Rare earth elements: Chemistry, fate and environmental impact
The use of rare earth elements (REEs) in industrial, medical and consumer products is increasing rapidly. REEs are important components of the new technologies such as wind turbines and electric vehicles, that are vital to the transition to a low carbon economy. Because of their chemistry and speciation, REEs are rarely present in a form that can cause toxicity, but REE concentrations are noticeably elevated in surface waters and sediments globally. This article explores the current and future uses of REEs and discusses the possible environmental impacts of REE emissions.
REEs, also known as the lanthanide elements, are a group of 15 elements with atomic numbers 57 to 71 (Figure 1). Despite their name, they are not particularly rare. For example, Ce (the most abundant REE) is more abundant than Cu or Pb and several REEs are more common than Sn, Mn, Ag or Hg. The REEs often occur together because they are chemically similar (all have an oxidation state of 3+, have similar ionic radii and all inhabit the same row in the periodic table), which enables them to substitute for each other in crystals. There is, however, a small but consistent decrease in ionic radius with increasing atomic number which occurs alongside a decrease in abundance. REEs with atomic numbers 57 to 63 are considered light REEs (LREEs) while those with atomic numbers 64 to 71 are considered heavy REEs (HREEs). The REEs with an even atomic number are more abundant than REEs with an odd atomic, according to the Oddo-Harkins rule that describes proton pairing during star formation.
REEs, also known as the lanthanide elements, are a group of 15 elements with atomic numbers 57 to 71 (Figure 1). Despite their name, they are not particularly rare. For example, Ce (the most abundant REE) is more abundant than Cu or Pb and several REEs are more common than Sn, Mn, Ag or Hg. The REEs often occur together because they are chemically similar (all have an oxidation state of 3+, have similar ionic radii and all inhabit the same row in the periodic table), which enables them to substitute for each other in crystals. There is, however, a small but consistent decrease in ionic radius with increasing atomic number which occurs alongside a decrease in abundance. REEs with atomic numbers 57 to 63 are considered light REEs (LREEs) while those with atomic numbers 64 to 71 are considered heavy REEs (HREEs). The REEs with an even atomic number are more abundant than REEs with an odd atomic, according to the Oddo-Harkins rule that describes proton pairing during star formation.
Figure 1. The lanthanides are referred to as rare earth elements.
Pure forms of REEs tend not to exist in the environment because they are very reactive and form part of the crystalline lattices of minerals. Even after weathering, they are usually incorporated into secondary minerals and remain immobilised. Many REEs were originally isolated in the 18th and 19th century, but only efficiently in the 20th century. Apart from Ce, La, and Nd, REEs were not commercially available until 1940s. The market for REEs has arisen in the last fifty years, and demand is expected to increase exponentially in the next twenty-five years. Up until 1985, most REEs originated from minerals mined at Mountain Pass in California. However, China emerged as a major producer in the 1990s with extraction of REEs from ores located at the Bayan Obo Iron deposit, Inner Mongolia. As a result, China now supplies more than 90% of the global output of REEs.
The properties of REEs are exploited to produce a wide range of consumer products. REEs have been used in China in fertilisers and as feed additives for livestock. There is some evidence linking this application to increased crop yields, and to gains in animal body weight and enhanced milk/egg production (2). REEs are incorporated into catalysts for petroleum cracking in oil refineries. Cerium carbonate and cerium oxide are used in diesel fuel additives and in catalytic converters, resulting in the release of CeO2 nanoparticles in vehicle exhausts. Phosphors for fluorescent lighting contain several REEs (Tb, Ce, and La for green and blue emissions, and Eu for red emission). However, the environmental impact of REEs in these phosphors is often overlooked because they also contain Hg, used to produce natural white light. Some MRI contrast agents, e.g. gadopentetic acid, contain Gd as a complex ion. Oral or intravenous administration of the contrast agent results in urinary excretion of the absorbed Gd, where it enters wastewater treatment plants (see below). Several applications of REEs are a direct result of legislation to improve air quality (e.g. Ce in catalytic converters) or the transition to a low carbon economy. For example, Nd, Dy and Sm are used in the manufacture of magnetic alloys for hybrid engines and in permanent magnets for wind turbines (as well as a whole host of consumer electric goods such as headphones, computer hard drives, and electric motors). Nickel-metal hydride rechargeable batteries for electric and hybrid vehicles may contain La Ce, Nd, or Pr as the anode.
Environmental fate
There are very few studies of the environmental fate of REEs. Because they are considered less toxic than many other elements (e.g. As, Cd, Cr, Hg, Pb), the environmental impact of REEs is often overshadowed by these other elements. For example, when REE ore is mined, the major hazards are associated with other toxic metals present in the REE-bearing ore and in waste streams. Waste materials from REE processing are naturally radioactive due to the presence of, for example, thorium and uranium radioisotopes. Clean-up costs for spillages of radioactive wastewater from the processing of REE ore at the Mountain Pass mine in California contributed to its decline and closure in the 1990s. Despite these hazards, REE mining is generally considered to have a low environmental impact. REEs tend to be present in carbonate minerals rather than in sulphides, meaning that acid mine drainage is less common at REE mines than other metalliferous mines.
The impact on the environment from the use of REEs was first detected in the early 1980s. Analysis of sediment cores collected from the San Pedro Shelf, just south of Los Angeles in California, revealed 890-6900 times more light REEs (La, Ce, Nd and Sm) than present in crustal material (3). The REEs were traced to a wastewater outfall discharging into the basin that served 3.5 million people, tens of thousands of businesses and industries, and, importantly, seventeen oil refineries. Most wastewater is discharged into rivers where REE contaminants may be deposited on land during flood events or carried out to sea. About 1.5 tonnes of anthropogenic La is discharged into the North Sea from the river Rhine every year. The source of this contamination is an industrial plant near the city of Worms in Rhineland-Palatinate, which uses fluid catalytic cracking agents. La concentrations (49 ppm) in the plant effluent exceed levels known to cause ecotoxicological effects (4).
Gadolinium from MRI contrast agents is the most ubiquitous REE discharged into rivers via wastewater treatment plants. The relatively stable Gd complex is soluble in aquatic ecosystems. As a result, influent and effluent of wastewater treatment plants is often enriched in Gd, but an elevated Gd concentration is not found in sewage sludge. Nozaki et al. (5) found evidence for elevated Gd concentrations in three rivers, which discharged into Tokyo Bay. They concluded that not only population size, but also the degree of modernisation of medical treatments affects REE enrichment in the environment. It has been suggested that, due to its solubility and ubiquity, Gd could be used as a tracer of other emerging contaminants discharged from wastewater treatment plants, which are more difficult to analyse, such as pharmaceuticals or personal care products (6).
In an ecological risk assessment of REEs on aquatic organisms in rivers, ecotoxicity tests were performed for several aquatic invertebrates. From the dose-response curves, Predicted No Effect Concentrations (PNECs) were computed for Ce, Gd and Lu. These PNECs were used to produce risk quotients by comparison with a global dataset of REE concentrations in rivers (7). Risk quotients were above 1 for some locations (e.g. effluent from a wastewater treatment plant downstream of a hospital or residential area), indicating that REEs were present at concentrations above their PNECs.
Toxicity
Because REEs are chemically similar, one might expect their ecotoxicological effects also to be similar. However, a current lack of knowledge on the mechanisms of toxicity of REEs precludes this assumption. Most ecotoxicology studies have focused on freshwater organisms, and there are only a few published salt water or terrestrial studies (8). Toxicity seems to decrease with increasing atomic number due to higher stability constants of the heavier REEs, indicating that the speciation of the element is of greater importance than its chemical identity. The influence of speciation (and thus bioavailability) on toxicity is paramount as the REEs are usually present in the 3+ oxidation state so they can form soluble complexes with nitrates, chlorides, and sulphates. In highly complexing media, REEs form insoluble species with carbonates, phosphates and hydroxides, which means their toxicity could be underestimated in these solutions. REE complexes become less soluble with decreasing temperature and increasing pH or redox potential. Thus, chemical speciation modelling is required to estimate the free ion concentration in solutions and to predict toxicity.
Most toxicological data relate to Ce and La, with a little information on Gd and Nd, but virtually nothing for the other REEs. REEs have a similar ionic radius to calcium and so can replace Ca in cell functions or block Ca channels. However, the most widely reported mechanism of toxicity is a redox imbalance in cells that leads to oxidative stress. Several studies have used reactive oxygen species (ROS) or antioxidant activities (superoxide dismutase, catalase, and glutathione peroxidase) as a toxic endpoint (9). There are reports that REEs have a beneficial effect at low concentrations and adverse effects at higher concentrations (hormesis). For example, the stimulation of antioxidant enzymes in plants with low La levels could be interpreted as a toxic response, but the effect has been found to be positive at low concentrations, helping the plant to defend against ROS caused by other stresses (8). This mechanism may be responsible for the increased crop yield and livestock growth rates observed in response to REE additions to fertilisers and feed additives in China (2).
Relatively little attention has been paid to the effects of REE exposure on human health. Studies have primarily focused on monitoring populations who reside next to REE mines in China. An investigation of residents living close to the Bayan Obo Iron deposit in Inner Mongolia found REE levels in hair samples above those from a control area (10). Concentrations were higher in male hair than female, and higher in miners than non-miners, indicating occupational exposure. Accumulation of the light REEs also increased with age, which suggests a slow excretion rate and the potential for bioaccumulation. A similar study conducted in a large-scale mining area located in Hetian Town of Changting County, Fujian Province, Southeast China showed that REEs from mining were present in the soil, water and vegetables produced in the vicinity of the mine, and these were associated with elevated concentrations in the hair and blood of local farmers (11). A risk assessment indicated that the ingestion of contaminated vegetables did not exceed values known to be hazardous for human health, but that long term exposure should be considered.
Recycling and recovery
China produces more than 90% of the global REE output and imposes a quota on the export of REEs. This restriction is encouraging the recovery and reuse of REEs. Less than 1% of REEs in ‘end-of-life’ products were recycled in 2011, but there is considerable scope for recycling the REEs used in permanent magnets, lamp phosphors, rechargeable batteries and catalysts (12). Because REEs usually occur together in a fixed ratio, some may be stockpiled while there is a high demand for others, and thus their demand is controlled by the element in most scarce supply. For example, because Nd is much less common than La or Ce, mining ore for Nd produces an over-supply of La and Ce. Future demand for electric vehicles and wind turbines will rely heavily on Dy and Nd, which are used in permanent magnets. Because these technologies are important components of our strategy to transition to a low carbon economy, the supply of REEs has been highlighted as critical to achieving carbon neutrality. It has been predicted that the supply of Dy and Nd will need to increase by 700% and 2600%, respectively, between 2010 and 2035 if atmospheric CO2 is to be stabilised at 450 ppm using current technologies (13). The REEs of greatest concern are therefore those that are of high importance to clean energy and also have a high supply risk (Nd, Dy, Eu and Tb). One of the best opportunities for recycling is permanent magnets which are based upon neodymium-iron-boron (NdFeB) alloys and also contain Dy, Pr, Gd and Tb, representing some of the most critical REEs in terms of future supply risk. There are also opportunities for recycling REEs from nickel metal-hydride batteries and lamp phosphors.
Several studies have sought to identify the opportunities for recovering REEs from mixed waste streams. Commercially recoverable quantities of REEs, primarily Ce, have been detected in landfills by Gutiérrez-Gutiérrez et al. (14), but recovery would only be economically feasible if other valuable materials were also recovered simultaneously. Bottom and fly ash from municipal solid waste incinerators may prove to be a promising source, as strong enrichment of Eu, La, Gd, and Tb correlates with phosphorus pentoxide (P2O5) and indicates that the source of REE anomalies in the ash are likely from REE phosphors present in fluorescent materials (15). REEs were found to be present at several orders of magnitude higher in acid mine drainage water than natural water (16) presenting an opportunity to provide a modest but continuous supply of potentially valuable resources as a by-product of otherwise costly remediation activities. Sewage sludge only seems to be a promising source of Gd though (presumably due to its use as a MRI contrasting agent), since other REEs show enrichment factors near unity indicating geogenic origin (17).
The maturation of technologies for recovering REEs from waste streams and recycling them in ‘end-of-life’ products may result in lower concentrations released into environmental media. However, this must be contrasted against an anticipated exponential growth in their use in consumable products. It is unlikely that the environmental impact of REEs, which seems to be lower than many other metallic elements, will drive increases in recycling and recovery. Therefore, the adoption of technologies to exploit “second-hand” REE resources will likely be dependent on the price of REEs on the global market.
References
1. ECG Bulletin, July 2013, pp. 3-5 and pp. 10-20, see http://www.rsc.org/images/ECG-Bulletin-July-2013_tcm18-233368.pdf.
2. G. Pagano et al., Rare earth elements in human and animal health: State of art and research priorities, Environmental Research 142, 215-220 (2015).
3. I. Olmez, E. R. Sholkovitz, D. Hermann, R. P. Eganhouse, Rare earth elements in sediments off southern California: a new anthropogenic indicator. Environmental Science & Technology 25, 310-316 (1991).
4. S. Kulaksız, M. Bau, Rare earth elements in the Rhine River, Germany: first case of anthropogenic lanthanum as a dissolved microcontaminant in the hydrosphere, Environment International 37, 973-979 (2011).
5. Y. Nozaki, D. Lerche, D. S. Alibo, M. Tsutsumi, Dissolved indium and rare earth elements in three Japanese rivers and Tokyo Bay: evidence for anthropogenic Gd and In, Geochimica et Cosmochimica Acta 64, 3975-3982 (2000).
6. P. L. Verplanck et al., Evaluating the behavior of gadolinium and other rare earth elements through large metropolitan sewage treatment plants. Environmental Science & Technology 44, 3876-3882 (2010).
7. V. González et al., Lanthanide ecotoxicity: First attempt to measure environmental risk for aquatic organisms, Environmental Pollution 199, 139-147 (2015).
8. V. Gonzalez, D. A. Vignati, C. Leyval, L. Giamberini, Environmental fate and ecotoxicity of lanthanides: are they a uniform group beyond chemistry?, Environment International 71, 148-157 (2014).
9. G. Pagano, M. Guida, F. Tommasi, R. Oral, Health effects and toxicity mechanisms of rare earth elements—Knowledge gaps and research prospects, Ecotoxicology and Environmental Safety 115, 40-48 (2015).
10. B. Wei, Y. Li, H. Li, J. Yu, B. Ye, T. Liang, Rare earth elements in human hair from a mining area of China. Ecotoxicology and Environmental Safety 96, 118-123 (2013).
11. X. Li, Z. Chen, Z. Chen, Y. Zhang, A human health risk assessment of rare earth elements in soil and vegetables from a mining area in Fujian Province, Southeast China, Chemosphere 93, 1240-1246 (2013).
12. K. Binnemans et al., Recycling of rare earths: a critical review, Journal of Cleaner Production 51, 1-22 (2013).
13. E. Alonso et al., Evaluating rare earth element availability: A case with revolutionary demand from clean technologies, Environmental Science & Technology 46, 3406-3414 (2012).
14. S. C. Gutiérrez-Gutiérrez, F. Coulon, Y. Jiang, S. Wagland, Rare earth elements and critical metal content of extracted landfilled material and potential recovery opportunities, Waste Management 42, 128-136 (2015).
15. V. Funari, S. N. H. Bokhari, L. Vigliotti, T. Meisel, R. Braga, The rare earth elements in municipal solid waste incinerators ash and promising tools for their prospecting, Journal of Hazardous Materials 301, 471-479 (2016).
16. C. Ayora et al., Recovery of Rare Earth Elements and Yttrium from Passive-Remediation Systems of Acid Mine Drainage, Environmental Science & Technology 50, 8255-8262 (2016).
17. P. Westerhoff et al., Characterization, recovery opportunities, and valuation of metals in municipal sludges from US wastewater treatment plants nationwide. Environmental Science & Technology 49, 9479-9488 (2015).
The properties of REEs are exploited to produce a wide range of consumer products. REEs have been used in China in fertilisers and as feed additives for livestock. There is some evidence linking this application to increased crop yields, and to gains in animal body weight and enhanced milk/egg production (2). REEs are incorporated into catalysts for petroleum cracking in oil refineries. Cerium carbonate and cerium oxide are used in diesel fuel additives and in catalytic converters, resulting in the release of CeO2 nanoparticles in vehicle exhausts. Phosphors for fluorescent lighting contain several REEs (Tb, Ce, and La for green and blue emissions, and Eu for red emission). However, the environmental impact of REEs in these phosphors is often overlooked because they also contain Hg, used to produce natural white light. Some MRI contrast agents, e.g. gadopentetic acid, contain Gd as a complex ion. Oral or intravenous administration of the contrast agent results in urinary excretion of the absorbed Gd, where it enters wastewater treatment plants (see below). Several applications of REEs are a direct result of legislation to improve air quality (e.g. Ce in catalytic converters) or the transition to a low carbon economy. For example, Nd, Dy and Sm are used in the manufacture of magnetic alloys for hybrid engines and in permanent magnets for wind turbines (as well as a whole host of consumer electric goods such as headphones, computer hard drives, and electric motors). Nickel-metal hydride rechargeable batteries for electric and hybrid vehicles may contain La Ce, Nd, or Pr as the anode.
Environmental fate
There are very few studies of the environmental fate of REEs. Because they are considered less toxic than many other elements (e.g. As, Cd, Cr, Hg, Pb), the environmental impact of REEs is often overshadowed by these other elements. For example, when REE ore is mined, the major hazards are associated with other toxic metals present in the REE-bearing ore and in waste streams. Waste materials from REE processing are naturally radioactive due to the presence of, for example, thorium and uranium radioisotopes. Clean-up costs for spillages of radioactive wastewater from the processing of REE ore at the Mountain Pass mine in California contributed to its decline and closure in the 1990s. Despite these hazards, REE mining is generally considered to have a low environmental impact. REEs tend to be present in carbonate minerals rather than in sulphides, meaning that acid mine drainage is less common at REE mines than other metalliferous mines.
The impact on the environment from the use of REEs was first detected in the early 1980s. Analysis of sediment cores collected from the San Pedro Shelf, just south of Los Angeles in California, revealed 890-6900 times more light REEs (La, Ce, Nd and Sm) than present in crustal material (3). The REEs were traced to a wastewater outfall discharging into the basin that served 3.5 million people, tens of thousands of businesses and industries, and, importantly, seventeen oil refineries. Most wastewater is discharged into rivers where REE contaminants may be deposited on land during flood events or carried out to sea. About 1.5 tonnes of anthropogenic La is discharged into the North Sea from the river Rhine every year. The source of this contamination is an industrial plant near the city of Worms in Rhineland-Palatinate, which uses fluid catalytic cracking agents. La concentrations (49 ppm) in the plant effluent exceed levels known to cause ecotoxicological effects (4).
Gadolinium from MRI contrast agents is the most ubiquitous REE discharged into rivers via wastewater treatment plants. The relatively stable Gd complex is soluble in aquatic ecosystems. As a result, influent and effluent of wastewater treatment plants is often enriched in Gd, but an elevated Gd concentration is not found in sewage sludge. Nozaki et al. (5) found evidence for elevated Gd concentrations in three rivers, which discharged into Tokyo Bay. They concluded that not only population size, but also the degree of modernisation of medical treatments affects REE enrichment in the environment. It has been suggested that, due to its solubility and ubiquity, Gd could be used as a tracer of other emerging contaminants discharged from wastewater treatment plants, which are more difficult to analyse, such as pharmaceuticals or personal care products (6).
In an ecological risk assessment of REEs on aquatic organisms in rivers, ecotoxicity tests were performed for several aquatic invertebrates. From the dose-response curves, Predicted No Effect Concentrations (PNECs) were computed for Ce, Gd and Lu. These PNECs were used to produce risk quotients by comparison with a global dataset of REE concentrations in rivers (7). Risk quotients were above 1 for some locations (e.g. effluent from a wastewater treatment plant downstream of a hospital or residential area), indicating that REEs were present at concentrations above their PNECs.
Toxicity
Because REEs are chemically similar, one might expect their ecotoxicological effects also to be similar. However, a current lack of knowledge on the mechanisms of toxicity of REEs precludes this assumption. Most ecotoxicology studies have focused on freshwater organisms, and there are only a few published salt water or terrestrial studies (8). Toxicity seems to decrease with increasing atomic number due to higher stability constants of the heavier REEs, indicating that the speciation of the element is of greater importance than its chemical identity. The influence of speciation (and thus bioavailability) on toxicity is paramount as the REEs are usually present in the 3+ oxidation state so they can form soluble complexes with nitrates, chlorides, and sulphates. In highly complexing media, REEs form insoluble species with carbonates, phosphates and hydroxides, which means their toxicity could be underestimated in these solutions. REE complexes become less soluble with decreasing temperature and increasing pH or redox potential. Thus, chemical speciation modelling is required to estimate the free ion concentration in solutions and to predict toxicity.
Most toxicological data relate to Ce and La, with a little information on Gd and Nd, but virtually nothing for the other REEs. REEs have a similar ionic radius to calcium and so can replace Ca in cell functions or block Ca channels. However, the most widely reported mechanism of toxicity is a redox imbalance in cells that leads to oxidative stress. Several studies have used reactive oxygen species (ROS) or antioxidant activities (superoxide dismutase, catalase, and glutathione peroxidase) as a toxic endpoint (9). There are reports that REEs have a beneficial effect at low concentrations and adverse effects at higher concentrations (hormesis). For example, the stimulation of antioxidant enzymes in plants with low La levels could be interpreted as a toxic response, but the effect has been found to be positive at low concentrations, helping the plant to defend against ROS caused by other stresses (8). This mechanism may be responsible for the increased crop yield and livestock growth rates observed in response to REE additions to fertilisers and feed additives in China (2).
Relatively little attention has been paid to the effects of REE exposure on human health. Studies have primarily focused on monitoring populations who reside next to REE mines in China. An investigation of residents living close to the Bayan Obo Iron deposit in Inner Mongolia found REE levels in hair samples above those from a control area (10). Concentrations were higher in male hair than female, and higher in miners than non-miners, indicating occupational exposure. Accumulation of the light REEs also increased with age, which suggests a slow excretion rate and the potential for bioaccumulation. A similar study conducted in a large-scale mining area located in Hetian Town of Changting County, Fujian Province, Southeast China showed that REEs from mining were present in the soil, water and vegetables produced in the vicinity of the mine, and these were associated with elevated concentrations in the hair and blood of local farmers (11). A risk assessment indicated that the ingestion of contaminated vegetables did not exceed values known to be hazardous for human health, but that long term exposure should be considered.
Recycling and recovery
China produces more than 90% of the global REE output and imposes a quota on the export of REEs. This restriction is encouraging the recovery and reuse of REEs. Less than 1% of REEs in ‘end-of-life’ products were recycled in 2011, but there is considerable scope for recycling the REEs used in permanent magnets, lamp phosphors, rechargeable batteries and catalysts (12). Because REEs usually occur together in a fixed ratio, some may be stockpiled while there is a high demand for others, and thus their demand is controlled by the element in most scarce supply. For example, because Nd is much less common than La or Ce, mining ore for Nd produces an over-supply of La and Ce. Future demand for electric vehicles and wind turbines will rely heavily on Dy and Nd, which are used in permanent magnets. Because these technologies are important components of our strategy to transition to a low carbon economy, the supply of REEs has been highlighted as critical to achieving carbon neutrality. It has been predicted that the supply of Dy and Nd will need to increase by 700% and 2600%, respectively, between 2010 and 2035 if atmospheric CO2 is to be stabilised at 450 ppm using current technologies (13). The REEs of greatest concern are therefore those that are of high importance to clean energy and also have a high supply risk (Nd, Dy, Eu and Tb). One of the best opportunities for recycling is permanent magnets which are based upon neodymium-iron-boron (NdFeB) alloys and also contain Dy, Pr, Gd and Tb, representing some of the most critical REEs in terms of future supply risk. There are also opportunities for recycling REEs from nickel metal-hydride batteries and lamp phosphors.
Several studies have sought to identify the opportunities for recovering REEs from mixed waste streams. Commercially recoverable quantities of REEs, primarily Ce, have been detected in landfills by Gutiérrez-Gutiérrez et al. (14), but recovery would only be economically feasible if other valuable materials were also recovered simultaneously. Bottom and fly ash from municipal solid waste incinerators may prove to be a promising source, as strong enrichment of Eu, La, Gd, and Tb correlates with phosphorus pentoxide (P2O5) and indicates that the source of REE anomalies in the ash are likely from REE phosphors present in fluorescent materials (15). REEs were found to be present at several orders of magnitude higher in acid mine drainage water than natural water (16) presenting an opportunity to provide a modest but continuous supply of potentially valuable resources as a by-product of otherwise costly remediation activities. Sewage sludge only seems to be a promising source of Gd though (presumably due to its use as a MRI contrasting agent), since other REEs show enrichment factors near unity indicating geogenic origin (17).
The maturation of technologies for recovering REEs from waste streams and recycling them in ‘end-of-life’ products may result in lower concentrations released into environmental media. However, this must be contrasted against an anticipated exponential growth in their use in consumable products. It is unlikely that the environmental impact of REEs, which seems to be lower than many other metallic elements, will drive increases in recycling and recovery. Therefore, the adoption of technologies to exploit “second-hand” REE resources will likely be dependent on the price of REEs on the global market.
References
1. ECG Bulletin, July 2013, pp. 3-5 and pp. 10-20, see http://www.rsc.org/images/ECG-Bulletin-July-2013_tcm18-233368.pdf.
2. G. Pagano et al., Rare earth elements in human and animal health: State of art and research priorities, Environmental Research 142, 215-220 (2015).
3. I. Olmez, E. R. Sholkovitz, D. Hermann, R. P. Eganhouse, Rare earth elements in sediments off southern California: a new anthropogenic indicator. Environmental Science & Technology 25, 310-316 (1991).
4. S. Kulaksız, M. Bau, Rare earth elements in the Rhine River, Germany: first case of anthropogenic lanthanum as a dissolved microcontaminant in the hydrosphere, Environment International 37, 973-979 (2011).
5. Y. Nozaki, D. Lerche, D. S. Alibo, M. Tsutsumi, Dissolved indium and rare earth elements in three Japanese rivers and Tokyo Bay: evidence for anthropogenic Gd and In, Geochimica et Cosmochimica Acta 64, 3975-3982 (2000).
6. P. L. Verplanck et al., Evaluating the behavior of gadolinium and other rare earth elements through large metropolitan sewage treatment plants. Environmental Science & Technology 44, 3876-3882 (2010).
7. V. González et al., Lanthanide ecotoxicity: First attempt to measure environmental risk for aquatic organisms, Environmental Pollution 199, 139-147 (2015).
8. V. Gonzalez, D. A. Vignati, C. Leyval, L. Giamberini, Environmental fate and ecotoxicity of lanthanides: are they a uniform group beyond chemistry?, Environment International 71, 148-157 (2014).
9. G. Pagano, M. Guida, F. Tommasi, R. Oral, Health effects and toxicity mechanisms of rare earth elements—Knowledge gaps and research prospects, Ecotoxicology and Environmental Safety 115, 40-48 (2015).
10. B. Wei, Y. Li, H. Li, J. Yu, B. Ye, T. Liang, Rare earth elements in human hair from a mining area of China. Ecotoxicology and Environmental Safety 96, 118-123 (2013).
11. X. Li, Z. Chen, Z. Chen, Y. Zhang, A human health risk assessment of rare earth elements in soil and vegetables from a mining area in Fujian Province, Southeast China, Chemosphere 93, 1240-1246 (2013).
12. K. Binnemans et al., Recycling of rare earths: a critical review, Journal of Cleaner Production 51, 1-22 (2013).
13. E. Alonso et al., Evaluating rare earth element availability: A case with revolutionary demand from clean technologies, Environmental Science & Technology 46, 3406-3414 (2012).
14. S. C. Gutiérrez-Gutiérrez, F. Coulon, Y. Jiang, S. Wagland, Rare earth elements and critical metal content of extracted landfilled material and potential recovery opportunities, Waste Management 42, 128-136 (2015).
15. V. Funari, S. N. H. Bokhari, L. Vigliotti, T. Meisel, R. Braga, The rare earth elements in municipal solid waste incinerators ash and promising tools for their prospecting, Journal of Hazardous Materials 301, 471-479 (2016).
16. C. Ayora et al., Recovery of Rare Earth Elements and Yttrium from Passive-Remediation Systems of Acid Mine Drainage, Environmental Science & Technology 50, 8255-8262 (2016).
17. P. Westerhoff et al., Characterization, recovery opportunities, and valuation of metals in municipal sludges from US wastewater treatment plants nationwide. Environmental Science & Technology 49, 9479-9488 (2015).
