117
1988; Ishchenko and Butnik 1990). According to some estimates (Sultanbaev 1974),
the content of
238
U in the clay fraction is up to three times higher than its concentration in unfractionated soil. At the same time, little information indicates a lack of a
close relationship between the sorption of radionuclides and the granulometric composition of soils. For example, for soils in the taiga zone, up to 50% of radionuclides
of the uranium series are sorbed on the surface of particles with a size of 1–100 μm,
and consolidation of aggregates in this range is accompanied by a decrease in the
content of radioelements (Taskaev 1979).
The strong fixation of
238
U in soils occurs due to its sorption by organic substances and clay minerals (illite, montmorillonite and kaolinite) (Taskaev 1979;
Alexakhin et al. 1990; Rusanova 1998; Ranson 1973). Limestone weakly absorbs
238
U, and silicon dioxide does not absorb the uranium at all (Alexakhin et al. 1990;
Savenko 2001). The sorption centres of two types are significant for sorption of
radionuclides by clay minerals. These centres are located on the basal surfaces of
mineral particles, where ion exchange takes place, and on the lateral faces with the
formation of strong sorption complexes. According to Kornilovich et al. (2001),
uranium was primarily absorbed on the lateral faces of clay minerals. The X-ray
diffraction analysis (Iskra and Bakhurov 1981) showed that binding takes place
according to the ion-exchange mechanism.
238
U competes with cations Ca
2+
, Ba
2+
and Mg
2+
for sorption centres, but not with Na
+
and K
+
. According to the structural
data, uranyl forms dissolved and absorbed by clay minerals are identical. In subacidic conditions, they correspond to the hexahydrated ion (Iskra and Bakhurov
1981). In the pH range of 6–12, the sorbed forms of uranyl are represented by
UO 2 OH
+
and (UO 2 ) 2 (OH). Absorption of UO 2 OH
+
predominates at pH 6; with a
decrease in acidity, the ratio of sorption of (UO 2 ) 2 (OH) increases. The isotherms of
238
U (VI) sorption by clay minerals have a maximum at pH 3–6 with a sharp decrease
in the absorption parameters to the sides of this interval (Kornilovich et al. 2001).
The influence of the capacity of cation exchange on the sorption of
238
U by soils
is practically absent (Bondietti and Tamura 1980). For example, the removal of
organic matter and free ferrous oxides from the clayey fraction of silty loam reduces
its cation-exchange capacity, without changing the absorption rate of
238
U. Along
with this, there is an opinion (Malikov et al. 1988a) that in surface media, the Fe
content controls the pH-dependent adsorption of
238
U. Oxide forms of metals (goethite, hematite, boehmite, etc.), on the one hand, themselves serve as sorbents of a
radionuclide, and on the other, when they are adsorbed, these oxides form films on
the surface of clay minerals changing their ion-exchange properties. The distribution coefficient of the radionuclide between solutions of its salts and Fe minerals
reaches 2 × 10
6
mL g
−1
, which is 100 times greater than for secondary minerals.
The behaviour of
238
U and its forms in soils are determined both by the state of
the radionuclide itself and by the physical and chemical properties of the soils. The
original form of the radionuclide fallout plays an essential role in the processes of
sorption and desorption of
238
U. For example, in 2 years later after application, the
sorption of
238
U by five types of soil (17 soil differences) with contrasting physical
and chemical properties and indicators of texture when introducing a radionuclide
in the form of a solution of uranyl nitrate is 82.2–99.2% and desorption by water is
The Behaviour of Uranium in Soils and the Mechanisms of Its Accumulation…
1988; Ishchenko and Butnik 1990). According to some estimates (Sultanbaev 1974),
the content of
238
U in the clay fraction is up to three times higher than its concentration in unfractionated soil. At the same time, little information indicates a lack of a
close relationship between the sorption of radionuclides and the granulometric composition of soils. For example, for soils in the taiga zone, up to 50% of radionuclides
of the uranium series are sorbed on the surface of particles with a size of 1–100 μm,
and consolidation of aggregates in this range is accompanied by a decrease in the
content of radioelements (Taskaev 1979).
The strong fixation of
238
U in soils occurs due to its sorption by organic substances and clay minerals (illite, montmorillonite and kaolinite) (Taskaev 1979;
Alexakhin et al. 1990; Rusanova 1998; Ranson 1973). Limestone weakly absorbs
238
U, and silicon dioxide does not absorb the uranium at all (Alexakhin et al. 1990;
Savenko 2001). The sorption centres of two types are significant for sorption of
radionuclides by clay minerals. These centres are located on the basal surfaces of
mineral particles, where ion exchange takes place, and on the lateral faces with the
formation of strong sorption complexes. According to Kornilovich et al. (2001),
uranium was primarily absorbed on the lateral faces of clay minerals. The X-ray
diffraction analysis (Iskra and Bakhurov 1981) showed that binding takes place
according to the ion-exchange mechanism.
238
U competes with cations Ca
2+
, Ba
2+
and Mg
2+
for sorption centres, but not with Na
+
and K
+
. According to the structural
data, uranyl forms dissolved and absorbed by clay minerals are identical. In subacidic conditions, they correspond to the hexahydrated ion (Iskra and Bakhurov
1981). In the pH range of 6–12, the sorbed forms of uranyl are represented by
UO 2 OH
+
and (UO 2 ) 2 (OH). Absorption of UO 2 OH
+
predominates at pH 6; with a
decrease in acidity, the ratio of sorption of (UO 2 ) 2 (OH) increases. The isotherms of
238
U (VI) sorption by clay minerals have a maximum at pH 3–6 with a sharp decrease
in the absorption parameters to the sides of this interval (Kornilovich et al. 2001).
The influence of the capacity of cation exchange on the sorption of
238
U by soils
is practically absent (Bondietti and Tamura 1980). For example, the removal of
organic matter and free ferrous oxides from the clayey fraction of silty loam reduces
its cation-exchange capacity, without changing the absorption rate of
238
U. Along
with this, there is an opinion (Malikov et al. 1988a) that in surface media, the Fe
content controls the pH-dependent adsorption of
238
U. Oxide forms of metals (goethite, hematite, boehmite, etc.), on the one hand, themselves serve as sorbents of a
radionuclide, and on the other, when they are adsorbed, these oxides form films on
the surface of clay minerals changing their ion-exchange properties. The distribution coefficient of the radionuclide between solutions of its salts and Fe minerals
reaches 2 × 10
6
mL g
−1
, which is 100 times greater than for secondary minerals.
The behaviour of
238
U and its forms in soils are determined both by the state of
the radionuclide itself and by the physical and chemical properties of the soils. The
original form of the radionuclide fallout plays an essential role in the processes of
sorption and desorption of
238
U. For example, in 2 years later after application, the
sorption of
238
U by five types of soil (17 soil differences) with contrasting physical
and chemical properties and indicators of texture when introducing a radionuclide
in the form of a solution of uranyl nitrate is 82.2–99.2% and desorption by water is
The Behaviour of Uranium in Soils and the Mechanisms of Its Accumulation…