Transition from NORM to TENORM

Oil and gas deposits contain naturally-occurring radionuclides, such as the 238U series, 232Th series and 40K, which have been identified as Naturally Occurring Radioactive Materials, abbreviated as ‘NORM’ (Attallah et al., 2020). The operation of oil and gas equipment involves the precipitation of alkaline earth metals such as sulfates, carbonate and silicates, resulting in the production of TENORM (Abdelbary et al., 2019). The radionuclides produced by the aforementioned isotopes and their decay products then often end up either dissolved or suspended in produced water, resulting in various forms of waste such as sludge, mineral scales, and thin-films (El Afifi et al., 2023).

Radioactivity of TENORM wastes

Current literature posit that TENORM wastes contain radioactive concentrations of Ra-226 levels significantly higher than what is permitted by the International Atomic Energy Agency (IAEA) (Hilal et al., 2014). Due to its extremely long half-life and high abundance of its parent nuclide (238U), 226Ra remains as the superabundant radium isotope found on TENORM wastes (Attallah et al., 2019). Similar to the nature of phosphogypsum discussed in the previous blog article, these TENORM wastes resulting from oil and gas processing also give rise to the generation of radon gas, which produces alpha particles as it decays (Attallah et al., 2019). According to Alfifi et al (2023), the radiological hazard parameters of scale and sludge residues and produced water—specifically pertaining to 222Rn levels—have been found to exceed well beyond the allowed safe limits.

Public exposure to TENORM wastes can occur through direct exposure pathways or through inhalation and ingestion from contaminated soil and water sources arising from the disposal of TENORM wastes (ALNabhani et al., 2016). In light of this, the International Atomic Energy Agency (IAEA) has put forth safety standards for industrial activities which involve NORM, specifically with regard to radiation protection and radioactive waste management for the oil and gas industry (Ali et al., 2021). Unfortunately, these proposed safety standards are inadequate at reducing radiation exposure risks, with current TENORM waste disposal methods often exacerbating the distribution of radionuclides and their decay products. ALNabhani et al (2017) posit that waste products from oil and gas production—of which contain varying levels of TENORM—are often disposed of above ground or underground, exposing workers to radiation. An elaboration of these disposal methods are illustrated in the figure below:

Lack of management strategies

While the presence of TENORM in wastes generated by the oil and gas industry is not a new discovery, there is still insufficient research and understanding on the impacts of TENORM on public health and safety (ALNahbani et al., 2016). Furthermore, while the IAEA has proposed recommended measures, the implementation of these measures vary across governing bodies, and there is still an absence of a standardised set of regulations with regard to TENORM management and disposal (Ault et al., 2014).

In the US, sludge containing TENORM contents is first dewatered and stored in tanks for later disposal, while produced waters are injected into deep wells, and scale is sandblasted with water, and the removed scale is then stored in drums for later disposal (US Environmental Protection Agency, 2015). Elsewhere, the open dumping of produced water is legal in Brazil, and for several European countries as well (Landa, 2007).

References

Abdelbary, H. M., Elsofany, E. A., Mohamed, Y. T., Abo-Aly, M. M., & Attallah, M. F. (2019). Characterization and radiological impacts assessment of scale TENORM waste produced from oil and natural gas production in Egypt. Environmental Science and Pollution Research, 26(30), 30836–30846. https://doi.org/10.1007/s11356-019-06183-x

AL Nabhani, K., Khan, F., & Yang, M. (2016). Technologically Enhanced Naturally Occurring Radioactive Materials in oil and gas production: A silent killer. Process Safety and Environmental Protection, 99, 237–247. https://doi.org/10.1016/j.psep.2015.09.014

Ali, M. M. M., Li, Z., Zhao, H., Rawashdeh, A., Al Hassan, M., & Ado, M. (2021). Characterization of the health and environmental radiological effects of TENORM and radiation hazard indicators in petroleum waste –Yemen. Process Safety and Environmental Protection, 146, 451–463. https://doi.org/10.1016/j.psep.2020.11.016

Ali, M. M. M., Zhao, H., Li, Z., & Maglas, N. N. M. (n.d.). Concentrations of TENORMs in the petroleum industry and their environmental and health effects. RSC Advances, 9(67), 39201–39229. https://doi.org/10.1039/c9ra06086c

ALNabhani, K., Khan, F., & Yang, M. (2016). The importance of public participation in legislation of TENORM risk management in the oil and gas industry. Process Safety and Environmental Protection, 102, 606–614. https://doi.org/10.1016/j.psep.2016.04.030

ALNabhani, K., Khan, F., & Yang, M. (2017). Management of TENORMs produced during oil and gas operation. Journal of Loss Prevention in the Process Industries, 47, 161–168. https://doi.org/10.1016/j.jlp.2017.03.016

Attallah, M. F., Abdelbary, H. M., Elsofany, E. A., Mohamed, Y. T., & Abo-Aly, M. M. (2020). Radiation safety and environmental impact assessment of sludge TENORM waste produced from petroleum industry in Egypt. Process Safety and Environmental Protection, 142, 308–316. https://doi.org/10.1016/j.psep.2020.06.012

Attallah, M. F., Hamed, M. M., & El Afifi, E. M. (2019). Remediation of TENORM scale waste generated from petroleum industry using single and mixed micelles solutions. Journal of Molecular Liquids, 294, 111565. https://doi.org/10.1016/j.molliq.2019.111565

Ault, T., Krahn, S., & Croff, A. (2015). Radiological Impacts and Regulation of Rare Earth Elements in Non-Nuclear Energy Production. Energies, 8(3), Article 3. https://doi.org/10.3390/en8032066

El Afifi, E. M., Mansy, M. S., & Hilal, M. A. (2023). Radiochemical signature of radium-isotopes and some radiological hazard parameters in TENORM waste associated with petroleum production: A review study. Journal of Environmental Radioactivity, 256, 107042. https://doi.org/10.1016/j.jenvrad.2022.107042

Hilal, M. A., Attallah, M. F., Mohamed, G. Y., & Fayez-Hassan, M. (2014). Evaluation of radiation hazard potential of TENORM waste from oil and natural gas production. Journal of Environmental Radioactivity, 136, 121–126. https://doi.org/10.1016/j.jenvrad.2014.05.016

Landa, E. R. (2007). Naturally occurring radionuclides from industrial sources: Characteristics and fate in the environment. In G. Shaw (Ed.), Radioactivity in the Environment (Vol. 10, pp. 211–237). Elsevier. https://doi.org/10.1016/S1569-4860(06)10010-8

US EPA, O. (2015, April 22). TENORM: Oil and Gas Production Wastes [Overviews and Factsheets]. https://www.epa.gov/radiation/tenorm-oil-and-gas-production-wastes

Phosphate rock mining

Phosphate rock is predominantly mined and processed to obtain phosphorus – a key ingredient used to enhance plant productivity (Chen & Graedel, 2015). Currently, there are two predominant mining approaches for phosphate rock: open-pit mining, and underground hard-rock mining (Steiner et al., 2015). Following extraction, the phosphate rock is then processed to produce phosphoric acid, which is then converted to phosphate fertilisers (Liang et al., 2017). Here, phosphate rock undergoes wet chemical processing, which uses sulfuric acid to first digest phosphate minerals (Liang et al., 2018).

As shown Fig. 2, the wet process of phosphate rock generates two products: phosphoric acid and phosphogypsum (principally CaSO4.2H2O), with the latter being a waste product. Phosphogypsum, while mostly made up of calcium sulfate dihydrate, contains several impurities such as phosphoric acid, phosphates, fluorides and organic matter, and is usually in the form of a grey, damp, powder or silt (Saadaoui et al., 2017). Additionally, phosphogypsum also contains the bulk of naturally-occuring uranium, thorium, radium, and heavy metals (US Environmental Protection Agency, 2019). While the characteristics of phosphogypsum is heavily dependent on the phosphate ore composition and quality, wet processing has been found to result in the selective separation and concentration of naturally-occurring radium in phosphogypsum (Sahu et al., 2014). To elaborate, about 80% or radium is concentrated in phosphogypsum, while almost 86% and 70% of uranium and thorium respectively are concentrated in phosphoric acid instead (Tayibi et al., 2009). This has therefore raised concerns about the leaching of radioactive elements in disposal sites, as well as the release of radon gas into the atmosphere.

Global phosphogypsum production is estimated to be around 280 million tonnes per year, with 28% of it disposed into water bodies, and 58% of it stored in tailing ponds (Turner et al., 2022). The disposal of phosphogypsum in water bodies or storage in ponds or leaps are often done without purification (Rashad, 2017), thereby resulting in extensive contamination. Currently, phosphogypsum is managed using wet stacking, in which filtered phosphogypsum is mixed with water and pumped into settling ponds, and the solid residue is then placed into stacks (Turner et al., 2022). These phosphogypsum stacks pose multiple environmental risks, as they are often not watertight nor covered with any inert material (Tayibi et al., 2009). The percolation of water often induces edge outflows that consist of toxic wastewater and leachates, introducing heavy metals and radionuclides into the waters and sediments of estuarine systems (Guerrero et al., 2019). This is exemplified in the outflow of leachates in the phosphogypsum stacks of Huevela, a region in southwestern Spain, which has since led to the deep pollution of underlying salt-marsh sediments (Guerrero et al., 2019). 

The open storage of these phosphogypsum stacks has also been found to contain significant levels of radioactivity due to high radon concentrations in the waste stacks. This has been posited to lead to an increase in radon inhalation rates and consequently, atmospheric radon concentrations (López-Coto et al., 2014). As discussed in a previous blog, radon exposure and inhalation has been found to lead to lung cancer, with radon itself being responsible for half of the total effective dose received by the population.

References

Ben Abdelouahed, H., & Reguigui, N. (2011). Radiotracer investigation of phosphoric acid and phosphatic fertilizers production process. Journal of Radioanalytical and Nuclear Chemistry – J RADIOANAL NUCL CHEM, 289, 103–111. https://doi.org/10.1007/s10967-011-1035-9

Chen, M., & Graedel, T. E. (2015a). The potential for mining trace elements from phosphate rock. Journal of Cleaner Production, 91, 337–346. https://doi.org/10.1016/j.jclepro.2014.12.042

Chen, M., & Graedel, T. E. (2015b). The potential for mining trace elements from phosphate rock. Journal of Cleaner Production, 91, 337–346. https://doi.org/10.1016/j.jclepro.2014.12.042

Guerrero, J. L., Gutiérrez-Álvarez, I., Mosqueda, F., Olías, M., García-Tenorio, R., & Bolívar, J. P. (2019). Pollution evaluation on the salt-marshes under the phosphogypsum stacks of Huelva due to deep leachates. Chemosphere, 230, 219–229. https://doi.org/10.1016/j.chemosphere.2019.04.212

Liang, H., Zhang, P., Jin, Z., & DePaoli, D. (2017). Rare-earth leaching from Florida phosphate rock in wet-process phosphoric acid production. Minerals & Metallurgical Processing, 34(3), 146–153. https://doi.org/10.19150/mmp.7615

Liang, H., Zhang, P., Jin, Z., & DePaoli, D. W. (2018). Rare Earth and Phosphorus Leaching from a Flotation Tailings of Florida Phosphate Rock. Minerals, 8(9), Article 9. https://doi.org/10.3390/min8090416

López-Coto, I., Mas, J. L., Vargas, A., & Bolívar, J. P. (2014). Studying radon exhalation rates variability from phosphogypsum piles in the SW of Spain. Journal of Hazardous Materials, 280, 464–471. https://doi.org/10.1016/j.jhazmat.2014.07.025

(PDF) Radiotracer investigation of phosphoric acid and phosphatic fertilizers production process. (n.d.). Retrieved March 24, 2023, from https://www.researchgate.net/publication/251415281_Radiotracer_investigation_of_phosphoric_acid_and_phosphatic_fertilizers_production_process

Rashad, A. M. (2017). Phosphogypsum as a construction material. Journal of Cleaner Production, 166, 732–743. https://doi.org/10.1016/j.jclepro.2017.08.049

Saadaoui, E., Ghazel, N., Ben Romdhane, C., & Massoudi, N. (2017). Phosphogypsum: Potential uses and problems – a review. International Journal of Environmental Studies, 74(4), 558–567. https://doi.org/10.1080/00207233.2017.1330582

Sahu, S. K., Ajmal, P. Y., Bhangare, R. C., Tiwari, M., & Pandit, G. G. (2014). Natural radioactivity assessment of a phosphate fertilizer plant area. Journal of Radiation Research and Applied Sciences, 7(1), 123–128. https://doi.org/10.1016/j.jrras.2014.01.001

Silva, L. F. O., Oliveira, M. L. S., Crissien, T. J., Santosh, M., Bolivar, J., Shao, L., Dotto, G. L., Gasparotto, J., & Schindler, M. (2022). A review on the environmental impact of phosphogypsum and potential health impacts through the release of nanoparticles. Chemosphere, 286, 131513. https://doi.org/10.1016/j.chemosphere.2021.131513

Steiner, G., Geissler, B., Watson, I., & Mew, M. C. (2015). Efficiency developments in phosphate rock mining over the last three decades. Resources, Conservation and Recycling, 105, 235–245. https://doi.org/10.1016/j.resconrec.2015.10.004

Tayibi, H., Choura, M., López, F. A., Alguacil, F. J., & López-Delgado, A. (2009). Environmental impact and management of phosphogypsum. Journal of Environmental Management, 90(8), 2377–2386. https://doi.org/10.1016/j.jenvman.2009.03.007

Turner, L. E., Dhar, A., Naeth, M. A., Chanasyk, D. S., & Nichol, C. K. (2022). Effect of soil capping depth on phosphogypsum stack revegetation. Environmental Science and Pollution Research, 29(33), 50166–50176. https://doi.org/10.1007/s11356-022-19420-7

US EPA, O. (2019, February 11). What kinds of consumer products contain radioactive materials? [Overviews and Factsheets]. https://www.epa.gov/radiation/what-kinds-consumer-products-contain-radioactive-materials

Introduction to rare earth elements

Rare earth elements (REEs) refer to the group of 15 lanthanides, and are usually categorised into either of the two groups: light rare earth minerals (LREEs) and heavy rare earth minerals (HREEs), based on their atomic weight as shown in the table below:

Lanthanides (elements with atomic numbers from 58 to 71)
LREEs	HREEs
Lanthanum (La) Cerium (Ce) Praseodymium (Pr) Neodymium (Nd) Promethium (Pm) Samarium (Sm)	Europium (Eu) Gadolinium (Gd) Terbium (Tb) Dysprosium (Dy) Holmium (Ho) Erbium (Er) Thulium (Tm) Ytterbium (Yb) Lutetium (Lu)
Non-lanthanide elements: Yttrium (Y) Scandium (Sc)

Table 1: List of the 15 REEs (Source: Riesgo García et al., 2017)

These REEs often make up key components in defence technologies, such as lasers and satellites (Royer-Lavallée et al., 2020) and in the renewable energy sector, such as wind turbines and hybrid cars (EuRare, 2017, as cited in Costis et al., 2021). To elaborate further, Balaram (2019) further breaks down the main uses of REEs into the following categories:

REE deposits and the extraction of REEs

REEs are naturally occurring in geological deposits, and develop in virtually all major rock types, and multiple significant REE deposits can be found worldwide:

Despite the natural abundance of REEs, in which the concentrations of REEs far exceed the concentrations of many other primary produced metals that are mined on an industrial scale, REEs are rarely found in mineable ore deposits (Balachandran, 2014). Furthermore, REEs often occur collectively and are found in various minerals in the Earth’s crust, such as silicate, carbonate or phosphate minerals (Hoshino et al., 2016; Royer-Lavallée et al., 2020). In essence, the nature of REEs co-occurring with mineral ores of base metals to form rare earth ores, along with the high reactivity of REEs with most non-metals (Gwenzi et al., 2018) pose additional challenges to the extraction of REEs.

There are two main methods for REE mining: the first involves the removal of topsoil and creating a leaching pond, in which chemicals are added to the extract earth to separate metals. The second method involves drilling holes into the ground and pumping chemicals such as ammonium sulfate and ammonium chloride into the earth, which in turn also creates a leaching pond (Nayar, 2021; Standaert, 2019). Given that both approaches involve creating a leaching pond containing harsh, toxic chemicals, these extraction methods pose huge risks of groundwater contamination, which can then in turn affect entire waterways (Nayar, 2021). 

Following extraction, the REEs ores often undergo leaching processes to further refine and isolate the rare earth metals. This involves dissolving the ores in a suitable leaching agent before extracting the rare earth metals using methods such as precipitation, fractional crystallisation, and chromatography (Shahbaz, 2022).

Problems associated with REE mining activities

1. Acid Mine Drainage

Acid mine drainage (AMD) refers to the formation and movement of highly acidic water rich in heavy metals. AMD contains sulfuric acid – the result of chemical reactions between surface water and shallow subsurface water with rocks that contain sulfur-bearing minerals (United States Environmental Protection Agency, 2022). The extraction of REEs often results in the generation of AMD, and the toxicity of AMD can lead to the degradation of soils, water reservoirs, and rivers, which can in turn jeopardise the ecosystems present in these biospheres (Gomes et al., 2022). 

2. Waste Products from REE-related leaching processes

The use of concentrated hydrochloric acid, sulfuric acid, and nitric acid to leach REE from REE-bearing ores via hydrometallurgical processes can also pose significant environmental threats (Edahbi et al., 2019). The process of leaching REE-bearing ores generates two main waste products: waste rocks that are extracted to reach the orebody, and tailings which consist of finely ground ore and other chemicals used during the hydrometallurgical processes (Filho, 2016). Similar to AMD, the interaction of these aforementioned waste products with water and oxygen can give rise to the oxidation of sulfide minerals (Nordstrom, 2012).

Mine tailings are often dumped in tailing ponds, and the REE concentrations in these residues are often extremely high. These high concentrations, coupled with the fine particle sizes of these tailings, can result in radioactive pollution due to the introduction and diffusion of radioactive thorium-containing dust (Binnemans et al., 2015) over large areas. This radioactive dust is airborne, and can also migrate through soils and water bodies (Krasavtseva et al., 2021). Apart from becoming a source of ionising radiation, the radioactive dust also results in particulate matter formation, posing significant health risks (Marx et al., 2018)

Case study: Bayan Obo Rare Earth Mine

The Bayan Obo Ore Deposit, located in Inner Mongolia, China, is the world’s largest REE deposit, with China making up 42% of the global REE reserve base, in which 80% of China’s LREE resources are found in the Bayan Obo region (Fan et al., 2016). This ore deposit is also known as one of the most heavily polluted places in the world, due to the mismanagement of mining waste products, resulting in extensive contamination of farmland, water supplies, and air (Gramling, 2023). 

Due to the lack of environmental awareness and inadequate technologies to manage REE mining waste, mine tailings and other waste materials generated from China’s mining activities have resulted in a series of environmental impacts. These include the creation of derelict lands, known as ‘mining brownfields’ in China, as well as extensive soil pollution and huge economic loss (Pan & Li, 2016). For the Bayan Obo Deposit, 90% of the tailings are stored in the tailing dam of the Baotou Iron and Steel Group Company—established in 1965 with poor support capacity—and has since resulted in the development of huge brownfields (Pan & Li, 2016). This has been attributed to a large amount of these tailings being discharged directly into Inner Mongolia’s grassland ecosystem (Guo et al., 2013).

Apart from the poor structural integrity of these tailing ponds, the open stockpiling of rare earth tailings in Bayan Obo has also resulted in the deterioration of grassland resources, air pollution, and water and soil pollution via erosion and leaching processes (Guo et al., 2013). Another point of concern for communities living in the Bayan Obo region is that the main source of drinking water is groundwater located in the vicinity of the Bayan Obo mining area, therefore rendering drinking water as one of the REE exposure pathways (Liang et al., 2018). The open tailing dumps and surface mining activities have also resulted in higher concentrations of total suspended particulate (TSP) and PM10 in the Bayan Obo region, and these airborne pollutants have been found to be associated with respiratory and cardiovascular diseases, skin cancer, and gastrointestinal issues (Wang et al., 2014).

References

Balachandran, G. (2014). Case Study 1—Extraction of Rare Earths for Advanced Applications. In S. Seetharaman (Ed.), Treatise on Process Metallurgy (pp. 1291–1340). Elsevier. https://doi.org/10.1016/B978-0-08-096988-6.09983-1

Balaram, V. (2019). Rare earth elements: A review of applications, occurrence, exploration, analysis, recycling, and environmental impact. Geoscience Frontiers, 10(4), 1285–1303. https://doi.org/10.1016/j.gsf.2018.12.005

Binnemans, K., Jones, P. T., Blanpain, B., Van Gerven, T., & Pontikes, Y. (2015). Towards zero-waste valorisation of rare-earth-containing industrial process residues: A critical review. Journal of Cleaner Production, 99, 17–38. https://doi.org/10.1016/j.jclepro.2015.02.089

China Wrestles with the Toxic Aftermath of Rare Earth Mining. (n.d.). Yale E360. Retrieved March 17, 2023, from https://e360.yale.edu/features/china-wrestles-with-the-toxic-aftermath-of-rare-earth-mining

Costis, S., Mueller, K. K., Coudert, L., Neculita, C. M., Reynier, N., & Blais, J.-F. (2021). Recovery potential of rare earth elements from mining and industrial residues: A review and cases studies. Journal of Geochemical Exploration, 221, 106699. https://doi.org/10.1016/j.gexplo.2020.106699

Edahbi, M., Plante, B., & Benzaazoua, M. (2019). Environmental challenges and identification of the knowledge gaps associated with REE mine wastes management. Journal of Cleaner Production, 212, 1232–1241. https://doi.org/10.1016/j.jclepro.2018.11.228

EJOLT. (n.d.). Bayan Obo world biggest rare earths mine, Baotou, Inner Mongolia, China | EJAtlas. Environmental Justice Atlas. Retrieved March 18, 2023, from https://ejatlas.org/conflict/bayan-obo-world-biggest-rare-earths-mine-baogang-group-baotou-inner-mongolia-china

Fan, H.-R., Yang, K.-F., Hu, F.-F., Liu, S., & Wang, K.-Y. (2016). The giant Bayan Obo REE-Nb-Fe deposit, China: Controversy and ore genesis. Geoscience Frontiers, 7(3), 335–344. https://doi.org/10.1016/j.gsf.2015.11.005

Gomes, P., Valente, T., Marques, R., Prudêncio, M. I., & Pamplona, J. (2022). Rare earth elements—Source and evolution in an aquatic system dominated by mine-Influenced waters. Journal of Environmental Management, 322, 116125. https://doi.org/10.1016/j.jenvman.2022.116125

Guo, W., Zhao, R., Zhao, W., Fu, R., Guo, J., Bi, N., & Zhang, J. (2013). Effects of arbuscular mycorrhizal fungi on maize (Zea mays L.) and sorghum (Sorghum bicolor L. Moench) grown in rare earth elements of mine tailings. Applied Soil Ecology, 72, 85–92. https://doi.org/10.1016/j.apsoil.2013.06.001

Gwenzi, W., Mangori, L., Danha, C., Chaukura, N., Dunjana, N., & Sanganyado, E. (2018). Sources, behaviour, and environmental and human health risks of high-technology rare earth elements as emerging contaminants. Science of The Total Environment, 636, 299–313. https://doi.org/10.1016/j.scitotenv.2018.04.235

Hoshino, M., Sanematsu, K., & Watanabe, Y. (2016). Chapter 279—REE Mineralogy and Resources. In B. Jean-Claude & P. Vitalij K. (Eds.), Handbook on the Physics and Chemistry of Rare Earths (Vol. 49, pp. 129–291). Elsevier. https://doi.org/10.1016/bs.hpcre.2016.03.006

Kirk Nordstrom, D. (2012). Models, validation, and applied geochemistry: Issues in science, communication, and philosophy. Applied Geochemistry, 27(10), 1899–1919. https://doi.org/10.1016/j.apgeochem.2012.07.007

Krasavtseva, E., Maksimova, V., & Makarov, D. (2021). Conditions Affecting the Release of Heavy and Rare Earth Metals from the Mine Tailings Kola Subarctic. Toxics, 9(7), 163. https://doi.org/10.3390/toxics9070163

Liang, Q., Yin, H., Li, J., Zhang, L., Hou, R., & Wang, S. (2018). Investigation of rare earth elements in urine and drinking water of children in mining area. Medicine, 97(40), e12717. https://doi.org/10.1097/MD.0000000000012717

Marx, J., Schreiber, A., Zapp, P., & Walachowicz, F. (2018). Comparative Life Cycle Assessment of NdFeB Permanent Magnet Production from Different Rare Earth Deposits. ACS Sustainable Chemistry & Engineering. https://doi.org/10.1021/acssuschemeng.7b04165

Not So “Green” Technology: The Complicated Legacy of Rare Earth Mining. (2021, August 12). Harvard International Review. https://hir.harvard.edu/not-so-green-technology-the-complicated-legacy-of-rare-earth-mining/

Pan, Y., & Li, H. (2016). Investigating Heavy Metal Pollution in Mining Brownfield and Its Policy Implications: A Case Study of the Bayan Obo Rare Earth Mine, Inner Mongolia, China. Environmental Management, 57(4), 879–893. https://doi.org/10.1007/s00267-016-0658-6

Peiravi, M., Dehghani, F., Ackah, L., Baharlouei, A., Godbold, J., Liu, J., Mohanty, M., & Ghosh, T. (2021). A Review of Rare-Earth Elements Extraction with Emphasis on Non-conventional Sources: Coal and Coal Byproducts, Iron Ore Tailings, Apatite, and Phosphate Byproducts. Mining, Metallurgy & Exploration, 38(1), 1–26. https://doi.org/10.1007/s42461-020-00307-5

Puko, T. (2020, April 26). Pentagon Invests in Strategic Metals Mine, Seeking to Blunt Chinese Dominance. Wall Street Journal. https://www.wsj.com/articles/pentagon-invests-in-strategic-metals-mine-seeking-to-blunt-chinese-dominance-11587924001

Rare earth mining may be key to our renewable energy future. But at what cost? (2023, January 11). https://www.sciencenews.org/article/rare-earth-mining-renewable-energy-future

Riesgo García, M. V., Krzemień, A., Manzanedo del Campo, M. Á., Menéndez Álvarez, M., & Gent, M. R. (2017). Rare earth elements mining investment: It is not all about China. Resources Policy, 53, 66–76. https://doi.org/10.1016/j.resourpol.2017.05.004

Royer-Lavallée, A., Neculita, C. M., & Coudert, L. (2020). Removal and potential recovery of rare earth elements from mine water. Journal of Industrial and Engineering Chemistry, 89, 47–57. https://doi.org/10.1016/j.jiec.2020.06.010

Shahbaz, A. (2022). A systematic review on leaching of rare earth metals from primary and secondary sources. Minerals Engineering, 184, 107632. https://doi.org/10.1016/j.mineng.2022.107632

Smith, M. P., Moore, K., Kavecsánszki, D., Finch, A. A., Kynicky, J., & Wall, F. (2016). From mantle to critical zone: A review of large and giant sized deposits of the rare earth elements. Geoscience Frontiers, 7(3), 315–334. https://doi.org/10.1016/j.gsf.2015.12.006

US EPA, O. (2015, September 15). Abandoned Mine Drainage [Overviews and Factsheets]. https://www.epa.gov/nps/abandoned-mine-drainage

Wang, L., Liang, T., Zhang, Q., & Li, K. (2014). Rare earth element components in atmospheric particulates in the Bayan Obo mine region. Environmental Research, 131, 64–70. https://doi.org/10.1016/j.envres.2014.02.006

Yuan, J., Ding, Z., Bi, Y., Li, J., Wen, S., & Bai, S. (2022). Resource Utilization of Acid Mine Drainage (AMD): A Review. Water, 14(15), Article 15. https://doi.org/10.3390/w14152385

Anthropogenic sources of radioactive pollution in oceans are many and can include the following:

Having discussed the issue of dumping nuclear waste in a previous blog post, this blog post will focus on radioactive pollution originating from nuclear accidents—such as the Fukushima Dai-ichi Nuclear Power Plant disaster—and how radioactive particles can be dispersed by ocean currents. This blog post will then conclude with a discussion on the impacts of these radioactive substances in oceans.

Apart from the atmospheric deposition of radioactive particles via nuclear fallout, the drainage of water from nuclear power plant (NPP) reactors following NPP-related disasters has also resulted in the contamination of neighbouring water bodies. Additionally, the rain-induced wash off has from the soil has also resulted in the introduction of radioactive substances into the sea (Prants et al., 2011). To better understand these processes, we will be examining the aftermath of the Fukushima Daiichi Nuclear Power Plant (FDNPP) and how the released radionuclides were dispersed.

The events of the Fukushima nuclear accident

The Fukushima nuclear accident occurred on 11 March 2011 following a 9.0-magnitude earthquake, coupled with a resultant tsunami event (Kim et al., 2013). Following this, the Fukushima Daiichi Nuclear Power Station lost external power supplies and AC power, which then resulted in reactors and spent fuel pods losing their cooling capabilities as well. This then led to explosions in three out of the six nuclear power units, along with serious damage towards the reactor core of another power unit (Hasegawa, 2012). The Fukushima nuclear disaster was rated 7 on the International Nuclear Event Scale, and has since been deemed as the first example of a “Quake and Nuke Disaster Complex”, as well as the first major accident of a NPP located on the coast (Hasegawa, 2012). Prants et al (2011) posit that the propagation of radioactive pollution from the NPP stems from two sources: the direct discharge of nuclear wastewater from the NPP, and deposition of radioactive substances as atmospheric precipitation into the ocean.

One of the most contentious issues regarding the aftermath of the Fukushima nuclear disaster is the release of around 10,400 cubic metres of contaminated wastewater into the Pacific Ocean in order to free up storage for wastewater with even higher contamination levels (World Nuclear Association, 2022). Additionally, despite receiving criticism for the initial release of contaminated wastewater, the Japanese government has announced plans in April 2021 to release huge quantities of contaminated nuclear wastewater into the Pacific Ocean from late 2022 to early 2023, over the next 30 years (Greenpeace, 2020, as cited in Yang et al., 2022). These actions were met with considerable opposition criticism from neighbouring countries (Norio et al., 2012), given the dispersal of radioactive substances via ocean currents as shown in the figure below:

Apart from contaminating oceans, the deposition of radionuclides on Japanese soils can also lead to a possibility of introducing contaminated sediments into rivers via runoff and erosion processes (Chartin et al., 2013). Alternatively, deposited radionuclides can also be deposited via soluble media—a process known as ‘liquid wash off’—into rivers and other water bodies (Pratama, 2015). Such occurrences can prove to be problematic, with coastal rivers becoming a constant supply of contaminated sediment to the Pacific Ocean (Chartin et al., 2013).

Apart from the volume of wastewater dumped, the contamination of the marine environment following the Fukushima nuclear disaster is also characterised by the location of the coastal waters of the FDNPP, which is located in the zone at where the Kuroshio and Oyashio currents interact (Bailly du Bois et al., 2011). These currents influence the extent of which the radioactive pollution is dispersed, with the Kuroshio current carrying the radioactive plume towards the centre of the Pacific Ocean (Jayne et al., 2009, as cited in Bailly du Bois et al., 2011). 

Impacts and concerns

One of the biggest concerns is the introduction of Caesium-137 and Caesium-134 into the marine environment, as these Cs isotopes are essentially soluble in seawater and can be transported over long distances by marine currents and dissipated throughout the ocean water masses (Sanchez-Cabeza et al., 2011, as cited in Bailly du Bois et al., 2011). Additionally, other radionuclides also tend to bind to suspended particles, resulting in sedimentary contamination as they deposit onto the seafloor (Evrard et al., 2011). As such, the nature of these radionuclides not only makes it more challenging to track and monitor the dispersal of radioactive substances in water bodies, but also that it is easier for these radioactive substances to affect marine life on a larger scale. The latter has been reflected in how 40% of fish species between April 2011 and April 2012 were found to have exceeded the Japanese radioactive regulatory limit of 100 Bq/kg-wet for radioactive Cs (Wada et al., 2013). In response to this, the Japanese government stopped the distribution of contaminated fishery products and contaminated feed for aquaculture (Morita et al., 2019).

A simulation conducted by Behrens et al., (2012, as cited in Koo et al., 2014) estimates that the dispersion and dilution of radioactive substances by ocean current activity would help to decrease peak radioactivity in the seawater off Fukushima gradually. However, while the ocean is able to dilute and disperse radioactive substances due to its large volume, the long half-life radionuclides, such as Caesium-137 and Caesium-134, are likely to still remain in the marine environment for prolonged periods (Yu et al., 2015). Both short- and long-lived radioactive elements can be absorbed by plankton and kelp, which can then accumulate in marine animals across the food chain (Grossman, 2011), ultimately affecting humans who consume them. 

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