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 TcO22 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).
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