Chapter 9 Health Risks Associated with Radionuclides in Soil Materials Abstract Radionulides in soils from natural and man-made sources constitute a direct route of exposure to humans. The most significant part of the total exposure is due to natural radiation. Soil- or rock-borne radionuclides generate a significant component of the background radiation people are exposed to. Naturally occurring radionuclides with half-lives comparable with the age of the earth and their corresponding decay products existing in terrestrial material, such as thorium (232Th), uranium (238U, 235U) and potassium (40K), are of great importance. Their spatial distribution depends on geological parent materials and plays an important role for radiation protection. Another source of exposure to natural radiation is expressed through high energy cosmic ray particles in earth’s atmosphere. Additional amounts of natural radionuclides are released into the environment through human activities such as mining and milling of mineral ores, processing and enrichment, nuclear fuel fabrications, and handling of the fuel cycle tail end products. Radionuclides produced by humans originate from nuclear industrial activities, nuclear reactor accidents, or military activities. The most important man-made radionuclide is cesium (137Cs) with a half-life of 30.17 years, which is released from nuclear fission and activation processes. A large amount of 137Cs was released into the atmosphere during the nuclear weapons tests in the 1950s and 1960s. Atmospheric deposition of this 137C has made it a typical background component of northern hemisphere top soils. The most severe civil nuclear reactor accidents, which also released large quantities of 137Cs, occurred at Chernobyl (April 26, 1986) and Fukushima Daiichi (March 11, 2011) power stations. During the years after those weapon tests and reactor accidents, the bioactivity and environmental mobility of 137Cs declined markedly, resulting in large changes in contamination of soils, surface water and foodstuffs. Radiation is harmful to life. Depending on the dose, it can cause cancer, genetic and organs damages, cell killing as well as rapid death. The potential damage from an absorbed dose depends on the type of radiation and the sensitivity of different tissues and organs. Radionuclides can generally be internalized through inhalation, ingestion, wound contamination and percutaneous absorption. If radionuclide contamination is likely, the first step is to remove sources of potential contamination. External decontamination procedures are vital in reducing the risk of additional internal contamination events. Isotope-specific pharmacological treatments can begin once thorough external decontamination is performed. For remediation of © Springer Science+Business Media B.V. 2018 R. Nieder et al., Soil Components and Human Health, https://doi.org/10.1007/978-94-024-1222-2_9 451 452 9 Health Risks Associated with Radionuclides in Soil Materials radionuclide-contaminated sites a number or remediation techniques is available. The choice of the technique requires consideration of performance, reliability and maintenance requirements, cost, available supporting infrastructure, risk to workers and public during implementation, environmental impact, future land use and regulatory and community acceptance. Keywords Alpha particles • Beta particles • Gamma particles • Cosmogenic radiation • Natural terrestrial radiation • Nuclear weapons tests • Nuclear fuel cycle • Nuclear accidents • Radionuclides in food and water • Radiation exposure • Clinical effects • Therapy • Remediation of radionuclide-contaminated sites 9.1 Types of Radiation Unstable atoms are said to be radioactive. In order to reach stability, these atoms give off, or emit, the excess energy or mass. These emissions are called radiation. The three main types of ionizing radiation are called alpha (α), beta (β) and gamma (γ) radiation. 9.1.1 Alpha Particles Alpha radiation is a particle, consisting of two protons and two neutrons, making it identical to a helium atom, but without the electrons. Compared to a β-particle the α-particle has a large size. It travels very fast and thus has a large amount of energy. When a positively charged α-particle passes near an atom, it excites its electrons and pulls an electron from the atom. This is the process of ionization. With each ionization, the α-particle loses some energy and slows down. At the end of its path it will take two electrons from other atoms and become a complete helium atom, which has no biological effect. 9.1.2 Beta Particles Beta particle emission occurs when the ratio of neutrons to protons in the nucleus is too high. In this case, an excess neutron transforms into a proton and an electron. The proton stays in the nucleus and the electron is ejected energetically. This process increases the number of protons by one and decreases the number of neutrons by one. Since the number of protons in the nucleus of an atom determines the element, the conversion of a neutron to a proton changes the radionuclide to a different element. Beta particles have either a positive or a negative charge. Most β-particles are negatively charged. They are much lighter and more penetrating than α-particles. 9.2 Determination of Radioactivity: Definition of the Units Used 9.1.3 453 Gamma Particles Gamma radiation is very high-energy ionizing radiation. Gamma photons have about 10,000 times as much energy as the photons in the visible range of the electromagnetic spectrum. Different from α- and β-radiation, γ-photons are pure electromagnetic energy having no mass and no electrical charge. Because of their high energy, gamma photons travel at the speed of light and can cover hundreds to thousands of meters in air before spending their energy. 9.2 Determination of Radioactivity: Definition of the Units Used Radioactivity can be quantified in different categories. The corresponding units are briefly described. Radioactivity is expressed with the unit Bequerel (Bq), Gray (Gy) describes the absorbed dose and Sievert (Sv) is representative for the effects on health of an individual which has been exposed to radionuclides (Table 9.1). 9.2.1 Bequerel Becquerel is the unit of radioactivity measurement and one Becquerel is equivalent to one disintegration per second. One Curie (originally used unit) is equivalent to the radioactivity of 1 g of radium which corresponds to 37 billion disintegrations per second. 9.2.2 Gray Gray represents the amount of radiation absorbed by any matter. One Gray corresponds to 1 J absorbed per kilogram of matter. Originally the R€ontgen, and more recently, the Rad (Radiation Absorbed Dose) were common. One Rad is equivalent to 10 2 Gy. Table 9.1 Current and former units for quantification of radioactivity Current unit Bequerel (Bq) Gray (Gy) Sievert (Sv) Former unit Curie (Ci) Rad (rad) Rem (rem) Compiled from different sources Equivalence 1 Ci ¼ 3.7  1010 Bq 1 rad ¼ 10 2 Gy 1 rem ¼ 10 2 Sv Quantity Radioactivity Absorbed dose Biological effect 454 9.2.3 9 Health Risks Associated with Radionuclides in Soil Materials Sievert Sievert is a unit of radiation weighted dose and is used to distinguish the effects induced on living tissues by the different types of radioactive particles and ionising radiation (alpha, beta, gamma, neutron), along with the radiosensitivity differences amongst different organs and tissues. The Sievert is a risk management unit used to assess the equivalent dose and the effective dose. The equivalent dose shows the effect of the different types of radioactive particles and ionising radiation on tissues. It corresponds to the absorbed dose, expressed in Gy, multiplied by a radiation weighting factors which depend on the energy delivered by the radiation. Originally, the Rem (R€ontgen Equivalent Man) was used. One Rem is equivalent to 10 2 Sv. The effective dose, besides the different types of radioactive particles and ionising radiation on tissues, also describes the more or less radiosensitive nature of the exposed tissue or organ. It corresponds to the equivalent dose, expressed in Sv, multiplied by a tissue weighting factor which depends on the radiosensitivity of the tissue. 9.3 Naturally Occurring Radionuclides Humans are exposed to ionizing radiation spontaneously emitted by naturally occurring atomic species since his existence on the earth. Three types of radiations, alpha, beta and gamma are emitted by different radioactive materials, which differ in their energy and penetrating power. Until recent times, life on earth was exposed to radiation only from natural sources. Sources of radiation can vary from place to place and there are areas in some part of the world in that the background radiation levels have been found to be abnormally high. Such areas are referred to as high background radiation areas. For example, the coastal regions of Espirito Santo and the Morro Do Forro in Brazil (Paschoa 2000), Ramsar and Mahallat in Iran (Ghiassi-nejad et al. 2002), the Southwest coast of India (Paul et al. 1998) and Yangjiang in China (Wei and Sugahara 2000) were identified as high background radiation areas. Monazite sands have been found to be the source of high radiation levels in parts of Brazil, China, Egypt and India (UNSCEAR 2000; Paschoa 2000; Ghiassi-nejad et al. 2002) while in parts of Southwest France, uranium minerals form the source of natural radiation (Delpoux et al. 1997), and in Ramsar, the very high amounts of 226Ra and its decay products brought to the surface by hot springs (Ghiassi-nejad et al. 2002) have been found to be the source. Humans have created other sources of exposures such as radioactive waste and industrial, medical and agricultural use of radioisotopes, fallout from weapon tests and radioactive releases from nuclear reactor operations and accidents. However, the major contribution to the average annual background radiation still arises from natural sources. 9.3 Naturally Occurring Radionuclides 9.3.1 455 Cosmogenic and Terrestrial Sources of Radionuclides Exposure of human beings to naturally occurring radiation arises mainly from two different sources, the first coming directly from cosmic radiation from the outer space and the second from terrestrial radioactive materials. Exposure through naturally occurring radiation accounts for up to 85% of the total annual exposure dose received by the world population (World Nuclear Association 2011). The interactions of cosmic ray particles in the atmosphere can create a number of radioactive nuclei such as 3H (half-life: 12.3 years), 7Be (half-life: 53.2 day), 14C (half-life: 5700 years) and 22Na (half-life: 2.6 years) (National Council on Radiation Protection and Measurements 1975). Apart from the exposure from cosmic radiation, natural exposures originate mainly from the primordial radionuclides, which are spread widely and are present in almost all geological materials on earth (Wilson 1994). These radionuclides are known as Naturally Occurring Radioactive Material (NORM). Only very long-lived nuclides, with half-lives comparable to the age of the earth, and their decay products, contribute to natural radiation background in significant quantities (United Nations Scientific Committee on the Effects of Atomic Radiation 2000). The majority of naturally occurring radionuclides belong to 238U and 232Th series and the single decay 40K. Generally 235U, 87Rb and other trace elements are negligible. The decay chain of 238U (232Th) includes eight (6) alpha decays and six (4) beta decays respectively, which are often associated with gamma transitions (IAEA 2006). Radionuclides of the Th and U decay series and their half-lives are given in Fig. 9.1. An overview of the radioisotopes produced by cosmic rays and terrestrial sources is given in Table 9.2. Radionuclides, which emit alpha or beta particles may be taken into the body by ingestion or inhalation and can give rise to internal exposures. Some of these nuclear species may also emit gamma rays following their Fig. 9.1 Radionuclides of the thorium and uranium decay series 456 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.2 Radioisotopes produced by cosmic rays and from terrestrial sources Cosmogenic radioisotopes C 32 Si 39 Ar 3 H 22 Na 35 S 7 Be 37 Ar 33 P 32 P 24 Na 14 Type of radiation β β β β β, γ β γ γ β β β, γ Terrestrial radioisotopes 222 Rn (Radon) 218 Po (RaA) 214 Pb (RaB) 214 Bi (RaC) 210 Pb (RaD) 210 Bi (RaE) 210 Po (RaF) 220 Rn (Thoron) 216 Po (ThA) 212 Pb (ThB) 212 Bi (ThC) Type of radiation α α β, γ α, β, γ β β α α α β, γ α, β, γ Adapted from Ramachandran (2011) Table 9.3 Concentrations of 238U, 232Th, and 40K (ranges and averages) in rocks and soils 40 Type of rock Igneous rocks Basalt (crustal average) Granite (crustal average) Sedimentary rocks Shale, sandstones Arkose Carbonate rocks Soils (global average) Continental crust (average) 232 K K (%) Bq kg 0.8 >4.0 2.7 2–3 0.3 1.5 2.8 238 Th μg g 1 Bq kg 300 >1000 3.0–4.0 17 10–15 70 0.5–1.0 3.0 7–10 40 800 600–900 70 400 850 12 2.0 2.0 9.0 10.7 50 <8.0 8.0 37 44 3.7 1.0–2.0 2.0 1.8 2.8 40 10–25 25 22 36 1 1 U μg g 1 Bq kg 1 Compiled from National Council on Radiation Protection and Measurements (1988) radioactive decay and represent the main sources of external (whole body) exposures to humans (Watson et al. 2005). Levels of the radioactivity in soils are related to the mineral composition of the parent rock from which soils have developed and the soil forming processes. Igneous rocks, such as granite, generally exhibit higher radioactivity compared to sedimentary rocks (excluding some shales and phosphate rocks). Table 9.3 gives typical natural radioactivity concentrations in rocks and soils. A number of studies focusing on naturally occurring radioactive materials in soil media provide information on levels of background radiation. Most soils contain 40 K and nuclides of the U and Th series, with broadly varying concentrations. For example, activity concentration levels arising from 238U, 232Th and 40K in surface soils in Cyprus ranged from 0.01 to 39.3, 0.01 to 39.8, and 0.04 to 565.8 Bq kg 1, respectively (Tzotzis et al. 2004). Mean natural radioactivity levels in the Firtina Valley (Turkey) of 238U, 232Th, 40K, and 137Cs were found to be 50, 42, 643 and 9.3 Naturally Occurring Radionuclides 457 85 Bq kg 1 in soil samples, and 39, 38, 573 and 6 Bq kg 1 in river sediment samples (Kurnaz et al. 2007). Radioactivity levels in soils, developed from various geological parent materials in the northern Highlands of Jordan, were 42.5, 49.9, 26.7 and 291.1 Bq kg 1 for 226Ra, 238U, 232Th, and 40K, respectively (Al-Hamarneh and Awadallah 2009). Radioactivity levels in sediment samples taken along the Upper Egypt Nile River for 238U, 232Th, and 40K ranged from 3.8 to 34.94, 2.8 to 30.10 and 112.31 to 312.98 Bq kg 1, respectively (El-Gamal et al. 2007). In soils from Bahawalpur (Pakistan) mean activity levels of 226Ra, 232Th, 40K, and 137Cs were 32.9, 53.6, 647.4 and 1.5 Bq kg 1 (Matiullah et al. 2004). In agricultural soils, application of fertilizers, particularly of inorganic P, has influenced radionuclide and trace element concentrations to a large extent (National Council on Radiation Protection and Measurements 1975). The application of phosphate fertilizers has substantially increased worldwide (Khater and Al-Sewaidan 2008). Fertilizers may also affect the chemical form of natural radionuclides in soils and thus their bioavailability (Klement 1982). 9.3.2 Natural Sources Modified by Humans The term NORM is used more specifically for all naturally occurring radioactive materials where human activities have increased the potential for exposure compared with the unaltered situation. The large production of NORM in the last decades and the potential long-term radiological hazards represent an increasing level of concern. NORM are found as products, by-products and/or wastes of industrial activities, such as production of non-nuclear fuels (e.g. coal, oil and gas), mining and milling of metalliferous and non-metalliferous ores (e.g. aluminum, iron, copper, gold and mineral sand), industrial minerals (e.g. phosphate and clays), radioisotope extraction and processing, as well as water treatments (IAEA 2003). 9.3.2.1 Fossil Fuels Oil and Gas Production Radioactivity in oil, gas and coal originates from cosmogenic and primordial radionuclides which are released to the environment through burning of these fuels (Ramachandran 2011). Rocks that hold oil and gas also contain U and Th at the order of some mg kg 1, corresponding to a total specific activity of some tens of Bq kg 1. However, these are commonly not mobilized from the rock formations. Oil and gas reservoirs contain a natural water layer (formation water) that lies under the hydrocarbons in that U and Th do not go into solution. However, the formation water tends to reach a specific activity of the same order of the rock matrix (Metz et al. 2003) due to dissolution of radium isotopes. In the water co-produced during 458 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.4 Radioactivity in oil and gas production Radionuclide 238 U 226 Ra 210 Po 210 Pb 222 Rn 232 Th 228 Ra 224 Ra Crude oil Bq g 1 <0.01 <0.04 <0.01 Natural gas Bq m 3 0.002–0.08 0.005–0.02 5–200,000 <0.002 Produced water Bq L 1 <0.1 0.002–1200 0.05–190 <0.01 0.3–180 0.05–40 Hard scale Bq g 1 0.001–0.5 0.1–15,000 0.02–1.5 0.02–75 Sludge Bq g 1 0.005–0.1 0.05–800 0.004–160 0.1–1300 0.001–0.002 0.05–2800 0.002–0.1 0.5–50 Compiled from Jonkers et al. (1997) oil and gas extraction, 226Ra, 224Ra, 228Ra and 210Pb are mobilized. These isotopes can precipitate out of solution, along with sulphate and carbonate deposits as scale or sludge in pipes. The immediate decay product of 226Ra is 222Rn which preferentially follows gas lines and is transformed through several rapid steps to 210Pb which can precipitate as a thin film in gas extraction equipment. The level radioactivity varies significantly, depending on the radioactivity of the reservoir rock and the salinity of the water co-produced from the well (Table 9.4). The radioactivity of scales and sludges can vary considerably and is much higher than that of formation water. Fracking for gas production releases significant radioactivity in drill cuttings and water. Exposure to radionuclides in the oil and gas industries poses a risk to workers during maintenance, waste transport, processing and decommissioning. In particular 210Pb deposits and films, as a beta emitter, is a concern only when pipe internals become exposed. Internal exposures can be reduced through hygiene practices. External exposure is generally low enough and does not require protective measures. Coal Production Coal contains uranium and thorium, their decay products and 40K. Total levels of individual radionuclides typically are not great and are generally about the same as in rocks near the coal. Enhanced radionuclide concentration in coal tends to be associated with the presence of other heavy metals and S. Characteristic values are given in Table 9.5. Coals from Australia, the US, may contain up to 4 ppm U, those from Germany up to 13 ppm U and coals from Brazil and China <20 ppm U. Concentrations of Th are often three times higher than those of U. The biggest producers of coal are China, the US and India, producing together more than two thirds of the world total. Coal has reached a global consumption of about 4.8 Pg in 2001 (IEA 2001). 9.3 Naturally Occurring Radionuclides 459 Table 9.5 Radioactivity in coal (in Bq kg 1; adapted from IAEA 2003) Country Australiaa Germanya Germanyb Greeceb Hungarya Polanda Romaniaa UKa USAa 238 U 8.5–47 117–390 20–480 <159 <415 7–19 6–73 226 Ra 19–24 10–145 <58 44–206 <557 8–22 9–59 210 Pb 20–33 210 Po 16–28 232 Th 11–69 10–63 <58 59–205 228 Ra 11–64 9–41 12–97 <123 <510 <580 12–78 3–52 <170 7–19 4–21 40 K 23–140 100–700 4–220 30–384 <785 55–314 a Hard coal Lignite b According to the World Coal Association 42% of the world’s electricity is currently generated by coal (WCA 2017). Most coal is used for electricity generation in coal-fired power plants. During combustion of coal, radionuclides are concentrated in fly ash and bottom ash. The concentrations can vary widely depending on the mineral composition of the coal, which can differ between mining regions. Differences also exist in residues from hard coal and lignite. The volatile 210Po shows a significant enrichment in fly ashes (Godoy et al. 2000). In older and smaller power plants, the absence or inadequacy of flue gas filtering and scrubbing may result in the atmospheric dispersal of fly ashes. In modern units, almost all of the fly ash is retained. While a lot of ash is deposited into surface impoundments and landfills, or is backfilled into the mines, some is also recycled together with gypsum for building construction. This may lead to exposure of the inhabitants to radiation. In 2003 the estimated worldwide fly ash production was 0.39 Pg (Mukherjee et al. 2008). In 2003 the US and EU fly ash production was about 0.06 Pg and 0.04 Pg, respectively. According to US GS (2002) the world coal consumption in 2035 is projected to reach about 9.4 Pg, corresponding to an estimated fly ash production of 0.75 Pg. In Table 9.6 some values for the radioactivity of coal ash and slag are given. A special situation arises when coal is burnt domestically. Emissions of particles from coal-based heating and cooking systems can be considerable, particularly in less developed countries (Clarke 1992). As coal mining produces a significant amount of waste rock and drainage water, the mining process itself also contributes to increased levels of radioactivity. Underground coal mines are subject to increased levels of Rn, while elevated levels of radium and 40K can be found in mining waste rocks and soil. Sediments discharged in waste water into the environment have been measured with activities as high as 55,000 Bq kg 1 of 226Ra and 15,000 Bq kg 1 of 228Ra (IAEA 2003). 460 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.6 Radioactivity in coal ash and coal slag from power plants (in Bq kg 1; adapted from IAEA 2003) Country Germanya Germanyb Egypta Egyptb Hungarya, b USAa 238 U series 6–166 68–245 16–41 41–90 200–2000 100–600 232 Th series 3–120 76–170 9–11 24–34 200–300 30–300 40 K 125–742 337–1240 300–800 100–1200 a Fly ash or bottom ash Slag b 9.3.2.2 Mining of Metal Ores Mineral Sands Mineral sands are sources of zircon, ilmenite, and rutile with xenomite and monazite which are mined in many countries. Mineral sand ore is important for the production of titanium, tin and zirconium bearing minerals and rare earth elements. The minerals have been concentrated by marine, alluvial and/or wind processes. These placer deposits can be found also in vein deposits, mostly disseminated in alkaline intrusions in hard rocks. The production amounts to millions of Mg per year of zirconium and titanium (from rutile and ilmenite). Thorium, tin and rare earth elements are associated. Depending on the placer geology, the radioactivity concentration of mineral sand can vary within a large range. The minerals in the sands are subject to gravity concentration and some of the concentrates are highly radioactive (Table 9.7). In terms of NORM, monazite is highly relevant. Monazite is a rare earth phosphate containing a variety of rare earth minerals such as cerium, lanthanum, xenotime and yttrium phosphate with traces of uranium and thorium (World Nuclear Association 2014). Industrial use of zirconium mainly occurs in the form of zircon (zirconium silicate). This mineral occurs naturally and requires little processing. It is used chiefly in foundries, refractories manufacture and the ceramics industry. Zircons typically have activities of up to 10,000 Bq kg 1 of 238 U and 232Th. No attempt is usually made to remove radionuclides from the zircon as this is not economical. Because zircon is used directly in the manufacture of refractory materials and glazes, the products will contain similar amounts of radioactivity. During mining and milling of zircon, care must be taken to keep dust levels down. When zircon is fused in refractories or ceramics manufacture, silica dust and fumes must be collected. This may contain the more volatile radionuclides, 210 Pb and 210Po, and the collection of these gases means that filters become contaminated. The main radiological issue is occupational exposure to these radionuclides in airborne dusts in the processing plant. Waste produced during zirconia/ zirconium production can be high in 226Ra, which presents a gamma hazard, and waste must be stored in metal containers in special repositories. Powders from filters used during zirconia manufacture have been assayed as high as 200,000 Bq kg 1 9.3 Naturally Occurring Radionuclides 461 Table 9.7 Radioactivity in mineral sand products 232 Mineral sand products Ore Mineral concentrate (primary separation) Ilmenite Rutile Zircon Xenomite Monazite Tailings waste 238 Th series μg g 1 5–70 80–800 Bq kg 40–600 600–6600 U series μg g 1 3–10 <10–70 Bq kg 1 70–250 <250–1700 50–500 <50–350 150–300 ~15,000 50,000–70,000 200–6000 400–4100 <400–2900 1200–2500 ~120 41,000–575,000 1500–50,000 <10–30 <10–20 150–300 ~4000 1000–3000 10–1000 <250–750 <250–500 3700–7400 ~100 25,000–57,000 12,000–60,000 1 Adapted from IAEA (2003) of 210Pb and 600,000 Bq kg 1 210Po (World Nuclear Association 2014). Tin is sometimes a by-product of mineral sand production. Slag from smelting tin often contains high levels of niobium and tantalum and so may form the feedstock for their extraction. It also typically contains enhanced level of radionuclides (World Nuclear Association 2014). Uranium Mining and Milling Uranium mining and extraction of U from the ore have generated the largest volume of radioactive waste worldwide. Globally, the total estimated volume of mill tailing is 938  106 m3 produced at roughly 4400 mines. The radioactivity of these tailings depends on the grade of ore mined and varies from less than 1000 Bq kg 1 to more than 100,000 Bq kg 1 (Abdelouas 2006). Kazakhstan has produced by far the largest volume of tailings (209  106 m3), while the production in the USA is about 120  106 m3. The radionuclides in uranium mill tailings include 238U, 235U, 234 U, 230Th, 226Ra, and 222Rn. The isotopes 238U and 230Th are long-lived emitters, while 222Rn has a short half-life (3.8 days). Radon concentrations persist in mill tailings because 222Rn is a decay product of the longer-lived 226Ra (half-live 1600 years), with the gamma radiation constituting the principal radiation risks from uranium tailings. In addition to radioactivity, uranium mill tailings are often associated with elevated concentrations of highly toxic heavy metals. Because of the oxidation of high sulfide content, uranium tailings generate acidic waters and accelerates the releases of radioactive and hazardous elements (Abdelouas 2006). Bauxite Mining for Alumina Production Bauxites are mined for alumina production. They generally contain concentrations of 232Th and 238U greater than the Earth’s crustal average. Activities for 232Th and 462 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.8 Radioactivity in red mud from alumina processing Country Greece Jamaica Hungary Turkey Germany Australia Activities (Bq kg 1) 232 Ra Th 13–185 15–412 370–1047 328–350 225–568 219–392 128–285 342–357 122  18 183  33 310  20 1350  40 226 References Papatheodorou et al. (2005)a Pinnock (1991)a Somlai et al. (2008)a Turhan et al. (2011)a Philipsborn and Kühnast (1992)b Cooper et al. (1995)b a Ranges indicating minima and maxima Meansstandard deviation b 238 U in bauxites were reported to be in a range of 400–600 Bq kg 1 for 238U and 300–400 Bq kg 1 for 232Th (UNSCEAR 2000). Parent rock composition significantly influences the 232Th and 238U abundances in bauxites. Those derived from acid igneous rocks show a concentration higher than those extracted from basic igneous rocks, whereas the bauxites mined from deposits of shales and carbonate rocks are characterized by intermediate concentrations. The process of lateritization during bauxite formation contributes to increase the Th: U ratio, which generally exceeds 4 (Adams and Richardson 1960). In 2009 the worldwide production of bauxite and alumina was 199  106 Mg and 123  106 Mg, respectively. Considering that the worldwide bauxite: alumina ratio is about 2.7, an amount of 1.7 kg of red mud is generated per kg of alumina (USGS 2002). During the process of extracting alumina from bauxite, over 70% of the radionuclides are concentrated in the red mud (Adams and Richardson 1960). Radionuclide activities in red mud reported in several studies are given in Table 9.8. Tantalum, Niobium and Tin production The main application of Ta is in the electronics industry as a major constituent of capacitors. Tantalum usually is associated with the chemically-similar niobium, often in tantalite and columbite. Tantalum ores, often derived from pegmatites, comprise a wide variety of more than a hundred minerals, some of which contain U and/or Th (Lehmann et al. 2013). Hence the mined ore and concentrate in some regions may contain both of these and their decay products in their crystal lattice. Concentration of the Ta minerals is generally by gravity methods (as with mineral sands), so the lattice-bound radioisotope impurities if present will report with the concentrate. While this has little radiological significance in the processing plant, concentrates shipped to customers sometimes exceed the Transport Code threshold of 10 kBq kg 1, requiring declaration and some special documentation, labeling and handling procedures. Some concentrates reach 75 kBq kg 1 (World Nuclear Association 2014). Niobium (Nb) slags can reach radioactivity levels in excess of 100 kBq kg 1. The largest producers of tantalum are Australia and Africa, most niobium comes from Brazil (World Nuclear Association 2014). In some situations 9.3 Naturally Occurring Radionuclides 463 Ta mineralisation is associated with tin-bearing minerals. The primary tantalum/tin ore contains trace quantities of 238U and 232Th associated with the minerals. Activities of 238U and 232Th in the ore are less than 60 Bq kg 1 and less than 5 Bq kg 1, respectively (Cooper 2005). After the initial dry and wet separation some 238U and 232Th may remain with the concentrates and are removed by acid leaching. Neutralisation of the leach solutions produces a solid tails containing 238U and 232Th. The levels of 238U and 232Th in the Ta products may reach radioactivity levels up to 75 kBq kg 1 (Cooper 2005). Copper Mining Copper is a metal of major importance with a range of uses, with the main application being for electrical installations and the electronics industry. Copper and gold are often associated with silver and uranium. After milling the ore, Cu is separated by flotation to produce a concentrate with a copper content of about 30%. The concentrates are smelted to remove volatile or less dense impurities. Further purification of the Cu melt from the smelter produces a primary form of the metal, known as blister copper. Electro-refining produces higher purity Cu. Most of the Au and U minerals are separated from the Cu concentrates during the flotation stage. Uranium and Th may be present in significant quantities in the Cu mineralization. Partitioning of 210Pb and 210Po from U occurs into the Cu concentrate during the smelting process. These radionuclides are vaporised at the smelting stage and may accumulate in dusts collected from off gases. Unless U is present in commercial quantities and separated during processing, the U will remain in the tailings from the flotation stage or will be present in the Cu concentrate and partition to the slag from the Cu smelter (Cooper 2005). 9.3.2.3 Phosphate Fertilizers Phosphate fertilisers for agricultural use are derived from phosphate rock. Superphosphate contains approximately 20% available P. Higher P contents are present in triple superphosphate, mono/diammonium phosphate, and dicalcium phosphate. The main phosphate-rock (phosphorite) deposits are both of igneous and sedimentary origin and are part of the apatite group. They are commonly encountered as fluorapatite and francolite, respectively. Uranium is incorporated in sedimentary phosphates through ionic substitution into the carbonate-flourapatitic crystals or by adsorption. Igneous phosphorite contains less U, but more Th. High contents of P usually correspond to high contents of U (50–300 mg kg 1; IAEA 2003). The world’s largest producers in 2007 were the USA, Morocco and China, the global total was 156 Mt P in the same year (World Nuclear Association 2014). Concentrations of radionuclides in phosphate rocks are given in Table 9.9. Normal superphosphate is produced by adding sulphuric acid to phosphate rock to form soluble mono-calcium phosphate. Phosphoric acid is used as the acidulating agent for higher grades of superphosphate and for ammonium phosphates. 464 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.9 Radioactivity in phosphate rocks Country USA Brazil Morocco South Africa Israel Jordan Australia 238 U (Bq kg 1) 259–3700 114–880 1500–1700 100–200 1500–1700 1300–1850 15–900 232 Th (Bq kg 1) 3.7–22 204–753 10–200 483–564 – – 5–47 226 Ra (Bq kg 1) 1.540 330–700 1500–1700 – – – 28–900 228 Ra (Bq kg 1) – 350–1550 – – – – – Adapted from IAEA (2003) Table 9.10 Radioactivity in phosphate fertilizers Product Phosphoric acid Superphosphate Triple superphosphate Ammonium phosphate Diammonium phosphate Dicalcium phosphate PK fertilizers NP fertilizers NPK fertilizers 238 U (Bq kg 1) 1200–1500 520–1100 800–2100 2000 2300 – 410 920 440–470 232 Th (Bq kg 1) – 15–44 44–48 63 <15 <37 <15 <30 <15 226 Ra (Bq kg 1) 300 110–960 230–800 20 210 740 370 310 210–270 Adapted from IAEA (2003) Phosphoric acid itself is produced by treating rock phosphate with excess sulphuric acid. The major solid waste product from the phosphate industry is the large quantities of calcium sulphate (phosphogypsum) arising during phosphoric acid production. Unless the acid is to be used for fertiliser production, purification of the phosphoric acid is carried out by solvent extraction (Cooper 2005). In phosphogypsum 80–90% of the 226Ra along with a high contents of 210Pb and 210 Po (Carvalho 1995; Beddow et al. 2006) are found. About 80–85% of the 238U (Beddow et al. 2006) and roughly 70% of 232Th (Tayibi et al. 2009) concentrate in phosphoric acid. The radioactivity of P fertilizers is variable and depends on the radionuclide content of the phosphate rock and on the method of production. Radioactivity levels of different P products are given in Table 9.10. 9.3.2.4 Ceramics and Building Materials Building materials may contain radionuclides due to their occurrence in raw materials (e.g. 238U and 232Th natural stones) or may be added with industrial products such as zircon sand or with industrial byproducts like coal ash, phosphogypsum and furnace slags. Generally, the radioactivity in the endproduct 9.3 Naturally Occurring Radionuclides 465 Table 9.11 Radioactivity in different building materials Material Concrete Clay bricks (red) Sand-lime bricks/limestone Natural building stones Natural gypsum Cement Tiles Phosphogypsum plasterboard Blast furnace slagstone 226 Ra (Bq kg 1) 1–250 9–2200 6–50 1–500 <70 7–180 30–200 4–700 30–120 232 Th (Bq kg 1) 1–190 <220 1–30 1–310 <100 7–240 20–200 1–53 30–220 40 K (Bq kg 1) 5–1570 180–1600 5–700 1–4000 7–280 24–850 160–1410 25–120 – Adapted from IAEA (2003) will be lower compared to the original by-product because of the presence of other inert material in the building material. Typical levels of radioactivity in different materials are given in Table 9.11. 9.3.3 Anthropogenic Radionuclides 9.3.3.1 Release of Radionuclides by Nuclear Weapons Testing The era of nuclear weapon testing began with the Trinity test of 1945. A total of 543 atmospheric weapon tests were carried during the period 1945–1980 giving a total yield of 440 Mt (comprising of 189 Mt due to fission and 251 Mt due to fusion) by all nuclear weapon countries all over the world. These were carried out at different locations on and above the earth’s surface. Depending on the location of the detonation the radioactive debris entered the local, regional, or global environment. For tests conducted on the earth’s surface, a portion of the radioactive debris is deposited at the site of the test (local fallout) and regionally up to several thousand kilometers down wind (intermediate fallout). Most radionuclides released in nuclear tests have a short half life. The most important fission products are given in Table 9.12. Additionally, radionuclides such as 236U, 237Np, 238-242Pu and 241Am are formed by neutron capture reactions. In terms of radioactivity, 3H, 90Sr, 137Cs and plutonium isotopes are currently the radionuclides of great importance. 9.3.3.2 Release of Radionuclides by Reactors and Reprocessing Plants Generation of electrical energy by nuclear means has grown steadily from the start of the industry in 1956. Currently, 439 reactors in 31 different countries are in operation (IAEA 2008). Nuclear power currently provides 16% of electricity in 466 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.12 Radionuclides released to the atmosphere due to nuclear testing Radionuclide H 14 a C 54 Mn 55 Fe 89 Sr 90 Sr 91 Y 95 Zr 103 Ru 106 Ru 3 Estimated release (EBq) 186 0.21 3.98 1.53 117 0.62 120 148 247 12.2 Half-life (days or years) 12.3 years 5730 years 312.3 years 2.7 years 50.5 days 28.8 years 58.5 days 64.0 days 39.3 days 1.0 years Radionuclide 125 Sb 131 I 137 Cs 140 Ba 141 Ce 144 Ce 239 Pu 240 Pu Pu Estimated release (EBq) 0.74 675 0.95 759 263 30.7 0.0065 0.0044 0.14 Half-life (days or years) 2.7 years 8.0 days 30.0 years 12.8 days 32.5 days 284.9 days 24,110 years 6560 years 14.4 years Adapted from UNSCEAR (2000) a All 14C was assumed to be due to fusion Russia, 18% in the UK, 30% in Japan, 32% in Germany, 48% in Sweden and Ukraine, 78% in France, 19% in the USA and 30% for the entire European Union (Hu et al. 2010). During the operation of a typical reactor, more than 200 radionuclides are produced most of which are relatively short-lived and decay to low levels within a few decades (Crowley 1997). A number of radionuclides are emitted from normal operation of NPP. For example, the annual discharge of gaseous 14C to the atmosphere from pressured water reactors in Germany was 280  20 GBq GWe 1 in 1999, on average 30% is thought to have emitted in the form of CO2, the rest in other forms. In France, 14C discharges were estimated to be 140 GBq year 1 per unit of 900 MWe and 220 GBq year 1 per unit of 1300 MWe (Roussel-Debet et al. 2006). Based on combined worldwide operable nuclear reactors of 3.72  105 MWe (World Nuclear Association 2007), the global annual discharge of 14C may amount to 60 TBq year 1. For comparison, cosmogenic natural production in the upper atmosphere is at a rate of approximately 1.54  103 TBq year 1 and all atmospheric nuclear tests emitted about 2.13  105 TBq of 14C. Monitoring of radionuclides in terrestrial and aquatic environments, including soil, plant and foodstuff samples, has been performed to assess the potential environmental contamination from normal operation of nuclear power plants (Hu et al. 2010). 9.3.3.3 Nuclear Waste Countries using nuclear power must deal with high-level radioactive waste. High level waste and spent nuclear fuel are stored at some 125 sites in 39 states, with over 161 million people residing within 121 km of temporarily stored nuclear waste (Hu et al. 2010). Spent nuclear fuel remains highly radioactive for thousands of years. Isolating this high-level waste from people and the environment has been an important and challenging issue for all countries that use nuclear power. Although 9.3 Naturally Occurring Radionuclides 467 high-level waste makes up only about 3% of the world’s total volume of radioactive waste, it contains roughly 95% of all the radioactivity in all kinds of radioactive wastes. Countries with high-level waste and spent nuclear fuel plan to dispose of these materials in a permanent repository which is commonly a deep geologic disposal facility. The waste is destined for vitrification in a borosilicate glass before permanent disposal. For example, in France and Germany, more than 500 Mg of waste have been vitrified (Hu et al. 2010). However, up to now, not a single country has disposed of high level waste. Along with the USA, Belgium, Canada, China, Finland, France, Germany, Japan, Russia, Spain, Sweden, Switzerland and the United Kingdom have invested significant resources in their radioactive waste management programs because of their reliance on nuclear energy (Hu et al. 2010). Deep geologic disposal has been expected to be the best method for isolating highly radioactive, long-lived waste (Witherspoon and Bodvarsson 2001). 9.3.3.4 Releases and Environmental Impacts from the Chernobyl and Fukushima Accidents Radioactive contamination of the environment has occurred as a result of nuclear accidents. The most notable accident is the one that destroyed Unit 4 of the Chernobyl nuclear complex in the Ukraine in April 1986. The accident at the Fukushima Daiichi nuclear power plant, following the earthquake and tsunami in March 2011, was one of the biggest nuclear disasters in recent years. In both accidents, most of the radioactivity released was due to volatile radionuclides, such as noble gases (85Kr, 133Xe), iodine (129I, 131I, 133I), cesium (134Cs, 136Cs, 137 CS) and tellurium (129Te, 132Te) (reviewed by Steinhauser et al. 2014). However, the amount of refractory elements (including actinides) emitted in the course of the Chernobyl accident was approximately four orders of magnitude higher than during the Fukushima accident. While for Chernobyl, a total release of 5300 PBq (excluding noble gases) has been estimated, for Fukushima, a total source term of 520 (340–800) PBq was established (Steinhauser et al. 2014). In the course of the Fukushima accident, more than 80% of the radioactivity may have been deposited in the Pacific Ocean. Initially, Chernobyl’s “exclusion zone” covered a radius of 30 km, corresponding to 2800 km2 around the nuclear complex. Fallout of hot particles in the vicinity of the reactor caused a considerable contamination of the soil surface, with 137Cs up to 106 Bq m 2 (Hu et al. 2010). About 116,000 people within this zone were evacuated to less contaminated areas in the months following the accident. The evacuation started only 3–11 days after the accident, which was already critical for parts of the population (Pr€ ohl et al. 2002). The exclusion zone was later extended and comprised 4300 km2 in 1996 (EC/IAEA/WHO 1996). The Chernobyl catastrophe resulted in significant contamination also outside the evacuation zone, as well as in many European states, including Belarus, Russia, Sweden, Finland, Norway, Germany, Slovenia, Croatia, Austria, Greece, and others (Fig. 9.2). 468 9 Health Risks Associated with Radionuclides in Soil Materials Fig. 9.2 Surface contamination with 137Cs in Europe after the Chernobyl nuclear accident (Steinhauser et al. 2014, p. 804; with kind permission from Elsevier) The contribution of 137Cs from the Chernobyl plume was significant even at 2000 km distance. However, the total surface contamination was at least two orders of magnitude below the level within the 30 km exclusion zone. Overall, there were 28 deaths from the acute radiation syndrome as a result of the Chernobyl accident (Eisenbud and Gesell 1997). The exclusion zone around the Fukushima reactor encompassed an area of less than 600 km2 (Yoshida and Takahashi 2012). Radionuclides were mainly deposited northwest of the reactor, causing the greatest contamination in a strip about 40 km in length (Hirose 2012). On 12 March 2011, a 20 km radius around the Fukushima reactor was designated as the “stay-away 9.3 Naturally Occurring Radionuclides 469 evacuation” zone. On 15 March 2011, the time of major releases of radionuclides, the surrounding area between 20 and 30 km was declared an “indoor evacuation” zone. On 16 March 2011, evacuees less than 40 years of age were instructed to leave the stay-away evacuation zone and to take pills or syrup of stable iodine (Hamada et al. 2012). Both the Chernobyl and Fukushima accidents caused radionuclide contamination of the atmosphere, hydrosphere, biosphere and pedosphere over the whole northern hemisphere. Not only the distance but also local weather conditions cause large variability in radionuclide concentrations in environmental media. After the Chernobyl catastrophe, maximum radioactivity in air resulting from the release of 131I and 137Cs was equivalent to 750,000 and 120,000 Bq m 3 at the reactor, 200 Bq m 3 and 9.9 Bq m 3 at 400 km distance, and < 12 Bq m 3 and < 10 Bq m 3 at distances greater than 1100 km from the reactor (Steinhauser et al. 2014). After the Fukushima accident, maximum values for 131I were 5600 Bq m 3 at the reactor (data for 137Cs are not available). Values for 131I and 137Cs at 190 km distance were 32 and 0.016 Bq m 3, and for distances greater than 1100 km, they did not exceed 0.026 and 0.0032 Bq m 3, respectively. Particularly the rainwater concentrations after the Chernobyl accident reflected the great influence of meteorological conditions on radionuclide concentrations. For example, in G€oteborg, Sweden (1300 km distance to reactor) maximum values for 131I and 137Cs corresponded to 3000 and 950 Bq L 1 while in Munich, Germany with a similar distance to Chernobyl (1400 km), the corresponding values were 58,000 and 6500 Bq L 1, respectively. The latter values were the highest that were determined for rainwater after the Chernobyl accident. Highest values after the Fukushima accident were determined at 200 km distance from the reactor (for 131I: 6072 Bq L 1 and for 137Cs: 752 Bq L 1). Maximum radionuclide concentrations in soils following the Chernobyl and Fukushima accidents are listed in Table 9.13. Part of the long-lived radionuclides (137Cs, 90Sr and 239+240Pu) can be present from previous accidents or nuclear weapon tests and may exhibit a significant background contamination that is difficult to distinguish from recent deposition. In summary, the impact of the Chernobyl accident was greater than of the Fukushima accident. The highly contaminated areas as well as the evacuated areas were smaller around Fukushima and the projected health effects in Japan were estimated to be lower than after the Chernobyl disaster. This is mainly due to the fact that evacuations and food safety campaigns worked quickly after the Fukushima accident. 9.3.3.5 Releases Through Medical Uses of Radiation In medicine, radiation is used for both diagnostic and therapeutic purposes. As the benefits of procedures become more widely disseminated, medical uses of radiation increase from year to year (UNSCEAR 2000). The physicians, technicians, nurses and others involved constitute the largest group occupationally exposed to anthropogenic sources of radiation. Occupational doses received by staff varies according to the source of exposure, thereby characterizing different occupational groups. 470 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.13 Activities of selected radionuclides (in Bq kg 1) in soil after the Chernobyl and Fukushima accidents Location Chernobyl Pripyat/Ukraine Christogalovka/ Ukraine Christinovka/ Ukraine Kupetsch/Ukraine V€ocklabruck, Austria Athens, Greece Mumbai, India Fukushima Fukushima Daiichi Fukushima Daiichi Sendai/Japan Kashiwa/Japan Mekong Delta/ Vietnam Vienna/Austria Distance to power plant (km) 131 137 90 239+240 4 5 n.d. n.d. 1,239,000 74,000 420,000 36,000 n.d. n.d. 64 n.d. 15,000 130 17.8 100 1250 n.d. n.d. 3460 506 44 n.d. 1.1 n.d. 1600 5100 n.d. 1.5 22,000 9.5 n.d. n.d. n.d. n.d. 0.9 4.3 95 195 4700 49,000 10,000 n.d. n.d. n.d. 1700 232 ~3 35 n.d. ~2200 ~0.5 n.d. ~0.5 n.d. 9000 ~0.4 n.d. n.d. I Cs Sr 1,790,000 2,740,000 5000 421,000 35 21.4 Pu Adapted from Steinhauser et al. (2014) n.d. not detected or not reported 9.4 Behaviour of Important Radionuclides in Soil-Water Systems Radionuclides of potential concern include long-lived fission products having relatively high to moderate waste inventories and high to moderate environmental mobilities (99Tc, 129I, 79Se, 126Sn), transuranic radionuclides of relatively high waste inventories and ingestion dose conversion factors (239+240Pu, 241Am, 242Cm, 237 Np), long-lived neutron-activation products contained in irradiated reactor waste materials, such as spent fuel disassembly hardware, spent control rods, reactor internals, and spent primary coolant demineralizer resin (10Be, 14C, 59+63Ni, 94Nb, 108 Ag) and naturally-occurring actinide radionuclides having relatively high waste inventories (235+238U, 232Th). The short-lived tritium (3H) is considered a suitable water tracer. The isotopes 90Sr and 137Cs are the major fission products, yet they do not pose long-term risk because of their relatively short half-lives and strong sorption in the subsurface. In general, the mobility of actinides in aqueous systems is low. Of particular importance to the environment and risk assessment are radionuclides such as 99Tc, 129I, and 237Np, because of their long half-lives (2.13  105, 1.57  107 and 2.14  107 years for 99Tc, 129I and 237Np, respectively) 9.4 Behaviour of Important Radionuclides in Soil-Water Systems 471 and presumably high mobility (Hu and Smith 2004; Arnold et al. 2006). Activities conducted at European nuclear reprocessing facilities (mainly Sellafield and La Hague) have led to the increased radioactivity in the Arctic marine environment due to discharge of 99Tc (Hu et al. 2010). The most important processes governing radionuclide interactions with soils include adsorption/desorption (including ion exchange) and precipitation/dissolution. Adsorption occurs primarily in response to electrostatic attraction. The degree of adsorption of radionuclides is governed by the pH of the solution because the magnitude and polarity of the net surface charge of a solid changes with pH. Mineral surfaces become increasingly more negatively charged as pH increases. The sorption of radionuclides in soils is frequently quantified by the partitioning coefficient (Kd). The Kd is a factor related to the partitioning of a radionuclide between solid and aqueous phases and is defined as the ratio of the quantity of the adsorbate adsorbed per mass of solid to the amount of the adsorbate remaining in solution at equilibrium. Radionuclides that adsorb very strongly to soil have large Kd values (typically greater than 100 mL g 1). Radionuclides that do not adsorb to soil and migrate essentially at the same rate as the water flow have Kd values near 0. The key geochemical processes affecting the mobility and bioavailability of selected radionuclides in soils are briefly described in the following paragraphs. 9.4.1 Nickel-59,63 (63,59Ni) In aqueous systems, the most important oxidation state of Ni is +2 (Baes and Mesmer 1976). At pH values less than 10, in aqueous systems the uncomplexed cation Ni2+ is the dominant Ni species. At pH values greater than 10, dissolved Ni is present in form of hydroxides. Nickel also forms aqueous complexes with ligands, such as dissolved chloride, sulfate and carbonate (Rai et al. 1984). The Ni concentrations in most soils are controlled by surface sorption processes. Nickel is known to be adsorbed by iron and manganese oxides and clays (Rai et al. 1984). Nickel is moderately to highly sorbed by soils with Kd values ranging from several tens to thousands milliliters per gram (Thibault et al. 1990). The adsorption of Ni2+ cations is greatest at pH values less than 11, decreases with deceasing pH, and minimal under acidic conditions. At pH values greater than 11, Ni adsorption to soil may decrease if the dominant Ni aqueous species is anionic, such as Ni(OH)3 . The sorption of Ni to clay minerals results in the formation of nickel hydroxide or nickel aluminum hydroxide surface precipitates (Scheckel and Sparks 2001). The formation of such surface precipitates strongly reduces Ni migration and remobilization in soil-water systems. 472 9.4.2 9 Health Risks Associated with Radionuclides in Soil Materials Selenium-79 (79Se) Selenium exists in the 2, 0, +4, and +6 oxidation states (Baes and Mesmer 1976). Dissolved Se is commonly present in the +6 oxidation state under oxidizing conditions as the dominant species HSeO4 and SeO42 at pH values less than and greater than 2, respectively. Under moderately oxidizing to reducing conditions, the Se(IV) species H2SeO3º (aq), HSeO3 , and HSeO32 may be dominant at pH values less than approximately 2.5, from 2.5 to 7, and greater than 7, respectively. The Se( 2) species H2Seº (aq) and HSe are the dominant aqueous Se species at pH values less than and greater than about 4, respectively, under highly reducing conditions. Dissolved Se in the 2, +4, and +6 oxidation states is present as anionic species at pH values greater than 4 under all redox conditions within the thermodynamic stability range of water. Due to the relatively high vapor pressure of these compounds, methylated forms of Se can be significant contributors to the mobility of Se. In some soil systems under moderately and highly reducing conditions, the concentration of dissolved Se may be controlled by the precipitation of solid forms of Se, such as Seº. Elemental Se is relatively insoluble in soils over a wide range of pH under moderately reducing conditions. In highly reducing and organic-rich systems containing dissolved sulfide or bisulfide, Se-sulfide solids and metal selenides, such as ferroselite (FeSe2), are insoluble and would limit the concentration of dissolved Se and its mobility. Selenium can be reduced to its lower oxidation states by abiotic and biotic processes, which may result in the bioreduction of Se to insoluble forms, such as solid elemental Se (Seº) or S1-xSe. This will limit Se mobility and bioavailability in soils. Because the dominant aqueous species of Se(IV) and Se(VI) are anionic over the pH range of most soils, the adsorption of Se to mineral surfaces would be expected to be low in most soils. However, hydrous oxides of Fe and Al and amorphous alumino-silicates have a high sorptive affinity for Se(VI) and Se(IV) (Rai et al. 1984). Values of Kd typically range from a few milliliters to several tens of milliliters per gram. The adsorption of Se also depends on pH with adsorption being strong under acidic conditions and decreasing with increasing pH. This pH dependency is consistent with that of other radionuclides and inorganic contaminants present primarily as anion. 9.4.3 Strontium-90 (90Sr) In aqueous solutions, dissolved Sr over the whole pH range up to pH 11 is present predominantly as uncomplexed Sr2+. At pH values greater than 11, the neutral carbonate complex SrCO3º (aq) is predicted to be the dominant aqueous complex (Robertson et al. 2003). In alkaline soils, the precipitation of strontianite (SrCO3) or coprecipitation in calcite may be important mechanisms for controlling the concentrations of dissolved Sr (Lefevre et al. 1993). Strontium as an alkaline-earth element can 9.4 Behaviour of Important Radionuclides in Soil-Water Systems 473 form similar solid phases as those with calcium. In certain soil systems, celestite (SrSO4) and strontianite are potentially two important solubility controls for strontium, but most strontium minerals are highly soluble. Celestite may precipitate in acidic soil environments at elevated concentrations of total dissolved Sr and sulfate. In contrast, strontianite is only stable in highly alkaline soils. However, Sr does not commonly precipitate as a pure mineral, because the total concentrations of dissolved Sr2+ in most environmental systems are less than the solubility limits of Sr-containing minerals and much lower than the concentrations of dissolved Ca2+. Because the ionic radii for Sr2+ (1.12 Å) and Ca2+ (0.99 Å) are similar, Sr2+ can substitute for Ca2+ in minerals to coprecipitate as a Sr-containing calcite (Ca1-xSrxCO3) (Veizer 1983). In most soils, Sr adsorption is controlled primarily by cation exchange. The important factors that affect adsorption and Kd values for Sr include CEC, pH, and concentrations of calcium and stable strontium naturally present in soil. Strontium Kd values increase with increasing CEC and pH. The correlation between strontium Kd values and pH is likely the result of H+ ions competing with Sr2+ for exchange sites. Strontium Kd values of less than 1 mL g 1 to more than 30,000 mL g 1 have been reported (Sheppard and Thibault 1991). The adsorption of Sr2+ has also been found to decrease with increasing concentrations of competing cations, such as Ca2+. Because Ca2+ concentrations in environmental systems are usually several orders of magnitude greater than stable strontium concentrations and many orders of magnitude greater than 90Sr concentrations, the significantly greater mass of calcium increases the possibility that Ca2+ will outcompete Sr2+, especially 90Sr. 9.4.4 Technetium-99 (99Tc) Technetium exists in oxidation states from +7 to 1. Depending on the redox conditions, Tc exists in two stable oxidation states. It forms a reduced species [predominantly Tc(IV)] at Eh values below about 220 mV under neutral pH conditions. At higher Eh, it occurs as Tc(VII)O4 . Technetium(VII) can be reduced to Tc(IV) by abiotic and biotic processes. This reduction results in a decrease in the dissolved concentrations of Tc due to the precipitation of the weakly soluble, amorphous TcO2
2 H2O. In reduced iron-sulfide systems, Tc(VII) can be reduced to Tc(IV) by coprecipitation with FeS solid (mackinawite) (Wharton et al. 2000). Dissolved Tc is present under aerobic conditions as the aqueous Tc(VII) oxyanion species TcO4 over the complete pH range of natural waters. The TcO4 anion is essentially nonadsorptive. Due to its weak interaction with mineral surfaces, TcO4 is one of the most mobile radionuclides in the environment. In contrast, transport of Tc(IV) species (TcO2  nH2O) are expected to be strongly retarded because of sorption and/or precipitation (Eriksen et al. 1992). Lieser and Bauscher (1987) by varying the redox potential observed a change in the Kd value of about three orders of magnitude over a small range of Eh at 170  60 mV and a pH of 7  0.5. In carbonate-containing waters, Tc(IV) carbonate complexes, such as TcCO3(OH)2º 474 9 Health Risks Associated with Radionuclides in Soil Materials (aq) and TcCO3(OH)3 , may become important aqueous complexes of Tc (Eriksen et al. 1992). Studies on the sorption of Tc on sediments, soils, pure minerals, oxide phases, and crushed rock materials consist primarily of measurements of Kd values for Tc (VII). The adsorption of Tc(VII) oxyanion TcO4 is very low to zero, i.e., Kd values of  0 mL g 1, at near neutral and basic pH conditions and to increase when pH values decrease to less than 5. However, Kd values for Tc(VII) sorbed on sediments high in organic matter can be considerable (Thibault et al. 1990). The sorption of TcO4 is positively correlated to the soil organic matter content (Wildung et al. 2000). Adsorption of Tc(VII) in experiments conducted with organic material as well as with crushed rock and Fe(II)-containing minerals has been attributed to the reduction of Tc(VII) to Tc(IV). Technetium(IV) is considered to be immobile because it readily precipitates as hydrous oxides and forms complexes with surface sites on Fe and Al oxides and clay minerals. 9.4.5 Tin-126 (126Sn) Tin exists in compounds in several oxidation states from 4 to +4 but only the +2 and +4 states are important for Sn in natural systems (Baes and Mesmer 1976). However, Sn2+ and its hydrolysis products are limited to very reducing conditions at or below the stability boundary for the breakdown of water. According to Séby et al. (2001) information regarding the hydrolysis of Sn(IV) and speciation in general is limited due to the low solubility of SnO2 (the mineral cassiterite) and its precipitation in laboratory studies of Sn(IV) speciation. Ashby and Craig (1988) found that Sn methylation may occur by abiotic and biotic process and thus may contribute to the mobility of Sn in the environment. As Sn(IV) will be present in the +4 oxidation state in most environmental systems, Sn is likely to be adsorbed in soils due to the low solubility of cassiterite, presupposed that no anionic aqueous complexes are formed. 9.4.6 Iodine-129 (129I) Similar to 99Tc, 129I has a complex chemistry in the environment which is caused by the multiple redox states. However, in contrast to 99Tc, iodine has a minimal retardation under reducing conditions when I is the predominant form. The 1, +5, and molecular I2 oxidation states are those most relevant for iodine in environmental systems. Different iodine species (I ,IO3 , and organic iodine species) are known to coexist in various aqueous systems (Hu et al. 2005). The iodide anion (I ) is highly mobile. Under more oxidizing conditions, iodine may be present as the iodate anion (IO3 ), which is more reactive than iodide. Adsorption of iodine species in soils seems to be controlled mainly by soil organic matter and Fe and Al oxides which becomes increasingly important under 9.4 Behaviour of Important Radionuclides in Soil-Water Systems 475 more acidic conditions. Results from numerous Kd studies suggest that the adsorption of iodide increases with increasing soil organic matter content, but the majority of the reported Kd values for iodide are limited to soils containing less than 0.2% organic matter contents (reviewed by Robertson et al. 2003). Values of Kd for iodide have been reported in the range from 1 to 10 mL g 1 for the pH range from 4 to 10, but most of the reported Kd values are typically less than 3 mL g 1. Because iodine is present as either the anions I or IO3 in most soils, their adsorption on soils and most minerals should be negligible at near neutral and alkaline pH conditions. Iodine volatilization from soils to the atmosphere may be a result of both chemical and microbiological processes. The chemical processes generally result in molecular iodine or hydrogen iodide, and the microbiological processes yield methyl iodide which is not strongly retained by soil components and is only weakly soluble in water (Whitehead 1984). 9.4.7 Cesium-137 (137Cs) Cesium in the environment exists in the +1 oxidation state. Compared to the other contaminants the speciation of Cs in environmental systems is relatively simple. Cesium does not form any important aqueous complexes with ligands and organic matter found in natural systems and Cs containing solids are highly soluble in aqueous systems. Therefore, the precipitation and coprecipitation of Cs-containing solids are not important processes in controlling the concentration of dissolved Cs in environmental systems. The behavior of Cs in soils is similar to K+. Cesium due to strong sorption on most minerals has large Kd values (Robertson et al. 2003). The sorption of Cs occurs primarily by ion exchange in most soil systems except when mica-like minerals are present. For example, when mica-like clay minerals such as illite and vermiculite are present, selective fixation of Cs between structural layers of these minerals occurs. The extent to which Cs is fixed will depend on the concentration of mica-like clays in the soil, and the concentration of major cations, such as K+, as K+ can effectively compete with Cs+ for ion exchange sites because its hydrated ionic radii are similar and smaller than those for the other alkali and alkaline earth ions. Cesium may also adsorb to iron oxides by complexation of cesium to surface mineral sites whose abundance is pH dependent (Schwertmann and Taylor 1989). In contrast, the sorption of Cs to humic substances is generally weak (Bovard et al. 1968). 9.4.8 Thorium-232 (232Th) In natural soil-water environments Th exists only in the +4 oxidation state. In the absence of dissolved ligands other than hydroxide, the uncomplexed ion Th4+ is the 476 9 Health Risks Associated with Radionuclides in Soil Materials main aqueous species at pH values less than about 3.5. At pH values exceeding 3.5, the hydrolysis of Th is dominated by the aqueous species Th(OH)22+ and Th(OH)4º (aq) with increasing pH (Robertson et al. 2003). At pH values greater than 5 the neutral hydroxide complex Th(OH)4º (aq) dominates the aqueous speciation of dissolved Th. Dissolved Th can form strong aqueous complexes with inorganic ligands, like dissolved carbonate, fluoride, phosphate, chloride, nitrate, and organic ligands. The complex ThF22+ may be dominant at pH values less than 5. Phosphate complexes may be important at acidic and near-neutral pH conditions. In addition, Th carbonate complexes Th(OH)3CO3 and Th(CO3)56 may be of some importance for Th mobility. Concentrations of Th in soils may also be controlled by adsorption processes. Oxides of Fe and Mn are expected to be important adsorbents of Th. Humic substances are considered particularly important in the adsorption of Th (Gascoyne 1982). Thibault et al. (1990) reviewed published Kd data for Th as a function of soil type and reported Kd values for Th that range from 207 to 1.3  107 mL g 1. 9.4.9 Uranium-235 (235U) Uranium in aqueous environments exists in different oxidation states (+3, +4, +5, +6). Uranium(VI), occurring as UO22+ (uranyl), and U(IV) in natural environments are the most common oxidation states of U. Under oxidizing to mildly reducing environments, U exists in the +6 oxidation state. Under reducing conditions, U (IV) is considered relatively immobile because U(IV) forms hardly soluble minerals, like uraninite (UO2). Because the oxidation state of U has a significant effect on its mobility in waste streams and the natural environment, the reduction of U (VI) to U(IV) by abiotic and biotic processes has recently received considerable attention. These processes may be useful for certain remediation technologies, such as permeable barriers composed of zero-valent iron particles, or sodium-dithionitereduced soils. Microbial reduction of U(VI) has also been suggested as a potential mechanism for removal of U from contaminated waters and soils (Lovley 1995). In carbonate-containing waters the aqueous speciation of U(VI) at near neutral and basic higher pH values is dominated by anionic aqueous carbonate complexes (e.g., UO2CO3º (aq), UO2(CO3)22 , and UO2(CO3)34 ). The formation of anionic U (VI) carbonate complexes at pH values greater than 6 result in an increase in U (VI) solubility, decreased U(VI) adsorption, and thus increased U mobility because anions do not readily adsorb to mineral surfaces at basic pH conditions. Uranium can also form stable complexes with other naturally occurring inorganic and organic ligands. Complexes of UO22+ with phosphate (UO2HPO4º (aq) and UO2PO4 ) may be important in aqueous systems with a pH between 6 and 9 when the total concentration ratio PO4(total)/CO3(total) is greater than 0.1 (Sandino and Bruno 1992). Complexes with sulfate, fluoride, and possibly chloride may be important uranyl species where concentrations of these anions are high. However, their stability is considerably less than the carbonate and phosphate 9.4 Behaviour of Important Radionuclides in Soil-Water Systems 477 complexes (Grenthe et al. 1992). Organic complexes may also be important to U aqueous chemistry, thereby increasing their solubility and mobility. The uncomplexed uranyl ion has a greater tendency to form complexes with fulvic and humic acids than many other metals with a +2 valence (Kim 1986) which was attributed to the greater “effective charge” of the uranyl ion compared to other divalent metals. In soils, U(VI) adsorbs onto a variety of minerals, including clays (ChisholmBrause et al. 1994), oxides and silicates (Waite et al. 1994), and organic material (Read et al. 1993). Environmental parameters controlling U adsorption include pH, redox conditions, and concentrations of complexing ligands. The adsorption of uranium to humic substances may occur through ion exchange and complexation processes that result in the formation of stable U(VI) complexes involving the acidic functional groups (Borovec et al. 1979). 9.4.10 Neptunium-237 (237Np) Neptunium may exist in the +3, +4, +5, +6, and +7 valence states, but only the +4, +5, and possibly +6 states are relevant to natural environments. Neptunium in aqueous systems has a large stability range for Np(V) (Lieser and Mühlenweg 1988). The pentavalent NpO2+ species is dominant at pH values <8 whereas Np (V) carbonate complexes tend to dominate at higher pH values (Kaszuba and Runde 1999). Since Np(V) solid phases are relatively soluble and Np(V) aqueous species do not easily sorb onto common minerals, Np(V) is relatively mobile in the environment (Nakata et al. 2002). Under reducing conditions, Np(IV) is present as the low solubility Np(OH)4 (aq) species at pH values >5 (Kaszuba and Runde 1999). Neptunium(V) species in soils to some extent adsorb to Fe oxides and clay minerals, but not on most common minerals (Nakata et al. 2002). Because NpO2+ does not compete favorably with dissolved Ca2+ and other divalent ions for adsorption sites on soils, the Kd values for Np(V) are relatively low (Kaplan and Serne 2000). Especially for iron oxides, the adsorption of Np(V) has a strong dependence on pH, (Kohler et al. 1999). Typically, the sorption of Np(V) on minerals is negligible at pH values less than pH 5, and increases at pH values between 5 to 7. This pH-dependency is expected for ions that are present in solution primarily as cations, such as NpO2+. 9.4.11 Plutonium-239 + 240 (239+240Pu) Under most environmental conditions Pu can occur in the +3, +4, +5, and +6 oxidation states. Under oxidizing conditions, Pu(IV), Pu(V), and Pu(VI) may exist, while Pu(III) and Pu(IV) would be present under reducing conditions (Allard 478 9 Health Risks Associated with Radionuclides in Soil Materials and Rydberg 1983). In aqueous systems, Pu(III) species, such as Pu3+, would be dominant up to pH values of approximately 8.5 under reducing conditions. The Pu (IV) species Pu(OH)4º (aq) is predicted to have a large stability range extending above near neutral pH values at moderately oxidizing conditions to pH values greater than 8 under reducing conditions (Robertson et al. 2003). Plutonium may form stronger complexes with dissolved carbonate, sulfate, phosphate, and fluoride, relative to those with ligands such as chloride and nitrate. Dissolved Pu may also form complexes with dissolved organic matter, such as fulvic and humic material. Associated with organic matter, Pu is mainly present in the +4 oxidation state (Nelson et al. 1987). Dissolved Pu in the environment is commonly present at less than 10–15 mol L 1 which indicates that adsorption may be the most important process affecting the retardation of plutonium in soils. However, Kd values for Pu can extremely vary, depending on the properties of the substrate, pH, and the composition of solution. According to Thibault et al. (1990) they may range from 27 to 190,000 mL g 1. If no complexing ligands are present, the adsorption of Pu commonly increases with increasing pH from about pH 5 to 9. At pH values exceeding 7, concentrations of dissolved carbonate and hydroxide will decrease Pu adsorption of and increase its mobility in soils as a result of the formation of strong mixed ligand complexes with Pu. According to Sanchez et al. (1985) increasing carbonate concentrations decreased the adsorption of Pu(IV) and Pu(V) on the surface of goethite [α-FeO (OH)]. At low pH in the presence of high concentrations of dissolved organic carbon, Pu-organic complexes may control Pu adsorption and mobility in soils (Robertson et al. 2003). 9.4.12 Americum-241 (241Am) Americium exists in several oxidation states (+3, +4, +5, and +6) but Am(III) is the most stable and important oxidation state in environmental systems. All of the higher oxidation states are strong oxidizing agents and stable only in systems containing no oxidizable substances (Ames and Rai 1978). Americium is present as Am(III) in all of the dominant species predicted to be stable for the Eh-pH region of environmental interest. The uncomplexed ion Am3+ is the dominant aqueous species at moderately to highly acidic conditions. At near neutral to alkaline pH conditions, Am(III) carbonate and hydroxyl complexes will dominate the aqueous species of Am(III). Sorption studies indicate that Am(III) readily sorbs to soils, pure minerals, and crushed rock materials and exhibits high Kd values that are often in the range of 1000 to greater than 100,000 mL g 1. Americium(III) is, therefore, considered one of the most immobile actinide elements in the environment. The adsorption of Am(III) increases with increasing pH with peak adsorption occurring between pH 5 and 6. This pH dependence is because of the dominant aqueous species of Am(III) in the pH range of natural waters which are primarily Am3+ complexes at acidic and cationic carbonate and hydroxyl complexes at basic pH 9.5 Routes of Exposure 479 values. Americium(III) is more mobile at low to moderate pH values where the net surface charge on minerals becomes more positive. The tendency of Am(III) to strongly sorb to soil particles suggests that there is a potential for colloid-facilitated transport of Am(III). Colloids of clay and humic acids are potentially important substances for the transport of actinides in soil/water systems (Robertson et al. 2003). 9.4.13 Curium-242 (242Cm) Curium can exist in the +3 and +4 oxidation states. However, the +3 state is the dominant oxidation state in natural waters. Curium(IV) is not stable in solutions because of self-radiation reactions (Onishi et al. 1981). In natural waters, Cm(III) is predicted to be the dominant species at pH values less than 7 and may form complexes with inorganic ligands (Wimmer et al. 1992). The complexes CmCO3+ and Cm(OH)2+ dominate the aqueous speciation of Cm(III) in the pH range from 7 to 9 and at pH values greater than 9, respectively. Other aqueous complexes of Cm(III) include those with sulfate, fluoride, chloride, and humic substances (reviewed by Robertson et al. 2003). Sorption studies indicate that Cm(III) readily sorbs to minerals, crushed rock, and soil materials. Compared to other actinides, Cm(III)is considered to be immobile in soil environments and exhibits high Kd values. Adsorption of Cm(III) is strongly pH dependent and increases with increasing pH with peak adsorption occurring between pH values of 5 and 6 (Robertson et al. 2003). Similar to the environmental behaviour of Am(III), the tendency of Cm(III) to strongly adsorb to soil particles suggests that there is a potential for colloid-facilitated transport of Cm(III). The mobility of Cm(III) in soils may thus be enhanced by its migration in colloid form (Kaplan et al. 1994). 9.5 Routes of Exposure There are two major routes of radionuclide exposure (Fig. 9.3). The first exposure route involving external contamination, originates from natural and anthropogenic sources of ionizing radiation. Part of the natural radiation is cosmic radiation from space and the other part is due to radioactive materials in soil and in building materials. Higher levels of natural radioactive material are left in products or on the land due to human activities. Skin can be contaminated by contact with aerosols or radionuclide-contaminated surfaces. For aerosols that deposit on the skin, factors such as particle size may be important in that larger particles may deliver larger doses to the skin, especially for beta-emitting radionuclides. Contamination of skin with beta-emitting radionuclides can also cause serious burns. 480 9 Health Risks Associated with Radionuclides in Soil Materials Fig. 9.3 Major exposure pathways of radionuclides The second major exposure route is by internal contamination, which occurs when the radioactive source is incorporated into the human body through inhalation, ingestion or through uptake via skin or wounds. As with other routes of exposure, the nature and physiochemical properties of the radionuclide are of major importance when determining the effect of the internalized isotope. Inhalation is the primary route of exposure for internalized radionuclides. Their fate depends on the size of the inhaled substances and the solubility of the radionuclide. Approximately 25% of inhaled radionuclides are immediately exhaled. Of the remaining 75%, particles greater than 10 μm tend to remain in upper areas of the lung while those less than 5 μm in diameter can reach the alveolar space (Morrow et al. 1967). In the lung, radionuclides such as tritium, phosphorus, and cesium are rapidly solubilized and enter the circulatory system. Less soluble radionuclides, such as the oxides of plutonium, uranium, cobalt, and americium, may be removed through phagocytosis by alveolar macrophages. Until being removed, the radioactive particle will continue to irradiate surrounding tissues. In most cases, the internalized radionuclide will have both soluble and insoluble components (Harrison and Muirhead 2003). Larger particles which are unable to access the alveolar space, may be removed from the lung via mucocilliary clearance and thus enter the gastrointestinal tract. In case of ingestion of radioactive material, the majority of radionuclides are poorly absorbed by the intestinal tract. Some exceptions include strontium, tritium, and cesium. The amount of damage inflicted will be determined by the transit time through the alimentary canal, with the greatest potential for damage occurring in the descending colon prior to the ingested radionuclide being excreted in the feces. Transit times are affected by a variety of factors, including diet, fluid intake levels, and physical activity, but generally range from 1 to 5 days (Eve 1966). The International Commission on Radiological Protection (ICRP 1999) has presented recommendations for the protection of the public in situations of prolonged 9.5 Routes of Exposure 481 exposure to radiation, including justification for remediation of contaminated sites. The ICRP dose limit for exposure of a member of the public from all relevant practices is 1 mSv year 1. Examples of dose rates from sources of major concern for the public are given in the following sections. 9.5.1 Exposure from Cosmogenic Radiation External exposure from cosmogenic radionuclides, except for 3H, 7Be, 14C and 22 Na, contribute only to a small extent to radiation doses. UNSCEAR (2000) for these radionuclides gave estimates of effective dose rates (μSv year 1) of 12 from 14C, 0.03 from 7Be, 0.15 from 22Na and 0.01 from 3H. Cosmic ray doses increase by a factor of about two for every 1500 m increase above the mean sea level (AMSL), resulting in 260 to 270 μSv year 1 at 0–150 m AMSL, 390 to 520 μSv year 1 at 1220 to 1830 m AMSL and 1070 μSv year 1 at 3400 m AMSL (Ramachandran 2011). However, contribution of this source compared to other sources (particularly terrestrial radiation) for most part of the global population is small. 9.5.2 Exposure from Natural Terrestrial Radiation Outdoor values of exposure from terrestrial radiation for different countries range between 18 and 93 nGy h 1 (0.16–0.82 mGy year 1) (Thorne 2003). The population weighted average may be around 59 nGy h 1 (0.52 mGy year 1). Indoor rates are commonly larger than outdoor rates. UNSCEAR (2000) estimated a global average effective dose rate of 0.48 mSv year 1 (0.41 mSv year 1 indoors and 0.07 mSv year 1 outdoors). Results for individual countries commonly range between 0.3 and 0.6 mSv year 1. External dose rates in regions of high natural background can be significantly increased. These include areas with monazite sand deposits, which have high levels of Th, regions with Andosols (volcanic soils) and precipitates of Ra associated with hot springs. For example, over themonazite sands in Kerala and Madras, India, absorbed dose rates in air of 200–4000 nGy h 1 (1.8–35 mGy year 1) have been detected. On monazite sand beaches of Guarapari (Brazil) dose rates of up to 90,000 nGy h 1 (0.79 Gy year 1) have been measured (UNSCEAR 2000). 9.5.3 Exposure from Nuclear Weapons Tests UNSCEAR (2000) has published global means of effective dose rates from fallout radionuclides. Average effective dose rates for external radiation peaked in 1962 and 482 9 Health Risks Associated with Radionuclides in Soil Materials 1963 at around 38 μSv year 1, with 95Zr and 95Nb contributing about 50% of the total. In 1999, the global average was 2.90 μSv year 1, almost entirely from 137Cs. Global average effective dose rates from ingested radionuclides peaked at 35 μSv year 1 in 1962, particularly from 131I (20.4 μSv year 1), 137Cs (10.3 μSv year 1) and 90Sr (3.1 μSv year 1). Until 1999, the effective dose rate was reduced to 0.90 μSv year 1, particularly from 137Cs (0.35 μSv year 1) and 90Sr (0.56 μSv year 1). The effective dose rate from 3H and 14C, with 12.7 μSv year 1 reached its maximum in 1962, with 7.2 μSv year 1 from 3H and 5.5 μSv year 1 from 14C. Until 1999, the total effective dose rate decreased to 1.7 μSv a 1, almost completely from 14 C. For inhalation, global means of effective dose rates with 36 μSv year 1 reached their maxima in 1963, particularly from144Ce (15.3 μSv year 1), 106Ru (9.3 μSv year 1), 90Sr (1.9 μSv year 1) as well as from radioisotopes of Pu and Am (7.7 μSv year 1). In 1985, inhalation doses made only a small contribution to exposure from fallout (Thorne 2003). The mean total effective dose rate from fallout in the Northern and the Southern Hemisphere in 1963 reached 125 μSv and 59 μSv year 1, respectively. The global mean is thus dominated by the Northern Hemisphere with a larger population compared to the Southern Hemisphere. The global averages for the Northern and Southern Hemisphere and both hemispheres in 1999 were 5.87, 2.68 and 5.51 μSv year 1, respectively (Thorne 2003). 9.5.4 Exposure from the Nuclear Fuel Cycle The global average effective dose from the nuclear fuel cycle may generate a per caput effective dose of 0.1 μSv to the world population. This would represent only 0.005% of the average exposures to natural sources of radiation. The average percentage contribution may be 37.2 from reactor fabrication, 34.4% from mining, 14.3% from mile and mill tailings, 7.2% from fuel processing each 2.9% from milling and transportation, 1.15% from airborne reactor release and 0.07% from fuel fabrication. The contribution from mining, milling and mill tailings is thus almost equal to that from reactor and fuel processing. 9.5.5 Exposure from Nuclear Accidents Radiation exposure from the soil and floral cover due to nuclear accidents will depend on a number of factors, particularly radionuclide composition of the fallout, the time of exposure after the fallout, the precipitation conditions at the time of deposition and afterwards, and upon the nature of the aerial biomass at the time of deposition. These factors probably account for the different external radiation dose rates per Becquerel per square meter implied in some of the post-Chernobyl reports. For example, in the Federal Republic of Germany “top level” ground deposits of 100,000 Bq m 2 representing an “absorbed dose” equivalent to 0.06 mSv hr. 1 9.5 Routes of Exposure 483 were indicated for persons “permanently on such ground”. Respective data from Sweden, in contrast, implied that in the “regions of highest contamination” local levels of the order of 1 MBq m 2 (i.e., for 131I and 137Cs) were reached and involved external irradiation dose rates of the order of 10 mSv hr 1, i.e., approximately 1 mSv hr 1 for a deposit of 100,000 Bq m 2 (IAEA 1986). 9.5.6 Contaminated Food and Water Concentrations of natural radioactivity in food are commonly in the range of 40–600 Bq kg 1 food. The radioactivity from potassium alone may be 50 Bq kg 1 in milk, 420 Bq kg 1 in milk powder, 165 Bq kg 1 in potatoes, and 125 Bq kg 1 in beef (Shahbazi-Gahrouei et al. 2013). According to Ramachandran and Mishra (1989) the concentration of 40K radioactivity in different foods varied from 45.9 to 649.0 Bq kg 1, that of 226Ra varied from 0.01 to 1.16 Bq kg 1, and that of 228Th from 0.02 to 1.26 Bq kg 1. Food, water, and air also contain trace amounts of alpha emitters from the uranium, thorium, and actinium series. Some of the radon (222Rn, and to a lesser extent 220Rn and 219Rn) gas diffuses into the food supply. For example, radon and its decay products, are deposited on the soil and the vegetation. To derive the corresponding dose in mSv year 1, it is necessary to take into account the energy and fraction deposited in the body, besides taking into account not only the radioactive lifetime but also the biological lifetime of the isotopes in the human body. The natural radioactivity from the 40K isotope, which is a constant fraction (0.0117%) of the potassium content in food, varies significantly with potassium concentration in different food sources. In Europe, following the accident of the Chernobyl atomic power station in 1986, high radioactivity levels of some wild-growing mushroom species were observed (Kalac 2001). Until 1985, activities of 137Cs, from nuclear weapons testing, were commonly below 1 kBq kg 1 dry matter. After the Chernobyl accident, activities up to tens of kBq kg 1 dry matter of 137Cs and to a lesser extent of 134Cs were observed in the following years in edible mushrooms. The species most heavily contaminated included Xerocomus badius, Xerocomus chrysenteron, Suillus variegatus, Rozites caperata and Hydnum repandum (Kalac 2001). Wildgrowing mushroom consumption contributed up to 0.2 mSv to the effective dose in individuals consuming about 10 kg (fresh weight) of heavily contaminated species per year. The radioactivity of cultivated mushrooms is negligible. Wild animals, consuming contaminated mushrooms, have elevated levels of radionuclides in their tissues. 9.5.7 Total Radiation Exposure Excluding extreme scenarios such as atomic power plant accidents and releases of radionuclides through nuclear weapons tests, it can be summarized that on a global 484 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.14 Radiation exposure and contribution from natural, modified and anthropogenic sources to the Indian population Source All natural Modified natural Nuclear weapons tests Nuclear fuel cycle Medical exposure Total Dose (mSv year 1) 2.299 0.00124 0.045 0.00005 0.048 2.393 Contribution (%) 96.07 0.052 1.88 0.0021 2.01 100.0 Adapted from Ramachandran (2011) scale, the natural sources, particularly soil materials, account for major share of total radiation exposure. An example is given in Table 9.14 for the Indian population. However, Ramachandran (2011) reported that this pattern is similar to the global situation. In future, these proportions may change, depending on the technological and industrial development, the energy policy and the trend in medical uses of radiation. 9.6 Clinical Effects of Radiation Ionizing radiation is a form of energy that may be destructive in biological systems and can cause serious diseases such as cancer and mutations and prodromal symptoms such as anorexia, nausea, vomiting, diarrhea, nervousness, confusion and consciousness in humans (Sharma et al. 2010). The types of radiation include alpha particles, beta particles, gamma particles, neutrons and X-rays (see Sect. 9.1). 9.6.1 Alpha Radiation Because of their large mass and charge, α-particles strongly ionize biological tissues. If the alpha particle is from radioactive material that is outside the human body, it will lose all its energy before passing through the dead outer layer of the human skin. This means that exposure to α-radiation only exists if α-particles are taken up by ingestion or inhalation. The α-particles can cause damage to tissues in the body. 9.6.2 Beta Radiation Most β-particles (except for tritium) have enough energy to pass through the dead outer layer of a person’s skin and irradiate the live tissue underneath. Human 9.8 Biological Significance of Radiation 485 exposure is also possible through ingestion or inhalation of β-particles that lose their energy by exciting and ionizing atoms along their path. When all of the kinetic energy is spent, negative β-particles (negatrons) become ordinary electrons and have no more effect on the body. A positive β-particle (positron) collides with a nearby electron, and this electron-positron pair turns into a pair of gamma rays called annihilation radiation, which can interact with other molecules in the body. 9.6.3 Gamma Radiation Gamma particles can pass through many kinds of materials, including human tissues. The gamma ray source can be relatively far away, like the radioactive materials in nearby construction materials, soil, and asphalt. A γ-ray may pass through the body without hitting anything, or it may hit an atom and give that atom all or part of its energy. This normally knocks an electron out of the atom (and ionizes the atom). This electron then uses the energy it received from the γ-ray to ionize other atoms by knocking electrons out of them as well. Since a γ-ray is pure energy, once it loses all its energy it no longer exists. Very dense materials, such as lead, are commonly used as shielding to slow or stop γ-photons. 9.7 Isotopes of Concern for Human Health There are numerous natural and manufactured radioisotopes that could result in internal contamination. Of these, about 40 radionuclides are potentially hazardous to humans (Casarett 1968). Only a limited number including plutonium, cesium, strontium, radioiodine, and tritium provide most significant health hazards to humans (Prasad 1995). Isotopes of serious human health concern and their target organs are listed in Table 9.15. 9.8 Biological Significance of Radiation Biological significance is a result of a combination of high decay energy, bioavailability and energy transfer to biological systems. The biological significance of radiation results from the enormous amount of energy contained in each emission. Visible light has an energy range of 1.77–4.13 electron volts. Most chemical changes occur within a range of 5–7 electron volts. Biologically significant radiation levels range from 18,610 electron volts for the weak beta emitting tritium (3H) (half-life: 12.35 years) over 511,630 electron volts for 137Cs (half-life: 30.174 years) to 5,155,400 electron volts for plutonium (239Pu) (half-life: 24,131 years). These highly energetic emissions carry enough energy to tear electrons from neutral atoms and molecules. External exposure to 239Pu poses 486 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.15 Radiation exposure and contribution from natural, modified and anthropogenic sources to the Indian population Element Americum Cesium Cobalt Iodine Iridium Phosphorus Plutonium Polonium Radium Strontium Tritium Uranium Isotope of concern 241 Am 137 Cs 60 Co 131 I 192 Ir 32 P 238 Pu, 239Pu 210 Po 226 Ra 90 Sr 3 H 235 U, 238U Type of radiation α, γ β, γ β, γ β, γ β, γ β α, (γ) α α, β, γ β, γ β α, β, γ Major target organs Lung, liver, bone Whole body, particularly kidney Whole body, particularly liver Thyroid Spleen Bone Lung, liver, bone Kidney, spleen Bone Bone Whole body Kidney, bone Adapted from Kalinich (2012) very little health risk, since plutonium isotopes emit α-radiation, and almost no β- or γ-radiation. In contrast, internal exposure to 239Pu is an extremely serious health hazard. It generally stays in the body for decades, exposing organs and tissues to radiation, and increasing the risk of cancer. Plutonium is also a toxic metal, and may cause damage to the kidneys. The radionuclide 137Cs, a β-emitter with a γ-component, is biologically significant due to its high energy level, its long halflife, its ubiquitous production during the fission process, and its tendency to follow the potassium cycle in nature, giving a whole body dose to those who ingest it. The weaker β-radiation of 3H is slightly more penetrating than α-radiation. Its biological significance comes from its ubiquitous production during the fission process, its tendency to follow the water cycle in nature, thus also contaminating soil materials, and its ability to become tissue bound in humans and the biotic environment. 9.9 9.9.1 Therapy Determination of Radioactive Contamination Contaminated persons could suffer acute symptoms of radiations injury and in the longer term could develop genetic damage or cancer. Early detection of the contaminants helps in early response for decontamination and decrease the potency of hazards. Information from the attack scene will provide the first information on the radioisotopes involved, but in many cases, the isotope and route of exposure are unknown. However, several simple assessments can determine whether the contaminant is a β or γ emitter, and may also indicate possible routes of exposure. A preliminary step is a body survey using a Geiger-Müller meter that is capable of detecting β and γ emitting isotopes, but not those emitting γ particles (Kalinich 9.9 Therapy 487 2012). The first scan should be made with the shield of the Geiger-Müeller meter open to detect the presence and location of β and γ contamination. The next scan is to be conducted with the closed shield. Results from this scan will indicate what proportion of the contamination is due to γ emitting isotopes alone (Kalinich 2012). Radioactive contamination by α emitters should also be determined. A variety of instruments is available to measure α radiation. However, special training in use of these instruments is necessary for making accurate measurements. 9.9.2 External Decontamination 9.9.2.1 Skin Decontamination External decontamination is essential to prevent the spread of radionuclides. Decontamination procedures decrease external radiation levels, allowing for a more accurate determination of internal radionuclide contamination. The primary objective is to remove the highest possible degree the radioactive contaminant from external body parts. The process requires removing the contaminants not just from the skin surface, but also from the clothing and protective equipments (Owens and Peacock 2004). The decontamination of skin involves different methods. The most commonly used one is simple washing with warm water and soap using face cloths and brushes. Apart from soap, the decontaminating agents include washing pastes, shampoos and various decontamination solutions (Turner 2007). To decontaminate radioactive material from eye or nose they should be washed with 0.9% saline solution, or if not available with tap water. Care should be taken that the water does not get swallowed and the solution for decontamination is not irritant to skin and other organs (Sharma et al. 2010). 9.9.2.2 Wound Decontamination Washing and cleaning of the wound in most cases is sufficient (Kalinich 2012). Adding a chelating agent to the washing solution in some cases supports contaminant removal. For wound contamination with plutonium, americium, or curium, chelation therapy with Ca-DTPA and Zn-DTPA is recommended. In case of wounds containing embedded radioactive fragments it may be necessary to surgically remove the fragment. The latter should be placed in shielded containers. 9.9.3 Internal Decontamination Radioactive contaminants in the body may cause significant health risks. The risks are long term and depend on type, nature and concentration of the radioactive material. Agents used for radionuclide decontamination can be categorized into 488 9 Health Risks Associated with Radionuclides in Soil Materials Table 9.16 Internal decontamination options Target radionuclide 241 Am 137 Cs Co Compound Calcium and zinc salts of DTPAa Prussian blue Calcium and zinc salts of DTPAa 60 131 I P 238 Pu, 239Pu Potassium iodide Phosphate, dibasic potassium and Na salts Calcium and zinc salts of DTPAa 210 2,3-Dimercaptopropanol 226 Ammonium chloride, Ca carbonate, Ca gluconate, Na alginate Al hydroxide, Al phosphate, ammonium chloride, Ca carbonate, Ca gluconate, Na alginate Sodium bicarbonate 32 Po Ra 90 Sr 235 U, 238U Application Intravenous infusion Oral Intravenous infusion Oral Oral Intravenous infusion Intramuscular injection Oral Oral Intravenous infusion or oral Adapted from Kalinich (2012) DTPA Diethylenetriamine pentacetic acid a absorption-reducing agents, blocking and diluting agents, mobilizing agents, and chelating agents (Kalinich 2012). However, a single compound does not work for all contaminants, demonstrating the need for identification of the radioactive source. An overview of internal decontamination options is given in Table 9.16. 9.9.3.1 Absorption-Reducing Agents Absorption can be reduced by lavage of stomach, emetics, purgatives, laxatives and ion exchangers. Stomach lavage is effective only if it is performed immediately after contamination and ingested dose is high. Commonly used emetic agents are apomorphine (5–10 mg, subcutaneous) and ipecac (1–2 g in capsule or 15 ml in syrup), which should be given concomitantly with 200–300 ml of water (Kalinich 2012). If the person is unconscious then emetics are contradicted and purgatives should not be used in individuals with abdominal pain. Certain non-absorbable binding resins are capable of preventing the uptake of various radioactive materials in the gut. For example, Prussian blue, which is a non-absorbable pigmented resin has been used orally to enhance the faecal excretion of cesium and thallium by means of ion exchange process. Several antacids (e.g. Al containing antacids, Al-hydroxide) are shown to reduce the absorption of radioactive strontium if given immediately after the exposure (Kalinich 2012). 9.9 Therapy 9.9.3.2 489 Blocking and Diluting Agents These agents are used to reduce the uptake of radionuclides from the blood tissues into target tissue (Torngren and Persson 1998). For example, potassium iodide is recommended to prevent radioactive iodine from being sequestered in the thyroid. Elements with chemical properties similar to the internalized radionuclide are also used as blocking agents. For example, calcium, and to a lesser extent phosphorus, can be applied to block the uptake of radioactive strontium. Diluting agents simply dilute the concentration of radionuclides in the body and decrease their absorption (Kalinich 2012). Water, for example, can be used to increase the excretion of tritium. 9.9.3.3 Mobilizing Agents Mobilizing agents enhance the release of deposited radionuclides from the tissue by increasing their natural turnover process rate. Examples of mobilizing agents include diuretics, propylthiouracil, expectorants, parathyroid extract and corticosteroids, ammonium chloride and sodium bicarbonate (Dubois et al. 1994). Ammonium chloride results in acidification of the blood and increased elimination of internalized radiostrontium. Sodium bicarbonate is used to increase urinary pH which is useful for preventing the precipitation of uranium that passes through the kidney (Kalinich 2012). 9.9.3.4 Chelating or Complexing Agents Chelating or complexing agents such as ethylenediamine tetraacetic acid (EDTA), diethylenetriamine pentacetate (DTPA), nitrilotriaceticacid (NTA) form complexes with certain radionuclides producing water soluble compounds which can be more rapidly eliminated from the body via excretion by the kidneys (Dalvi et al. 1980). Calcium and zinc salts of DTPA are approved as decorporating agent. DTPA forms water soluble, stable complexes with transuranium elements and increase their elimination from the body. Both Ca-DTPA and Zn-DTPA are safe for the treatment of plutonium, americium, or curium (Sharma et al. 2010). Ca-DTPA is administered as a single intravenous injection or inhaled immediately possible after contamination, and repeated doses of Zn-DTPA administered intravenously which may be given daily as maintenance therapy, as necessary. Uranium contamination has been treated with oral sodium bicarbonate, regulated to maintain an alkaline urine pH, and accompanied by diuretics (Sharma et al. 2010). 490 9.10 9 Health Risks Associated with Radionuclides in Soil Materials Measures for Remediation of Radionuclide Contaminated Sites 9.10.1 Classification of Radionuclide-Contaminated Sites The size of radionuclide-contaminated land may stretch over different ranges, from small- and mid-size areas with diameters from several tens to hundreds meters to large-size areas with diameters of thousands of meters. For smaller areas, mechanically or physio-chemically based technologies may be convenient. Larger areas such as the landscapes contaminated by the Chernobyl or Fukushima Daiichi accidents for several reasons may be mainly suited for phytoremediation technologies. The classification of contaminated areas relates to the phenomenological description of the site and the processes most important for the fate of the contaminants, such as the basic chemistry, generation, internal transport, outflow, adsorption, etc. These processes are in turn strongly influenced by the characteristics of a given site, including the geochemical constituents. In terms of remediation, the major radioactive contaminants can be grouped in two classes, including (i) radionuclides originating either from original ore or from nuclear waste, such as 238Pu, 239Pu, 240Pu, 241Pu, 241Am, 237Np, 234U, 235U, 238U, 230Th, 226Ra, 228Ra, 222 Rn, 210Pb, 231Pa, 227Ac, and (ii) radioisotopes of other elements, originating from low- or intermediate level radioactive waste, like 14C, 36Cl, 63Ni, 90Sr, 93Zr, 94 Nb, 99Tc, 107Pd, 126Sn, 129I, 135Cs and 137Cs. This information is required for adequate site restoration. 9.10.2 Mechanically or Physio-chemically Based Technologies The application of remediation techniques requires consideration of performance, reliability and maintenance requirements, cost, available supporting infrastructure, risk to workers and public during implementation, environmental impact, and regulatory and community acceptance. The choice of remediation technology for radioactively contaminated land also needs to consider future land use. Remediation technologies can include complete or partial remove, stabilization, immobilization (e.g. cement-based solidification, chemical immobilization) or isolation of the contamination (US EPA 1996; Mallett 2004). The most commonly used remediation techniques include (i) excavation of contaminated soil for on-site storage or offsite disposal, (ii) covering with inert materials for reducing or avoiding external exposure, creation of dust or infiltration by rainwater (and hence transfer of radioactivity to groundwater), and (iii) vertical and horizontal in-ground barriers to prevent contamination migration (IAEA 2002). The suitability of restoration techniques for treating different radionuclides depends upon their chemical and physical properties as well as on site-specific 9.10 Measures for Remediation of Radionuclide Contaminated Sites 491 Table 9.17 Effectiveness of restoration techniques in treating radionuclide-contaminated sites Restoration technique Excavation Soil washing Flotation Filtration Chemical solubilisation Desorption Biosorption Surface barriers Sub-surface barriers Solidification Vitrification Chemical immobilization Radionuclide Co Sr x x x x x x x x x Cs x x x x x x x x x x x x x x x x Ra x x x x x x Th x x x x x x x x x x x x x U x x x x x x x x x x x x Pu x x x x x x x x x x x Compiled from US EPA (1996) factors. Excavation of the contaminant, containment (surface and sub-surface barriers) and immobilization (solidification and chemical immobilization) are likely to be suited for most radionuclides (Table 9.17). However, while excavation removes contaminated soil from the site it creates large quantities of contaminated wastes. Excavated wastes necessitate long-term monitoring which limits the amounts of materials requiring disposal at landfill sites. Disposal of excavated materials is a problem worldwide. Potential alternative technologies for remediation of radionuclide contaminated soils include soil washing, desorption, electrochemical remediation and vitrification. Although very energy consuming, the latter method has been shown to be very efficient in immobilizing radionuclides in the long term. For example, on a test site in Australia, using an array of four electrodes, 3–6 Mg of soil per hour were melted to depths of 7–15 m. Retention of Am, Sr and Pu in the vitrified residue approached 100% while retention of Cs was >90% (IAEA 1999). The above methods are also used for remediation of sites contaminated with toxic elements. They have been described in detail in Chap. 8. In the latter, the costs of different remediation technologies are indicated as well. However, for remediation of radionuclide-contaminated soils, additional high costs arise for the specific disposal of radioactive materials. These have been indicated with 2000–3000 EUR m 3 for disposal of radioactive soil, and radioactive waste from soil washing, flotation, filtration, desorption, chemical solubilisation and biosorption (Zeevaert and Bousher 1999). Indirect costs arising from monitoring and loss or gain of income and taxes, are site-specific and cannot be calculated from the characteristics of the remediation technologies. Another important aspect in mechanically based site remediation is soil compaction caused by heavy equipment that is necessary to remove or transport soil. Compaction deteriorates physical conditions of the soil that have major impacts on 492 9 Health Risks Associated with Radionuclides in Soil Materials chemical and microbiological processes which ultimately affect plant growth. Treating the soil with dispersing and chelating chemicals removes not only radionuclides but also soil nutrients which are necessary for microbial and plant growth. Dispersing chemicals often adversely affect not only soil chemical processes, but also soil physical processes. After the soil is replaced, establishment of plants on a physically, chemically and microbiologically compromised soil is a great challenge (Entry et al. 1996). 9.10.3 Phytoremediation Radionuclides released into the environment are taken up by plants and redistributed throughout the ecosystem. A serious problem may arise when agroecosystems become contaminated. Soils contaminated with radionuclides provide a particular challenge to soil decontamination and hence a useful perspective on phytoremediation. The term phytoremediation has already been defined in Chap. 8, as this technology is being used for toxic element-contaminated sites as well (McGrath et al. 2002). Phytoextraction, phytostabilization, and phytovolatilization have been suggested as suitable remediation techniques for radionuclides (Beresford 2006). Besides the fact that phytoremediation is a non-intrusive technique, estimates from the USA suggest that this method is considerably more cost-effective than other remediation techniques (Fiore et al. 2000). However, major limiting factors for phytoextraction are low radionuclide availability in soils and limited translocation from roots to surface plant parts. For effective application of this technique the radionuclide uptake by plant roots needs to result in translocation to shoots. The major part of 137Cs taken up by plants accumulates in roots (Clint and Dighton 1992). Phytoremediation studies have been performed with a wide range of radionuclides, including 3H, 90Sr, 95Nb, 99 Tc, 106Ru, 144Ce, 226,228Ra, 239,240Pu, 241Am, 228,230,232Th, 244Cm and 237Np (Dushenkov 2003). 9.10.3.1 Phytoextraction A wide array of plant species occupying different habitats have been shown to accumulate large amounts of radionuclides from contaminated soils. Some are given in Table 9.18. For effective remediation it is necessary to grow plants producing high amounts of biomass that should exceed 3 Mg dry matter ha 1 in a cropping season and accumulate >1000 mg radionuclide kg 1. A variety of factors that may help to bring radionuclides into the soil solution and make it more available for plant uptake. These include manipulation of soil pH, addition of chelators, amending soil with chemicals stimulating radionuclide desorption, interaction with microorganisms and plant exudates. Bioavailability of radionuclides is strongly influenced by pH. For example, uranyl cations (UO22+) that dominate at 9.10 Measures for Remediation of Radionuclide Contaminated Sites 493 Table 9.18 Plants with potential for phytoextraction of various radionuclides Plant species Annual plants Sinapis ssp. Amaranthus retroflexus Radionuclide References 137 Cs Cs Dushenkov et al. (1999) Fuhrmann et al. (2002) Huang et al. (1998) Huang et al. (1998) Lasat et al. (1998) Negri and Hinchman (2000) Negri and Hinchman (2000) Cs Cs Dahlman et al. (1969) Salt et al. (1992) 137 Brassica ssp. U Phaseolus acutifolius Phalaris arundinaceae Grasses Festuca ssp. Lolium perenne Trees Acer rubrum Cocos nucifera Pinus radiata Eucalyptus tereticornis 137 137 137 137 244 Cs Cs, 90Sr, U Cm, 137Cs, 238Pu, 226Ra and 90Sr Cs 137 Cs, 90Sr 137 Cs, 90Sr 137 Pinder et al. (1984) Robison and Stone (1992) Entry et al. (1993) Entry et al. (1996) pH 5 were more readily taken up and translocated in plants compared to hydroxyl (pH 6) and carbonate (pH 8) complexes of U. Chelators (see Chap. 8) have shown a tenfold increase in 137Cs availability and an almost hundredfold increase in Pb and U availability (Huang et al. 1998). In greenhouse experiments, shoots U concentration in plants grown in a U-contaminated soil increased to more than 5000 mg kg 1 in citric acid treated soil compared to 5 mg kg 1 in control pots (Huang et al. 1998). In hydroponic experiments it was shown that shoots to roots ratio of 137Cs was significantly higher in mycorrhizal plants of Calluna vulgaris L. (heather) compared to non-mycorrhizal plants (Clint and Dighton 1992). In the Chernobyl exclusion zone mustard was used to noticeably reduce 137Cs activity (Dushenkov et al. 1999). Using three subsequent mustard crops, 137Cs activity was reduced from 2558 to 2239 (averages of the exclusion zone) Bq kg 1 soil. Within only one growth period, the portion of the area with 137Cs levels exceeding 3000 Bq kg 1 soil could be reduced from 29.4% to 7.7%. At a contaminated field site within the Brookhaven National Laboratory (New York, USA) about 50% of 137Cs and 90Sr was estimated to be removed by twice-yearly cropping of Amaranthus retroflexus within 15 and 7 years, respectively (Fuhrmann et al. 2002). Trees have also been shown to take up substantial amounts of radionuclides. Acer rubrum, Liriodendron tulipifera and Liquidambar stryaciflua accumulated high quantities of a large spectrum of radionuclides (Pinder et al. 1984). Cocos nucifera took up substantial amounts of 137Cs from soils contaminated by nuclear weapons testing on Bikini Atoll (Robison and Stone 1992). Entry et al. (1993) documented that Pinus radiata seedlings accumulated more 137Cs and 90Sr than Pinus ponderosa. Entry and Emmingham (1995) found that potted Eucalyptus 494 9 Health Risks Associated with Radionuclides in Soil Materials tereticornis seedlings removed 31.0% of the 137Cs and 11.3% of the 90Sr in sphagnum peat soil after one month of exposure. To increase plant uptake of radionuclides or toxic elements, genetic modification of plants has also been suggested (Entry et al. 1997; McGrath et al. 2002). However, Wolfe and Bjornstad (2002) warn that the development of genetically engineered plants for phytoremediation may not be accepted by the society. An alternative may be to exploit phylogenetic and ecological correlates of leaf mineral contents to select plant species with high radionuclide accumulation potentials (Broadley et al. 1999; Jansen et al. 2002). 9.10.3.2 Phytostabilisation This technique is applied for radionuclide-contaminated sites mainly to hold contaminants in place to prevent secondary contamination and exposure. Capturing radionuclides in situ is often the best solution at sites with low contamination levels or vast contaminated areas where a large-scale removal action or other in situ remediation is not practicable. Phytostabilisation can result in a remarkable risk reduction, especially if radionuclides with relatively short half-lives are present. Fast growing and deep-rooted plants have gained popularity in their use for radionuclides stabilization in soil. Establishment of a vegetation cover prevents the formation of windblown dust, a major pathway for human exposure at radionuclide-contaminated sites (Berti and Cunningham 2000). The ability of plants to transpire high volumes of water prevents migration of leachate towards groundwater or surface waters. This may be useful in confining areas leaking radioactive materials. Phytostabilization may be particularly suitable for controlling tailings from strip and open pit uranium mines. It is important to keep in mind that while phytostabilization of radionuclides may reduce the environmental and human health risk it does not remove the source of radioactivity from the site (Dushenkov 2003). Thus, the potential risk of radiation exposure remains, which needs be considered when decision on the best remediation approach is made. 9.10.3.3 Phytovolatilization This method exploits plants’ ability to transpire high amounts of water and is currently used for tritium (3H) remediation. Tritium is a radioactive isotope of hydrogen and is found in the environment typically as tritiated water. It decays to a stable helium with a half-life of about 12 years. As a weak β emitter, 3H is easily shielded by air and skin and produces almost no external radiation exposure. Its incorporation in water and organic compounds, however, presents a health hazard when absorbed into the body. The most cost-effective approach for the minimization of 3H risk is neither removal of 3H from the effluent nor a typical isolation of the water, but rather a change in the path of 3H exposure to the public (Dushenkov 2003). A reduction in dose by 40% can be achieved by releasing tritiated water as References 495 water vapor to the atmosphere, as opposed to allowing it to flow off site in surface water streams (Fulbright et al. 1996). For example, tritiated water applied as irrigation to the floor of forests either evaporates from the soil surface, is absorbed into the transpiration stream by plant roots or drains through the soil to the water table. Only a small portion of 3H that was absorbed by plant roots remains in the plant tissues in the form of exchangeable hydroxyl ions or is incorporated into organic molecules through photosynthesis (Dushenkov 2003). References Abdelouas A (2006) Uranium mill tailings: geochemistry, mineralogy, and environmental impact. 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