recovered at the site of its production. Therefore, the management of hard coal mining and production wastes is a legal, environmental and economic challenge for the industry.

Coals come from reducing environments and as such, they commonly contain iron sulfides such as pyrite. During coal beneficiation these sulfide fractions, if present, report to the wastes. Where such wastes are exposed to oxygen and moisture, the microbially-mediated decomposition of the pyrite and other sulfide minerals can cause the formation of acid mine drainage (AMD). AMD is characterised by low pH and varies widely in composition, with elevated concentration of sulfates, iron, manganese, aluminium, other toxic and radioactive ions as well as excessive total dissolved solids common in AMD from sulfidic coal wastes.

AMD from underground and surface mines, waste dumps and tailing dams is one of the oldest and most consistent industrial problems facing mining regions in Europe and elsewhere, affecting at least 5000 Km of Europe's rivers (Jarvis and Younger, 2000; Lottermoser, 2010). The impact of acid mine drainage is primarily on the receiving watercourses, where its effects are complex. It is a multi-factor pollutant, and affects ecosystems through a number of direct and indirect interactions that can be both chemical and physical (e.g. Gray, 1997). Chemical effects are the result of pH, dissolved solutes and salinity, whereas physical effects are caused by the precipitation of secondary minerals and metal oxides. AMD will affect different ecosystems in different ways, and it is difficult to distinguish which component will have what effect, or which, if any, may be more important.

The longevity of AMD genesis is one of the key aspects to the problem. Unlike other industries where cessation of operations will lead to a significant reduction in pollution, the reverse is often true for the mining industry. Uncontrolled oxidative dissolution of exposed sulfide minerals will lead to continuing pollution on a time scale often greater than the entire life of the mine.

A mine waste hierarchy; the order in preference for mine waste management strategies, is shown in Fig. 1. The prevention of waste production is unachievable; therefore, reuse and recycling are the most-favoured, viable options. The most common options for the reuse of coal production wastes include: (i) backfill for open voids; (ii) landscaping material and revegetation substrate at mine sites; (iii) aggregate for civil engineering projects; and (iv) feedstock for cement and concrete.

However, the presence of sulfides causes geotechnical and environmental problems as a result of their decomposition. This unstable sulfide content renders such wastes unsuitable for use in civil engineering projects or backfill/landscaping of the mine, limiting the possibilities to reuse or recycle many coal production wastes. (Significant impacts from the reuse of sulfide-bearing coal wastes in civil engineering projects are not uncommon, e.g. the Buków flood polder, ground levelling and backfilling; Stefaniak and Twardowska, 2010; Szczepañska-Plewa et al., 2010). Therefore, the majority of (acidogenic) mine wastes produced at mine sites are still placed into storage facilities (Lottermoser, 2011).

Current BAT (best available techniques) for managing acidogenic coal wastes and residues is to prevent seepage from waste dumps and dewater tailings (EC, 2009; Verburg et al., 2009): covering them with appropriate materials (composite/vegetative/wet covers, etc.) in order to prevent moisture and oxygen ingress, thus preventing AMD formation or to store pyritic material below the water table. This solution is costly and does not address the issue of AMD generating potential. The capping materials have a limited lifetime, and constant monitoring is required. Should the material be (re-)exposed to oxygen and moisture, microbial weathering and thus AMD generation will (re)commence. With increasingly extreme weather events more likely as a result of climactic changes and increasing pressure to reuse brownfield sites, the disturbance of apparently stable mine waste impoundments seems certain.

The environmental consequences of poorly managed or improperly stabilised waste piles are well documented and greater effort needs to be made to provide reuse and recycling options that are environmentally sound. Clearly, a major step required to enable alternative options for acidogenic waste use is therefore the removal of labile sulfide content (Kazadi Mbamba et al., 2012).

Biohydrometallurgy (or biomining) is a credible biotechnology used in the mining industry. It exploits the actions of lithotrophic microorganisms to recover metal from their ores. These organisms get their energy from the oxidation of iron and/or reduced inorganic sulfur compounds (RISCs), producing sulfuric acid and ferric iron. Ferric iron (Fe3+) is the primary oxidizing agent, attacking sulfide minerals (MS), as seen in Eq. (1), and the role of the organisms is the regeneration of Fe3+ from Fe2+ and the oxidation of sulfur compounds to produce sulfuric acid (Eqs. (2), (3)).MS+2Fe3+M2++2Fe2++S02Fe2++0.5O2+2H+2Fe3++H2OS0+1.5O2+H2O2H++SO42

The result is a highly corrosive “biolixiviant” solution which attacks the mineral matrix in which target metals are entrained or form an integral part (for example copper from chalcocite or gold entrained within arsenopyrites).

Biomining was successfully used in the re-processing of sulfidic mine wastes at the Kasese Cobalt Company site in Uganda where cobalt was produced from old copper mining tailings (Morin and d'Hugues, 2007). By contributing to the stabilisation of those wastes, this biohydrometallurgical operation has also drastically decreased the AMD discharge in the environment. While this can be applied to mine wastes with an economically viable grade of valuable metals, in many cases the value of the metals within AMD-producing wastes is not sufficient to cover the costs of reprocessing them directly. This is especially the case when considering pyritic coal production wastes, which do not normally come from valuable metal-bearing assemblages.

Studies since the 1960's have shown bioleaching can effectively remove inorganic and organic sulfur in coal prior to combustion (Cardona and Márquez, 2009; He et al., 2012; Hoffmann et al., 1981; Olson and Kelly, 1991; Schippers et al., 1999). Up to now this is not done commercially, despite the design and operation of pilot scale systems to remove both pyrite and organic sulfur (Cara et al., 2005; Milan et al., 2017; Ors et al., 1991; Rossi, 2014). Again, such approaches were designed for the removal of sulfur from coal, not coal production wastes. Such an approach uniquely for desulfurisation of waste is difficult to justify financially over preventative methods: once the sulfide is leached, there is a need to spend extra money on neutralisation of acid and management of iron (Klein, 1998). Conversely, in the CEReS process, the acid and ferric iron is a desirable product. CEReS uses biodesulfurisation to remove pyrite (and other sulfides/metals) from coal production wastes, and makes use of acid and ferric iron generated (as a biolixiviant), rather than requiring immediate neutralisation.

Supplying and securing mineral resources with minimum environmental footprint is a serious challenge, especially for the European Union which consumes 25 to 30% of the world's metal but accounts for around 5% of the world's mining output (Östensson, 2006). European dependency on metal import is growing every year despite efforts in the development of recycling technologies and material science. This has been further emphasised in the “Criticality Report” compiled for the European Commission in 2010 and revised in 2014,2 in which 20 mineral raw materials have been explicitly named as highly critical for the industrial development and economic security of the European Union. These tensions highlight the need to associate the identification of new potential resources that could be used for the recovery of rare and valuable materials with the development of recycling processes in order to close the gap in raw materials. Improving their mode of production and developing a circular component of the economy is imperative.

Among the different types of secondary post-consumption wastes, electronic-wastes (e-wastes) represent the fastest growing and most problematic waste stream in the world. In the EU, 10–12 million tonnes are produced pa (Balde et al., 2015; Goodship et al., 2019; Huisman et al., 2008). The Commission seeks to address this through DIRECTIVE 2012/19/EU on waste electrical and electronic equipment (WEEE).

E-wastes contain a wide range of different metals and other compounds, many of which are highly toxic to the environment and human health. For example, a UK study has found that the presence of e-wastes in municipal waste is the major single source of toxic elements in the potentially biodegradable fraction (Papadimitriou et al., 2008). In Japan, more than half of the copper from WEEE ends up in landfill or is lost (Oguchi et al., 2012). EEE is made up of many individual components. These include printed circuit boards (PCBs), cathode ray tubes (CRTs; from older PC monitors and TVs), batteries, internal and external wiring and the equipment casing. Each has its own unique composition and associated environmental hazards. Unsurprisingly, the Basel Convention has identified e-waste as toxic. As a direct result, e-wastes can only be exported where it can be shown that the wastes will be managed in an environmentally sound manner in the country of import. Despite this, and despite improving efforts to collect and recycle e-waste within the EU, significant amounts find their way to non-developed nations (Breivik et al., 2014), and the severe environmental implications of their improper storage and processing are well documented (e.g. Nnorom and Osibanjo, 2008).

The majority of the value in e-waste is in the PCBs. On average 90% of the intrinsic economic value of PCBs is in the precious metals that they contain (Cui and Zhang, 2008; Luda, 2011). These metals make up the majority of the economic value of WEEE, and are in concentrations at least 10 times higher than their typical mineral ores (Huang et al., 2009; Tuncuk et al., 2012). Many of these are of significant strategic importance and are reaching their extraction peaks. The average price of rare earth ores has almost doubled since 2007 (Kingsnorth, 2011). China supplies 95% of the world's rare earths, and a recent 40% reduction in Chinese export quotas demonstrated the strategic importance of a reliable and secure source of rare earths. Nevertheless, China produces 90% of the world's electronic goods (Widmer et al., 2005). This is an ‘open loop’ in that it can be viewed as a net export of valuable raw materials.

The US Environmental Protection Agency has identified e-wastes as a good way of generating a source of valuable metals, in what it terms “Urban Mining” (EPA, 2011). Recovering metals from e-waste is potentially more energy efficient than mining raw material. For example, recycling metals directly can lead to huge energy savings: 95% for aluminium, 85% for copper, 65% for lead and 60% for zinc (Cui and Forssberg, 2003; Nnorom and Osibanjo, 2008). Therefore, it is desirable to reprocess e-wastes not just on environmental grounds, but also economic.

The best options for dealing with WEEE and e-waste are direct reuse and resale or remanufacturing via refurbishment (Cui and Zhang, 2008). However, given the short lifespan of many electronic devices, particularly computers and mobile phones, this is not always achievable as a major destination. Landfill is the least desirable option, and most countries are increasingly limiting this due to land contamination issues and associated cost of toxic compound handling and disposal. Therefore, recycling is the most pragmatic approach for value recovery and environmental protection (Lundstedt, 2011).

PCBs are a mixture of polymers, ceramics and metals tightly bonded together. The board itself is made up of fibreglass-reinforced thermosetting matrix, which may contain up to 15% bromine (used in bromo-phenol flame-retardants; Luda, 2011). They are among the most complex sub-components of e-waste and most difficult to reprocess. The metal content of the PCBs is highly variable, depending on the type and make of the equipment. The metal content determines process selection and economics. For example, PCBs from mobile phones tend to contain greater concentrations of metals than those from PCs, while boards from TVs contain less than 100 ppm gold are usually considered low-grade (Kasper et al., 2011).

Pyrometallurgy is the traditional choice for metal refining from processed (usually upgraded) e-waste, resulting in the production of precious metal-bearing copper bullion (Tuncuk et al., 2012). This means that selective recovery of individual metals is effectively impossible by this route, and further recovery processes are needed. It can be done within existing smelters treating mineral concentrates, where e-waste may be combined (10–15%) with a copper concentrate (Cui and Zhang, 2008). However, it is energy intensive and requires a relatively high grade feed material, and the ceramics contribute to final slag volume. Proper emissions control is also necessary due to the production of dioxins, furans and other polybrominated organic compounds and polyaromatic hydrocarbons during PCB incineration (Huang et al., 2009).

Well-regulated smelters have processed 15–20 thousand tonnes of e-waste with 95% metal recovery and minimal generation of dioxins, though the economic viability is often questionable; reports indicate that the cost of smelting PCBs approximately equates to 50% of the revenue generated (Lehner, 1998; Mark and Lehner, 2000; PHA, 2006).

At present within the EU there are only three dedicated smelters (the Umicore plant in Hoboken, Belgium, the Rönnskär Smelter in Sweden and the Kayser Recycling Smelter in Aurubis, Germany) that can handle e-wastes (Khaliq et al., 2014). A requirement however for the PCB's to enter as input steams in these installations is that precious metals concentration should exceed a given cut-off grade. Given the disparate nature of e-waste production, there are logistical limitations to collection and transport. Moreover, in such a process critical metals such as REE are lost to the final slag, and there is limited or no recovery of other products such as halogens or fuels. Thus, there is still significant waste in the system and loss of valuable resources as a result.

Compared to pyrometallurgy, hydrometallurgical processes offer relatively low capital cost and are particularly suitable for small-scale installations (Tuncuk et al., 2012). An added advantage is their flexibility, offering a possibility for selective extraction of base and precious metals of interest in e-waste and PCBs. Since major metals exist in their elemental or alloy form in PCBs, their hydrometallurgical extraction has been tested using various oxidants (lixiviants; hydrogen peroxide, oxygen, ferric iron, etc.) under acidic (HCl, H2SO4, HNO3 etc.) or ammoniacal and chloride leaching environments. While cyanide is the most economically feasible of common leaching methods, it is also the highest in terms of toxicity. Research and development in hydrometallurgical applications remains rather scattered and mostly at lab-scale. Therefore, the generation of operational and cost data via pilot scale tests is essential. Moreover, the costs of the lixiviants contribute to overall operating expenditure of hydrometallurgical options which may include treatment of heavily polluted by-products in special waste disposal facilities, and further limits the grade of PCB that can be treated economically. Consequently, there are only a limited number of hydrometallurgy operations in Europe for e-wastes and these are reserved for high grade materials; there is currently no suitable option for low-grade PCBs and a lack of PCB processing capacity more generally.

Studies into the bioleaching of e-waste have mainly involved the treatment of printed circuit boards (PCBs). The use of organic acids produced by various fungi or biogenic cyanide has been examined, particularly for the recovery of gold and other noble metals (Brandl et al., 2001; Brandl et al., 2008; Chi et al., 2011; Faramarzi et al., 2004). However, such approaches require the selective cultivation of specific microorganisms in circum-neutral media rich in organic substrates. This requires aseptic growth conditions and is unlikely to be practical (or economic) when treating large volumes of non-sterile e-wastes. Therefore, the use of ferric iron and/or proton lixiviants produced by extreme acidophiles is preferable. There is no need for sterile conditions, and media are simple, comprising key nutrients such as sources of nitrogen, potassium and phosphorus.

In one-step tests (where the e-waste and microorganisms are introduced in a single vessel), reported copper recovery efficiencies vary widely from less than 4% to 100% (Rivero-Hudec et al., 2009; Wang et al., 2009; Xiang et al., 2010; Zhu et al., 2011) with leaching times typically greater than 10 days and pulp densities around 1% or less. Performance decreases significantly with increasing pulp density (Brandl et al., 2001; Liang et al., 2010; Wang et al., 2009; Xiang et al., 2010; Zhu et al., 2011). Generally, the lower the initial pH, the better the performance, though optimal pH ranges are apparent below which the organisms suffer. The same is true of ferrous iron concentration; optimal performance is reported between 7 and 9 g/L. Above this, factors such as proton consumption during iron oxidation and ferric iron precipitation affect final metal recoveries (Choi et al., 2004; Xiang et al., 2010; Zhu et al., 2011). The waste tends to be acid-consuming, probably as a result of several factors including the dissolution of acid-soluble metals. Performance may be better where constant pH is maintained (Yang et al., 2009), but this is not universal (Vestola et al., 2010). Several studies have demonstrated improved leaching performance in media containing elemental sulfur as well as ferrous iron (Ilyas et al., 2007; Liang et al., 2010; Wang et al., 2009).

The toxicity of the e-waste on the microorganisms has been shown to be the major problem preventing efficient leaching. Up to 90% copper recovery from a 10% pulp density within 18 days has been reported using a culture adapted to high metal concentrations (Ilyas et al., 2007). However, the crushed PCBs were pre-washed in saturated NaCl and contained relatively low levels of copper (8.5% w/w). Nevertheless, adaptation to elevated metal concentrations rather than prewashing the e-waste seems to be more important in improving leaching rates.

Staggering the production of the lixiviant and the addition of the e-waste in a two-step process has been shown to greatly increase final copper recoveries and leaching rates (Liang et al., 2010; Xiang et al., 2010; Yang et al., 2009; Zhu et al., 2011) but has not been tested at pulp densities greater than 2%. Furthermore, if the addition of the e-waste at the second step results in a loss of culture viability, then fresh media and inocula will be required for each subsequent run and it may not be possible to subculture from one run to the next.

In mineral bioleaching the source of iron and sulfur which is oxidised by the microbial community to produce the oxidising lixiviant solution is inherent in the form of pyrite or other sulfide minerals (it is an autocatalytic process). In the current e-waste bioleaching practices, this must be provided in addition to the nutritive medium. This is usually in the form of ferrous sulfate with acid provided directly via pH control with sulfuric acid or through the addition of elemental sulfur. Furthermore, PCBs are highly acid-consuming and require a high degree of pH modification to maintain an acidic environment necessary for the microbial action and maintained metal solubility. These chemicals adding increases the operating cost of the process.

While such lab-scale studies provide evidence that biohydrometallurgical reprocessing of e-wastes is technically possible, the wider economics of such a process are unlikely to be favourable with the current state-of-the-art. CEReS will decouple lixiviant production from PCB leaching in a two-step process to avoid issues of toxicity.

The lixiviant will be derived from a net acid generating source (sulfidic mine wastes) as opposed to ferrous sulfate (the oxidation of which is acid consuming; Eq. 2) to overcome acid-consumption issues. Such an approach has been successfully demonstrated a varying scales of application to waste PCBs (Bryan et al., 2015; Guezennec et al., 2015) and other forms of post-consumer, metallic waste (Lewis et al., 2011). The use of biohydrometallurgy to generate the lixiviant from mine wastes is preferable to chemical oxidation options as it can be done at standard temperature and pressure, does not require expensive or environmentally damaging reagents and traditionally offers lower operating expenditures as a result. In this context, CEReS concept can be seen to be simultaneously upgrading the metal content of mine wastes while providing a source of lixiviant to the e-waste making reprocessing economically viable both in terms of metal recovery but also operating expenditure.

Section snippets

CEReS

The co-processing approach proposed by CEReS employs AMD-generating coal production wastes as a cheap source of leaching solution (lixiviant) to recover metals from e-wastes. The novel flow-sheet will (i) remove the AMD-generating potential of coal wastes, ensuring their long term environmental stability while expanding avenues for their safe reuse; and (ii) enable selective recovery of base metals from waste PCBs, while concentrating precious and critical as well as rare earths into enriched

Coal production waste

In Poland, especially in the Upper Silesian Coal Basin, there are about 160 potential mine waste dump sites. Some of those possess potential risk for the generation of AMD (acid mine drainage) due to the large presence of pyrite. Tauron Wydobycie S.A. own the Brzeszcze, Janina & Sobieski coal mines, and all generate sulfidic wastes during coal production.

Four waste types where considered from two mines, Janina and Sobieski: sludge (−0.1 mm), spiral tails (0.1–2 mm), jig tails (2–20 mm) and

Pyrolytic cracking

The treatment and recovery sector for scrap metal, end-of-life vehicles (ELVs), and WEEE generates a large amount of waste shredder residues (SR). These residues offer an ideal opportunity to recover materials that have become highly significant both economically and environmentally. In addition to this the European Union has set ambitious recovery targets for these materials. For example, 95% of an ELV must be recycled by 2015.

COMET Traitements SA developed and implemented new treatment and

Bioleaching and lixiviant production

Complete biooxidation of sulfidic minerals involves the action of a consortium of both iron- and sulfur-oxidising extremophile microorganisms adapted to an inorganic and acidic environment. Many biomining microorganisms occurring naturally on mineral ores are known (Hallberg and Barrie Johnson, 2001; Rawlings and Johnson, 2007). Autotrophic species of the iron- and sulfur-oxidizing Acidithiobacillus genus and the iron-oxidizing Leptospirillum genus are significant contributors to commercial

Process integration, economic and environmental assessment

Process integration includes efficient use of raw materials, energy efficiency and emissions reduction (Friedler, 2010; Smith, 2000). The implementation of unit operation models within a steady-state simulator enables to facilitate both process integration and flow-sheet optimization (Brochot et al., 2002). The simulator is used to analyse, model and optimize the possible interactions between the equipment in the flow-sheet, in order to maximize resource valorisation, and minimize utilities

Conclusions

The project successfully demonstrated the key aspects of the CEReS flowsheet at laboratory and pilot scale. The process was able to convert scrap PCBs into combustible fuels, a halogen brine and a metal-rich char. AMD-generating coal production waste could be stabilised through accelerated weathering and used in construction products. The lixiviant produced was suitable for leaching the copper from the char and this refined into copper cathodes. In silico integration and simulation demonstrated

Declaration of Competing Interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

Acknowledgements

This project has received funding from the Research Fund for Coal and Steel under grant agreement No 709868. Dedicated to the memory of Dr. Björn Debecker.

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