**3. The interactions of radioactive contaminants with soil matrix and the methods of their identification**

#### **3.1. The nature of radionuclide interactions with soil components**

140La, 134Ce, 144Ce/144Pr are important pollutants at the reactor stage; 90Sr, 125mTe/129Te, 131I, 134Cs, 137Cs may be released during fuel element transport; 90Sr, 95Zr/95Nb, 106Ru, 131I, 137Cs, 144Ce/144Pr, and actinides are important at the fuel reprocessing stage; 90Sr, 106Ru, 137Cs, and 144Ce/144Pr contamination may occur during fission product solidification, whereas leaching from the final disposal may result in soil contamination with 90Sr, 137Cs, and actinides [15]. In addition to fission products, several corrosion products may become significant soil pollutants. Namely, during nuclear reactor operation, most metallic surfaces oxidize and form a layer of corrosion film rich in oxides of structural elements. This layer is exposed to high pressures and temper‐ atures, where radionuclides are generated under the neutron activation [16]. Depending on the composition of the reactor materials and their trace elements, reactor type and design, thermal power, years of irradiation and shutdown period, the corrosion products and their relative proportions are different. The products of steel corrosion are 55Fe, 59Ni, 63Ni, 94Nb, 60Co, 39Ar, 54Mn, with the 60Co and 55Fe being the most important in the first 10 years following the closure of a reactor, and 63Ni, 94Nb, 108Ag in the next 50 years. Reinforced concrete's corrosion

14C, 41Ca, 152,154Eu after 50 years from the reactor shut‐down. Taking into account both fission and corrosion products, 10–20 years after the reactor shutdown the most abundant radionu‐

Another key source of soil contamination with radionuclides is nuclear weapons tests, particularly atmospheric, which have started in 1945 in the USA [17]. In the period 1945–1980, the power of USA atmospheric tests (428 megatons) was approximately equivalent of the size of 29,000 Hiroshima bombs [17]. Finally, in 1990, thanks to the moratorium signed by SSSR, UK and USA, nuclear testing was stopped. Atmospheric detonations produce radioactive debris of different particle size, which are partitioned in the tropo‐ and stratosphere and my precipitate over a period of a few minutes to 1 year, or longer [18]. The concern is especially focused onto released Pu isotopes, due to the high biological toxicity and long half‐lives of its

, 81 × 106

244Pu) [19]. Furthermore, 137Cs, 90Sr, 241Am, and 131I are the released radioactive isotopes with major impact on the environment and irradiation of the human body [20]. The mentioned isotopes were predominantly found in most of the nuclear test sites worldwide, especially in

Nuclear accident are the events that led to significant consequences to people, the environment or the facility, such as the ones in Chernobyl (Ukraine, 1986) and Fukushima (Japan, 2011). These two events caused global contamination of the environment, including air, water, soil, and living organisms. Huge amounts of radioactive elements especially 131I, 137Cs, 90Sr and the sum activity of 239Pu and 240Pu were dispersed into environment [23]. Some 40% of Europe has

illustrated by the fact that the maximum radioactive contamination in the soil in the 1993 was

H, 14C, 152,154Eu originates from graphite.

H, 60Co,55Fe, and 137Cs, whereas in the

years, respectively, for 239Pu, 242Pu, and

[24]. The size of the disaster can be

H becomes the most prominent after 10 years, and

H, 14C, 41Ca, 55Fe, 60Co, 152,154Eu, whereas 3

period 20–30 years, 63Ni, 137Cs, 60Co, and 90Sr generally prevail [16].

, 373 × 103

Considering these two groups of materials, 3

256 Soil Contamination - Current Consequences and Further Solutions

relevant isotopes (e.g., 24.2 × 103

western US soil [21, 22].

clides in contamination residues generally include 3

been exposed to Chernobyl's 137Cs at a level 4–40 kBq/m2

found to be 3500 times higher than the level before Chernobyl accident.

products are 3

Interactions of contaminants with soil matrix, and their variation with environmental param‐ eters, are essential for radionuclide transport and fate, as well as for the risks to the living organisms and the environment. The uptake of radionuclides by soil can occur through diverse modes of interactions, while at the same time, other mechanisms are responsible for their elimination from the soil matrix (**Figure 1**). Due to the dynamic nature, heterogeneity and the overall complexity of the soil as a system, studying, understanding, and predicting the radionuclides behavior are the major challenges.

Factors influencing radionuclide distribution in the soil include the source term and the release conditions, transport and dispersion mechanisms, and the properties of the ecosystem [30]. Source term (ions, colloids, particles, oxidation states, etc.) influences mobility properties of radionuclides, since the transfer of mobile species in the ecosystem is faster in respect to the transfer of particles. Furthermore, the properties of the particular radionuclide, its chemical form and the reactivity, control the nature of its retention in the soil and the affinity to certain soil constituents.

Soil properties are primarily grouped into physical (texture, structure, porosity, water, air, and heat regimen), chemical (chemical and mineralogical composition, pH, microelements, micronutrients, salinity (EC), cation exchange capacity (CEC), organic matter, etc.) and bio‐ logical (macroflora, macrofauna (rodents, insects, woodlice, mite, snails, millipedes, spiders, worms), microflora (bacteria, actinomycetes, fungi, and algae), and microfauna (nematodes and the protozoa)) [31, 32]. All five basic components of the soil, that is, minerals, water, organic matter, gasses, and the microorganisms, affect the binding and retention of the pol‐ lutants to a greater or lesser extent, depending on the pollutant type.

**Figure 1.** The mechanisms of radionuclide binding and elimination from soil matrix.

The interactions between radionuclide and the soil include physical (reversible) sorption governed by the uncompensated charges on the surface of the soil particles, and the chemical (principally irreversible) sorption through high affinity, specific interactions, and establish‐ ment of covalent bonds [33, 34]. The primary minerals in soil, mainly quarts and feldspar, are derived from the parent rock and make up most of the sand and silt fraction. Due to the relatively low specific surface area, their role in contaminant interaction is the smallest, and the attachment occurs through reversible sorption [35]. Secondary minerals, such as clay, result from physical, chemical, and biological weathering processes. Because of the unbalanced charges of structural ions, they are the carriers of permanent surface charge, which in combi‐ nation with small particle size and large specific surface area make them important matrices for contaminant retention. Furthermore, oxides and (oxy)hydroxides of Fe and Al are abundant in amorphous form, with pH‐dependent surface charge. Soil organic matter consists of chains of carbon atoms, containing polar and/or ionized surface functional groups, such as OH– and COOH–. Consequently, clay minerals, Fe, Al‐oxides, and organic matter undergo a variety of interactions with contaminants.

#### **3.2. Assessment of the radionuclide mobility in the soil**

The bonds established between the particular radionuclide and the particular soil type can be assessed by different analytical approaches. Chemical reagents of various composition, strength, and selectivity are the most widely used, in the single stage or sequential extraction protocols [30, 36–41]. The aim of such tests was assessment of the transport mechanism in a soil profile and the potential toxicity, with implications to the risks to the biota and the ground water reservoirs. In general, weaker bonds between the pollutant and the soil components signify higher mobility of radionuclide, its increased possibility to reach the plants and soil organisms and to enter into the food chain. However, the mobility and bioavailability of closely related, they cannot be equalized in the interpretation. Bioavailability processes are defined as the physical, chemical, and biological interactions that determine the exposure of plants and animals to chemicals associated with soils and sediments, they incorporate a number of steps and represent the amount of a contaminant that is absorbed following skin contact, ingestion, or inhalation [42]. On the other hand, the bioaccessibility of the contaminant is defined as its fraction soluble in the gastrointestinal tract and available for absorption.

logical (macroflora, macrofauna (rodents, insects, woodlice, mite, snails, millipedes, spiders, worms), microflora (bacteria, actinomycetes, fungi, and algae), and microfauna (nematodes and the protozoa)) [31, 32]. All five basic components of the soil, that is, minerals, water, organic matter, gasses, and the microorganisms, affect the binding and retention of the pol‐

The interactions between radionuclide and the soil include physical (reversible) sorption governed by the uncompensated charges on the surface of the soil particles, and the chemical (principally irreversible) sorption through high affinity, specific interactions, and establish‐ ment of covalent bonds [33, 34]. The primary minerals in soil, mainly quarts and feldspar, are derived from the parent rock and make up most of the sand and silt fraction. Due to the relatively low specific surface area, their role in contaminant interaction is the smallest, and the attachment occurs through reversible sorption [35]. Secondary minerals, such as clay, result from physical, chemical, and biological weathering processes. Because of the unbalanced charges of structural ions, they are the carriers of permanent surface charge, which in combi‐ nation with small particle size and large specific surface area make them important matrices for contaminant retention. Furthermore, oxides and (oxy)hydroxides of Fe and Al are abundant in amorphous form, with pH‐dependent surface charge. Soil organic matter consists of chains of carbon atoms, containing polar and/or ionized surface functional groups, such as OH– and COOH–. Consequently, clay minerals, Fe, Al‐oxides, and organic matter undergo a variety of

The bonds established between the particular radionuclide and the particular soil type can be assessed by different analytical approaches. Chemical reagents of various composition, strength, and selectivity are the most widely used, in the single stage or sequential extraction protocols [30, 36–41]. The aim of such tests was assessment of the transport mechanism in a soil profile and the potential toxicity, with implications to the risks to the biota and the ground

lutants to a greater or lesser extent, depending on the pollutant type.

258 Soil Contamination - Current Consequences and Further Solutions

**Figure 1.** The mechanisms of radionuclide binding and elimination from soil matrix.

interactions with contaminants.

**3.2. Assessment of the radionuclide mobility in the soil**

Review of the literature shows that a wide spectrum of single‐stage extraction methods for soil analysis is in use [36, 42]. Basic groups of reagents include acids, chelating agents, and salts; moreover, reagent concentrations and other experimental conditions are considerably different (**Table 1**). In contrast to the well‐established methods for the determination of soil major nutrients and fertility, the procedures for the extraction of pollutants are not standar‐ dized.


**Table 1.** The common leaching solutions in single stage soil extraction analysis [36, 42].

In addition to acidic and salt‐containing solutions, the chelating agents are applied, due to their efficiency in extracting potentially bioavailable soluble complexes of radionuclides with organic matter. The results of leaching tests represent a rough measure of mobility, as the actual mobility in the field depends also on moisture, leaching, root uptake of nutrients, activity of microorganism, and many additional factors. Furthermore, the agreement between chemically extracted and fractions available to biota should be confirmed empirically, for wide variety of contaminated samples [43].

Speciation analysis is conducted for the identification and determination of the different chemical and physical forms of elements in the soil matrix [44]. The distribution of radionu‐ clides is related to their affinity towards certain soil components; thus, they can exist as a free ions or in the form of soluble complex ions in interstitial solution; as exchangeable ions attached to the soil surface, they can be associated with soil organic fractions, occluded, or co‐precipi‐ tated with metal oxides, carbonates, phosphates, or other secondary minerals, and incorpo‐ rated inside the crystal lattices of primary minerals.

The sequential extraction protocols were primarily developed for the determination of the distribution of stable macro‐ and micro‐constituents of the soil. Identification of the mobility and availability of trace elements, both the essential ones and the pollutants, is particularly important for the improvement and protection of the plant development and growth, and for the health of the ecosystem as a whole. Different sequential extraction methods have been proposed to separate the fractions of elements from various pools. The so‐called Tessier method [44] and the method proposed by the European Community Bureau of Reference the BCR method [45], are the two commonly used protocols, while many others are based on their modifications. Additionally, a modified version of Tessier's method was proposed at the Speciation Workshop organized by the National Institute of Standards and Technology (NIST), in order to optimize the protocol of soil extraction and select operationally defined fractions which can be separated by appropriate chemical reagents [46].

Evaluation of element distribution in soils by the sequential extraction is based on the as‐ sumption that mobility decrease with each extraction step (**Figure 2**), implying that under natural conditions elements in water soluble and exchangeable fractions are the most mobile and bioavailable, whereas those in residual fractions are the most tightly bound.

**Figure 2.** Common phases in sequential extractions based on Tessier's protocol [44].

The lack of the standardized procedure for the determination of pollutant mobility makes the interpretation and the comparability of the results difficult. In addition, the effect of the reagents may be questionable. For example, extracting solution having pH 5, used for the dissolution of carbonate phase, may also sequester ions specifically sorbed onto surface of other soil constituents [47]. The fractions of pollutants are defined only operationally; thus, instead of being associated with the terms *mobility* and *bioavailability,* they should actually be related to the extracting solution or the applied protocol [48]. Nevertheless, in the scientific and the technical literature, free ions, water‐soluble complexes of radionuclides and the species associated by reversible, physical sorption, are commonly considered as the *mobile fraction* [49]. On the other hand, the term *inert species* refers to fraction of colloids and particles deposited in soils, together with the fraction of radionuclides irreversibly bound to or incorporated into the mineral lattices. The results of sequential extractions can be used for the calculation if the mobility factors of radionuclides (MF) [49]:

The sequential extraction protocols were primarily developed for the determination of the distribution of stable macro‐ and micro‐constituents of the soil. Identification of the mobility and availability of trace elements, both the essential ones and the pollutants, is particularly important for the improvement and protection of the plant development and growth, and for the health of the ecosystem as a whole. Different sequential extraction methods have been proposed to separate the fractions of elements from various pools. The so‐called Tessier method [44] and the method proposed by the European Community Bureau of Reference the BCR method [45], are the two commonly used protocols, while many others are based on their modifications. Additionally, a modified version of Tessier's method was proposed at the Speciation Workshop organized by the National Institute of Standards and Technology (NIST), in order to optimize the protocol of soil extraction and select operationally defined fractions

Evaluation of element distribution in soils by the sequential extraction is based on the as‐ sumption that mobility decrease with each extraction step (**Figure 2**), implying that under natural conditions elements in water soluble and exchangeable fractions are the most mobile

The lack of the standardized procedure for the determination of pollutant mobility makes the interpretation and the comparability of the results difficult. In addition, the effect of the reagents may be questionable. For example, extracting solution having pH 5, used for the dissolution of carbonate phase, may also sequester ions specifically sorbed onto surface of other soil constituents [47]. The fractions of pollutants are defined only operationally; thus, instead of being associated with the terms *mobility* and *bioavailability,* they should actually be related to the extracting solution or the applied protocol [48]. Nevertheless, in the scientific and the technical literature, free ions, water‐soluble complexes of radionuclides and the species associated by reversible, physical sorption, are commonly considered as the *mobile fraction* [49].

and bioavailable, whereas those in residual fractions are the most tightly bound.

which can be separated by appropriate chemical reagents [46].

260 Soil Contamination - Current Consequences and Further Solutions

**Figure 2.** Common phases in sequential extractions based on Tessier's protocol [44].

$$MF = \frac{\text{Mobile species} \,(\text{Bqm}^{-2})}{\text{Total deposition} \,(\text{Bqm}^{-2})} \times 100 \,(\% \text{)}\tag{1}$$

where the mobile species include the fraction such as H2O and CH3COONH4 extractable and that taken up by vegetation from the same site.

Although none of the methods can provide the absolute quantities associated to the specific component of the soil, such analyses represent valuable tool in elements mobility and availa‐ bility assessment, and for tracking the effectiveness of soil remediation actions. Experiences achieved by practicing single and sequential chemical extractions reveal advantages of these methods but also a need for further research and developments due to increasing soil con‐ tamination which requires fast, reliable, and cost‐effective assessment.

In addition to extraction methods, studies of the radionuclide retention mechanism can be complemented by the determination of the type of the surface complexes, identification of the radionuclide incorporation in the crystal lattice of existing minerals, or the formation of new solid phases, etc., for which instrumental techniques are applied (X‐ray powder diffraction (XRD), Fourier transform infrared spectroscopy (FTIR), X‐ray photoelectron spectroscopy (XPS), X‐ray absorption spectroscopy (XAS), scanning electron microscopy with energy dispersive X‐ray spectroscopy (*SEM/EDS*), etc.) [50, 51].

Furthermore, bioassay tests involving plants, animals, and microorganisms, are valuable for the analysis of radionuclide mobility and bioavailability [52]. Soil‐to‐plant transfer factors (TF) have been widely used in radioecology, in order to quantify the availability of soil radionu‐ clides for plant uptake [36]:

$$TF = \frac{\text{Plant activity concentration} \left(\text{Bq} \,\text{kg}^{-1}\right)}{\text{Total soil activity concentration} \left(\text{Bq} \,\text{kg}^{-1}\right)} \tag{2}$$

As soil‐to‐plant transfer considerably differs between different plant species and the seasons, this method also gives crude estimations of potential radionuclide bioavailability. In spite of limitations, transfer factors are currently accepted as the most practical way of describing plant uptake. Also, several in vitro methods have been developed for the prediction of the relative bioavailablity of the contaminants, using physiologically based fractionation schemes [42, 53]. These methodologies mimic key processes that take place in vivo, such as contaminant dissolution, and after establishing a strong correlation between the in vivo and in vitro results, these methods have a potential to overcome the time and expense limitations of in vivo studies.

#### **3.3. Factors influencing radionuclide mobility in the soil**

A capacity of the soil itself to immobilize radionuclide is the main factor controlling activity concentrations available to biota, and it operates in conjunction with the numerous external factors. Soil texture and structure, mineral composition, organic components, redox potential (Eh) and pH, as well as rainfall, climate changes, and soil management, are recognized as important for radionuclide mobility [54]. The pH of the soil, cation exchange capacity (CEC), and total organic carbon (TOC) are the physicochemical characteristic most often correlated with the distribution of the radionuclides [40]. Alkaline soils are characterized by the presence of carbonates and have a high saturation of base cations (K+ , Ca2+, Mg2+, and Na+ ), whereas acidity in soils comes from H+ and Al3+ ions in the soil solution and sorbed to soil surfaces. The surface charge of minerals is a major contributor to soils CEC and influences the soil's ability to retain important nutrients and the pollutants. The texture of a soil is based on the relative content of sand (0.05–2.00 mm), silt (0.002–0.05 mm), and clay (<0.002 mm) fraction. Due to the finest granulation, clays minerals exhibit the largest surface area, important for soil chemistry and CEC, but also for water‐holding capacity important for transporting nutrients and pollutants to soil organisms and plants. In addition, soil organic matter significantly contrib‐ utes to the soil CEC and to the water‐holding capacity.

**Radionuclide Cs Sr, Ra U, Pu I Chemical form Cs+ Sr2+ PuO2 2+, Pu(NO3) 3+ I2, I− , IO3 − CH3I Mobility** pH decrease Increase Increase Increase Clay content decrease Increase Increase Increase Sand content decrease Decrease Decrease Decrease Increase Humus content low Not clear Decrease Decrease CEC decrease Increase Increase Increase Aging Decrease Weak effect Decrease

Based on the literature data, the influence of soil properties and other condition on the mobility of some important pollutants is given in **Table 2**.

**Table 2.** The effect of soil physicochemical properties and aging on the mobility of radionuclides [55, 56].

Apart from soil type, different sources of variability may influence the fractionation patterns and cause the shift from less available to more available fractions, or vice versa. Generally, the increase of contaminant concentration not only increases the overall activity in the soil but also leads to redistribution from the less to the more available fractions [57]. Radioactive contam‐ ination introduces new elements into the ecosystem and, in distinction from the transport of stable elements and NORM, transfer of contaminants through the trophic chains occurs under non‐equilibrium conditions. Consequently, ageing affects a decrease in the chemical mobility and biological availability of most of the radioactive pollutants [58]. The ageing process actually involves a set of reactions related to the enhancement of radionuclide sorption and fixation by the soil solid phase (i.e., the precipitation or penetration into the crystalline lattices of different mineral constituents. Aging exhibits a different effect on different ions. Increased contact times (months to years) were found to affect gradual reduction of Co2+ ions mobility [57, 59]. Time‐dependent studies on the variation in Cs+ bioavailability have revealed that over years, a decrease in the labile fraction of 137Cs in soils was correlated with a decrease in soil‐to‐ plant transfer [60]. In contrast, due to low sorption affinity of 90Sr towards soil constituents, impact of aging is very weak considering 90Sr speciation [40]. The behavior of Sr2+ and its uptake by living organisms are controlled by its similarity to calcium; thus, regardless of the soil type, contamination level, and aging time, it was largely found in water‐soluble and ion‐exchange‐ able fractions of soil. Seasonal effects may also cause variations in radionuclide mobility, and these effects can be controlled by appropriate sampling plan [57].
