Assessment of Radionuclide Transfer in Terrestrial Ecosystem

Abstract

Radioactivity in the terrestrial environment is present since time immemorial and contributed by natural radionuclides of ²³⁸U and ²³²Th series and artificial radionuclides of ¹³⁷Cs, ⁹⁰Sr, etc. The major portion of food consumed by humans is grown in the terrestrial environment, and these radionuclides deliver radiation dose to humans and non-human biota through various exposure pathways. Soil is an important constituent of terrestrial environment comprising of mixture of sand, silt and clay in varying proportions, which give each soil type, characteristic texture and properties. The important physicochemical processes by which soils provide the nourishment for plants are controlled largely within the clay fraction of soil. The formation of soil in various agroclimatic zones is influenced by climatic factors of precipitation and temperature. Uptake of radionuclide depends on its availability in root zone of plant and its chemical form available for transport to root zone and translocation to edible portions of the plant. The major pathways and compartments considered for assessment of the impact to terrestrial environment are foliar interception, translocation and uptake from root zone by plants. Partitioning of radionuclide between soil and interstitial water is described by distribution coefficient K_d expressed in L kg⁻¹. There is a strong dependence of K_d for uranium, particularly within the pH range of 5–7, where K_d (U) is observed to be 10 times higher. The sorption behavior (K_d Iodine) of iodine is more complex. Data generated on soil-to-plant transfer factor (TF) or concentration ratio (CR) is an important parameter used in the Radiological Environment Impact Assessment (REIA) models. Important factors affecting TF are chemical and physical characteristics of soils, agricultural practices adopted, crop types, frequency of rain events and dietary consumption practices. Studies indicated that sorption of radiocesium in soils is influenced by ionic exchange processes at frayed edge sites (FES) regular exchange sites present in soil. Soils containing 14–50% organic matter, effectively complex Ra, more than its parent element Uranium. Lichens contain significantly higher concentrations of ¹³⁷Cs, ²¹⁰Po and ²¹⁰Pb as compared to vascular plants.

1 Introduction

Radionuclides are ubiquitously present in the terrestrial environment and have originated from primordial times. They are being continuously added into the ecosystem through cosmogenic sources and also through anthropogenic activities. These radionuclides navigate diverse pathways, posing potential exposure risks to both humans and non-human biota, resulting in internal and external radiation exposure. Notably, food serves as a conduit for both natural and artificial radionuclides, contributing to internal doses when ingested [1]. In this chapter, a detailed discussion would be presented on these radionuclides, specifically in the terrestrial environment.

2 Input of Radionuclides to Terrestrial Ecosystem

Main sources of radioactivity in the terrestrial environment are primordial and human induced in origin. Predominant, naturally occurring long-lived radionuclides are ⁴⁰K (T_1/2 of 1.28 × 10⁹ yrs), ²³²Th (T_1/2 of 1.41 × 10¹⁰ yrs), and ²³⁸U (T_1/2 of 4.5 × 10⁹ yrs) and their progeny. The natural radionuclide composition of soils and rocks plays a key role in determining the terrestrial component of background radiation. It is an important factor influencing the variability in radiation levels across different geographical locations. Thus, there are large variations in the exposure from terrestrial radiation depending on factors such as geographical locations, the amount and type of radionuclide present in the soil and habitat of the population, e.g., external terrestrial radiation in India varies from 576 µGy/y to 910 µGy/y [2, 3]. UNSCEAR report suggests that about 95% of the people reside in areas where, the average dose rate varies from 0.3 to 0.6 mSv/y and about 3% are exposed to 1 mSv/y. Some places on earth lead to high levels of terrestrial radiation exposure, e.g., Pocos de Caldas in Brazil can cause an average radiation exposure of about 250 mSv/y. In South West Coast of India, a long strip of land containing Th-rich sands delivers an average dose of 3.8 mSv/y and at some locations the dose rates are even as high as 17 mSv/y. Ramsar in Iran has springs rich in Radium and levels of dose could be as high as 400 mSv/y. The majority of the world population receives an average dose of 0.35 mSv/y ± 25% due to external/outdoor terrestrial irradiation.

The important sources of artificial radioactivity in terrestrial environment are from global dispersal of radioactive materials due to nuclear weapons tests conducted in the atmosphere, especially during the late 1950s and early 1960s, fallout in specific regions due to nuclear tests conducted in proximity, such as the Marshall Islands and Mururoa Atoll, disposal of radioactive wastes in certain areas, like the dumping of wastes in the Sea of Japan, instances where radioactive materials are accidentally released, such as accidents involving the SNAP-9A satellite, nuclear-powered vessels, and nuclear weapons, accidents at nuclear power plants [4]. Routine operations in nuclear power plants result in small, controlled discharges of radioactive substances into the environment but accidents at nuclear power plants had caused significant releases of radioactivity into the environment. In addition to this, some contribution is from technologically enhanced naturally occurring radionuclides due to usage of fossil fuel based, processes and industries. Medical and other industrial procedures also employ radioactivity to some extent and these can contribute to radioactivity in the terrestrial ecosystem.

The anthropogenic radionuclides ¹³⁷Cs (cesium-137) and ⁹⁰Sr (strontium-90), formed by nuclear fission are of concern due to their impact on humans and the environment. The reduction in their environmental concentrations over the last few decades is attributed to factors such as the cessation of atmospheric testing, dilution, and radioactive decay. The surveillance of natural radionuclides in the environment initially did not have a priority compared to man-made radioactivity during the early stages of the nuclear industry. However, the scenario has changed, and monitoring of natural radionuclides has gained importance with the increase in non-nuclear applications, such as phosphate processing plants, offshore oil and gas installations, and ceramic industries. These activities can lead to a build-up of natural radioactivity, particularly ²²⁶Ra, ²²⁸Ra, ²¹⁰Po, and ²¹⁰Pb [5, 6].

3 Sources of Radionuclides in Terrestrial Environment

3.1 Sources of Natural Radionuclides in Terrestrial Environment

3.1.1 Cosmogenic Production

Cosmogenic nuclides are continuously formed by the bombardment of upper atmospheric gases with cosmic radiation from space. Several important cosmogenic nuclides include ³H (tritium), ⁷Be (beryllium-7), ¹⁴C (carbon-14), and ²²Na (sodium-22). These nuclides enter the terrestrial environment through fallout from the atmosphere, settling on exposed surfaces of soil or vegetation, and subsequently getting assimilated. About 99% of tritium is incorporated into water and follows the water cycle from the atmosphere to the sea. Production of tritium ranges from 50 to 70 PBq (0.15 to 0.20 kg) annually and a global natural inventory of approximately 1300 PBq (3.5 kg) of tritium has been evaluated by UNSCEAR.¹⁴C is produced in the upper troposphere and the stratosphere by thermal neutrons absorbed by nitrogen atoms. The highest rate of ¹⁴C production occurs at altitudes of 9–15 km and at higher geomagnetic latitudes. Production rates vary due to changes in the cosmic ray flux caused by solar cycle modulation and variations in the Earth's magnetic field.

3.1.2 Primordial Origin

These group of nuclides are present in the terrestrial environment since formation of earth and consist mainly of radionuclides in the series of ²³⁸U (uranium-238), ²³²Th (thorium-232) and ²³⁵U (uranium-235), and ⁴⁰K, the radioactive isotope of potassium. Natural processes like weathering, resuspension, washout, and river transport contribute to the distribution of radionuclides from terrestrial to other environmental compartments. The radioactivity levels in soil are influenced by the type of rock from which the soil originates. Igneous rocks like granite tend to have higher radiation levels, while sedimentary rocks generally have lower levels. Shales and phosphate rocks are exceptions with relatively high radionuclide content. Phosphate rocks, often used as fertilizers, can have significantly higher concentrations of radionuclides, particularly uranium. Most crustal rocks contain a few ppm of uranium, averaging around 2.8 ppm. Phosphate rocks, on the other hand, may contain much higher amounts, around 120 ppm or more Radionuclides such as ²³⁸U and ²³²Th, along with their daughters like ²²⁶Ra and ²²⁸Ra, are widespread in both abiotic and biotic materials. This ubiquitous presence contributes to the overall terrestrial environmental radiation [5, 6, 7].

3.1.3 Coal Burning

Fossil fuels are integral part of the energy sources in the modern world. Emissions due to household coal burning were reported to be three times higher than that for electricity generation. Fly-ash generated from coal burning contains, average concentration of ²³⁸U as 1850 Bq kg⁻¹, ²²⁶Ra as 370 Bq kg⁻¹, and ²¹⁰Pb as 3700 Bq kg⁻¹. ²²²Rn is released during the burning and excavation of coal, uranium ores etc. and decays to ²¹⁰Pb in the atmosphere. These are referred to as technologically enhanced naturally occurring radionuclides.

3.2 Sources of Anthropogenic Radionuclides in Terrestrial Environment

3.2.1 Nuclear Weapon Tests

Global nuclear fallout from atmospheric weapons tests carried out in the 1950s and 1960s, close-in fallout, fallout from the Chernobyl and Fukushima accidents, contributions from nuclear weapon test sites, disposal of nuclear wastes into the world's oceans and seas, accidental losses, nuclear submarine accidents and discharges of radionuclides from nuclear installations have resulted in planned and accidental release of radionuclides to terrestrial environment [4].

Refractory radionuclides (e.g., ⁹⁰Sr, actinides) were assumed to be deposited locally, indicating a more limited spread from their sources whereas Volatile Radionuclides (e.g., ¹³⁷Cs, ¹³¹I) were deposited locally and regionally, suggesting a wider distribution over larger geographical area [8]. Anthropogenic radionuclides are predominantly short-lived, with only a few long-lived radionuclides. The contribution of long-lived radionuclides to the inventory of artificial radionuclides in the terrestrial environment is small. Presently, ¹³⁷Cs is estimated to be the main source of anthropogenic terrestrial radioactivity along with other important radionuclides of ⁹⁰Sr, ^239+240Pu, ²⁴¹Am, ³H, and ¹⁴C, released in large quantities during nuclear tests [9]. Global and local fallout accounts for 90% of the total ¹³⁷Cs isotope radioactivity whereas Reprocessing plants (7%) and Chernobyl accident (3%) contribute to the remaining 10% of ¹³⁷Cs isotope radioactivity [10]. The expected ratio in global fallout is 1.45, based on measured fission product yields from megaton weapons and data on half-lives and decay schemes. This ratio is assumed to be consistent in all fallout on soil [11]. Among released fission product radionuclides, ¹³⁷Cs need special mention because of its production in large quantities during fission, having intermediate half-life, decay by high energy emissions, chemically reactive and high solubility.

Tritium was released during atmospheric nuclear tests conducted in the past decades. About 420 kg of tritium was released in the northern hemisphere, and 140 kg in the southern hemisphere. Nuclear tests conducted from 1945 to 1963 released about 560 kg of tritium into the environment. About nine-tenths (90%) of the released tritium went into the sea, one-tenth (10%) entered continental waters and approximately 1% entered the atmosphere.

Global fallout depended on the atmospheric transport of debris from the stratosphere to the troposphere. Maximum transfer occurred at mid-latitudes. The global precipitation pattern and the locations of the nuclear test sites also influenced the global distribution of weapons fallout. The Northern hemisphere reflected larger fallout compared to the Southern hemisphere. This discrepancy is attributed to the relatively low number of test explosions in the Southern hemisphere and limited atmospheric exchange between the northern and southern stratospheres. Approximately 76% of the fallout reached the Northern hemisphere, and 24% reached the Southern hemisphere. Fallout was maximal at mid-altitudes (30–60° latitude) and minimal at the equator and poles.

3.2.2 Nuclear Fuel Cycle

Each part of the nuclear fuel cycle from mining and milling, through fuel fabrication, reactor operation and reprocessing of spent fuel and waste management practices releases radioactive materials into the environment. Radioactive waste disposal activities are the greatest contributor to the public dose, according to the recent estimates of UNSCEAR 2000. Due to advent of new environmentally benign technologies, and modifications in dose calculation methodologies, the significance of mining and milling activities has been re-evaluated.

3.2.3 Nuclear Accidents: Chernobyl and Fukushima Daiichi

The Chernobyl accident in 1986 released radionuclides during explosion and included very short-lived fission products, resulting in very high dose rates in the adjacent areas. Due to graphite fire, significant releases of semi-volatile or refractory elements (0.4–1.5% of the reactor core inventory) such as radio-strontium, actinides and volatile radionuclides from Chernobyl core occurred and were mainly deposited within a distance of 100 km from the site due to the large particle sizes. Release of ¹³¹I, ¹³⁷Cs, and ⁹⁰Sr was estimated to be 1760 PBq, 85 PBq and 10 PBq, respectively, as per UNCEAR 2008 estimates.

The recent accident of March 2011 at Fukushima was solely of chemical nature and only gas phase radionuclides of about 520 PBq were released of which ¹³¹I, ¹³⁷Cs, and ⁹⁰Sr were estimated to be 150 PBq, 12 PBq, and 0.2 PBq, respectively. The overall releases from Fukushima Daiichi were about 10% of the Chernobyl releases.

In terms of radionuclide signatures, radiocesium releases from Chernobyl and Fukushima were clearly distinct. Fukushima-derived radiocesium carried a ¹³⁴Cs/¹³⁷Cs activity ratio of 0.98 ± 0.01 Bq/Bq, whereas radiocesium from Chernobyl had a ¹³⁴Cs/¹³⁷Cs signature of 0.5–0.6 Bq/Bq.

3.2.4 Releases from Nuclear Reactors Under Normal Operation

Monitoring of nuclear reactors released radionuclides that are ensured before discharge to the environment as part of routine practice. The monitoring process includes the measurement of various radionuclides, such as noble gases, tritium, and traces of fission and activation products. Effective doses to the members of public from the discharged effluents are assessed annually and observed to be insignificant. Nuclear facilities adhere to stringent regulations governing waste management practices strategically designed to minimize stress on the environment. These regulations are designed to ensure the safe and controlled disposal of radioactive waste, minimizing potential risks. However, it is a source of input for radionuclides in the environment posing a potential long-term risk in future.

3.2.5 Other Sources of Radioactivity

Accidents related to nuclear weapons and satellites powered by trans-uranium elements resulted in the dispersion of radionuclides into the environment. Depleted uranium (DU) is a by-product resulting from the production of enriched uranium for various applications, including nuclear reactors and weapons. Due to its unique properties, DU finds applications in both military and civil sectors. Common uses include ammunition, counterweights in aircraft, radiation shielding, and as a chemical catalyst.

4 Mechanism of Transport of Radioactivity and Concept of Transfer Factor

Assessment of impact and fate of released radionuclides as planned or accidental in the environment is carried out by compartmentalization of the ecosystem where the radionuclide is present. Compartmentalization of ecosystems provides a structured approach to understand the movement and distribution of radionuclides. It is required to ensure the radiological protection of the environment, the human population, surrounding biota and the whole ecosystem in-totality. The pathway followed by a radionuclide once it reaches the environment is divided into compartments and the ratio of the concentration of radionuclide in one compartment to the other under equilibrium is defined as transfer factor. Equilibrium assumption is valid during controlled and continuous releases from a nuclear facility where movements into and out of compartments stabilize. But in the case of short term or acute releases, during an accident, equilibrium cannot be assumed, and variation in transfer rates between compartments must be considered.

The major pathways and compartments considered for assessment of the impact to terrestrial environment are (Fig. 1).

Fig. 1

Processes for transfer of radionuclides through terrestrial environment

• 1. Foliar interception and translocation in plants: aerial dry deposition

• 2. Wet deposition and translocation in soils: K_d

• 3. Uptake from root zone by plants.

Dry deposition, wet deposition, interception on foliar surfaces, weathering and translocation define the various processes by which radioactivity released to environment reaches agriculture systems and finds its way to soil, the ultimate sink. The uptake of radionuclides by plants through the rooting zone and their accumulation in the edible parts of the important vegetation or crops is the next compartment where the radionuclides can be transferred to humans directly or via the animal-meat-milk pathway. Most of the studies are concentrated on transfer of radionuclides through food chains to humans and do not consider the impact or transfer to non-human species specifically. However, most of the times they are also applicable for assessments of radionuclide transfer to non-human species.

4.1 Foliar Interception and Translocation in Plants: Aerial Dry Deposition

Radionuclides released into the atmosphere undergo dispersion and can deposit on various surfaces, including leaves, plant canopies, and open grasslands. The particle size spectrum is a crucial parameter for dry deposition, influenced by the characteristics of the release and the distance traveled from the release point. Dry deposition is particularly effective for small particles and reactive gases as these are more readily deposited on plant surfaces. Dry and wet deposit interception represents a critical link between the dispersion of radionuclides released to the atmosphere and their transport in food chains. The development of the plant canopy and current weather conditions plays a substantial role in dry deposition. Atmospheric conditions, wind patterns, and the morphology of the plant canopy impact how radionuclides settle on vegetation.

Translocation is the process involving the redistribution of radionuclides from contaminated parts of the plant to other parts that are not directly contaminated. It is quantified using a ratio defined as the activity of the edible part within 1 m² of crops at harvest time (expressed in Bq m⁻²) divided by the foliage activity of 1 m² of the crop at the time of deposit (expressed in Bq m⁻²). The resulting value is typically expressed as a percentage. The vegetative growth versus percentage translocation for cesium showed that translocation factor values for cesium vary among different crops, such as wheat, barley, and rye and range from 1 to 27%. Maximum translocation occurs around 45 days before harvesting, corresponding to the flowering stage of the plants. Translocation factor values for strontium also vary for crops like wheat, barley, and rye. Depending on the time between foliar contamination and harvesting, translocation factor values for strontium range from 0.007% to 8.5%. Maximum translocation value is reached when foliar contamination occurs about 30 days before harvesting, corresponding to the grain growth stage of the plants [12].

4.2 Wet Deposition and Translocation in Soils (K_d)

Radioactivity dispersed in the atmosphere is removed from the system by aerosols of water in the atmosphere or rainfall and get deposited majorly on the open grasslands and soil. They may be intercepted by the plant canopies and the standing crop but are mostly washed off depending on the precipitation period and intensity. In case of long-term precipitation and irrigation, there are indications that the interception of wet deposited polyvalent cationic radionuclides may be larger by a factor of 2 or 3, than given by the approach based on water storage capacity and leaf area index. Due to the negative charge of the plant surface, the radionuclide retention of cations by the leaf is enhanced. The deposited radioactivity in the soil gets transported horizontally and vertically which again is the characteristic of the soil and the radionuclide. The main factor affecting these processes in soil is the distribution coefficient K_d. The radionuclide then becomes available in the system for uptake by plants or direct intake to burrowing animals or ingested with grass by the ruminants and herbivores.

4.3 Uptake from Root Zone by Plants

Data generated on uptake of radionuclides by edible part/compartment of a plant, i.e., the soil-to-plant transfer factor (TF) or concentration ratio (CR) is an important parameter used in the Radiological Environment Impact Assessment (REIA) models. Soil-to-plant transfer factors were defined on a dry weight basis for both plants and soil to minimize uncertainties in the assessment process. A root zone depth of 10 cm for grass and 20 cm for all other crops (fruit trees included) was standardized by International Union of Radioecology (IUR) for assessing activity in the soil, which was available for uptake by plants [13].

The TF is defined as a ratio of the activity concentration of radionuclide in the whole plant or plant compartment (Bq kg⁻¹ dry weight) to that in the soil (Bq kg⁻¹ dry weight). TF of individual radionuclides was calculated using the following relationship [14]

$$ {\text{TF}} = {}^{{\text{P}}}{\text{A}}_{{\text{R}}} /{}^{{\text{S}}}{\text{A}}_{{\text{R}}} $$

where ^PA_R is the activity concentration (Bq kg⁻¹ dry weight) of “radionuclide R” in the plant (or plant compartment), and ^SA_R is the activity concentration (Bq kg⁻¹ dry weight) of “radionuclide R” in soil. For stable elements, it is the ratio of the concentration of stable elements (mg kg⁻¹ dry weight) in the plant (or plant compartment) to soil (mg kg⁻¹ dry weight).

Aggregated transfer factors (T_ag) defined as the ratio of radionuclide activity concentration in natural or semi-natural products (Bq kg⁻¹ fresh or dry) to the total soil deposition (Bq m⁻²) are also used for assessing radionuclide transfers to wild animals, particularly those hunted for meat and various types of natural or semi-natural vegetations. They quantify radionuclide availability to animals or other products [15].

Soil-to-plant transfer factor data is often obtained in controlled environments (pot experiments) or open fields using radionuclides. Typical observation periods (less than one year) from farming to harvest may not be suffice for all radionuclides to attain equilibrium with the environment. Freshly added radionuclides may lead to higher TF values in the short term, especially when not in equilibrium due to higher bioavailability. TF of Technetium generated in laboratory conditions is significantly higher than data from field experiments [16].

As part of IAEA initiative program, significant work had been carried out in compiling transfer parameter data for human food chains. These are Biosphere Modelling and Assessment (BIOMASS) in 1996–2002, Environmental Modelling for Radiation Safety (EMRAS I and II) in 2003–2007 and 2009–2011, and Modelling and Data for Radiological Impact Assessments (MODARIA I and II) in 2012–2015 and 2016–2019 [28]. IAEA Technical Report Series No 472 continues to be an important document to provide guidelines for REIA of nuclear facilities to ensure compliance. Working group four of the IAEA program MODARIA-II focused on transfer processes and data for radiological impact assessment including improving distribution coefficient K_d values in soils. Globally generated data exhibits a wide range of TF values for the individual radionuclides (or elements) for the same plant (and compartment) due to fluctuations concerning geological and geographical locations [17].

5 Factors Affecting Radionuclide Transfer to Plants

Transfer parameter values depend on climatic factors that are associated with specific environmental, agricultural, and managerial features. Important factors influencing the radionuclide transfer to humans from crops are chemical and physical characteristics of soils, agricultural practices adopted, crop types, frequency of rain events and dietary consumption practices. It is therefore appropriate to collate adequate transfer factor data for different climatic conditions like temperate, tropical, arid, and semi-arid environments [18].

Studies indicate that transfer factor values vary widely (Fig. 2) due to their dependence on various factors such as the form in which the activity (particles, aerosols, solutions) is present in the soil, the physicochemical properties of the radionuclide, time after entry into the soil, type and characteristics of the soil environment, crop type, crop management practices (fertilizers, irrigation, plowing, liming) climate conditions, and the experimental conditions of data collection [19,20,21].

Fig. 2

Factors affecting root uptake of radionuclides

5.1 Deposition Properties

Radionuclides released during normal and accidental conditions from nuclear fuel cycle can be grouped according to their ionic form and solubility in soil–water system. Radio isotopes of actinides, including Pu, Th, and Am, exist in the most stable states, mainly as oxides, carbonates or hydroxides and exhibit extremely low solubility in the soil–water system. Different actinides show variability in biotic transfer rates. Np and Cm exhibit slightly higher biotic transfer rates compared to Pu, Th, and Am. Uranium displays the widest range in bioavailability because it forms both cationic and anionic species [19, 21]. After release into the soil, radionuclides form complexes and different species due to their interaction with anions and organic matter of soil. Elements like technetium, selenium, and iodine, forming anions, become more mobile and plant-available. However, elements like Ni, Cs, Sr, Ag, Sn, Nb, and naturally occurring radionuclides (e.g., ^226/228Ra, ²¹⁰Pb, ²¹⁰Po) exist in cationic forms, forming stable compounds that limit their mobility and bioavailability [22,23,24,25,26].

5.2 Time After Contamination

Due to the presence of microorganisms in soil, radionuclides may undergo fixation in soil minerals over time, thereby reducing their migration to root zone and resulted in reduced uptake by farm crops with time had been observed in various agricultural ecosystems [27]. Soil-to-plant transfer of ¹³⁷Cs is observed to reduce significantly over time and a factor of 10 reductions is noted once ¹³⁷Cs achieves fixation in soil and equilibrium conditions. The availability of ⁹⁰Sr also decreases with time, but it is less pronounced compared to ¹³⁷Cs [28].

5.3 Crop Properties and Cultivation Practices

Transfer of radionuclides from soil to crop is highly variable due to inherent biological variability in different varieties and species within same plant. Variability arises due to differences in metabolic and biochemical mechanisms governing radionuclide uptake, varying requirements of micro and macronutrients by plants, variation in roots penetration depth in the soil, yield and vegetative period of plant, etc. Radionuclides tend to accumulate in greater concentrations in leaves and stems compared to generative parts. Legumes exhibit more effective accumulation of ⁹⁰Sr compared to cereals. Soil properties are affected by crop cultivation practices and lead to redistribution of radionuclides in the root zone, which changes radionuclide uptake by crops. For example, under irrigation, radionuclide accumulation in plants is enhanced by a factor of 1.2–2.0. Ploughing decreases the transfer of radionuclides to crops (factor of two). Liming reduces plant uptake of ⁹⁰Sr and ¹³⁷Cs (factor of 2–3 reduction) and application of fertilizers increases radionuclide uptake (factor of 2–5). Hence, cultivation practices should be considered in estimating parameters for soil-to-plant transfer and uptake by plants [26].

5.4 Radionuclide Properties

Natural radionuclides of ²³²Th, ²³⁸U, ²²⁶Ra, ²¹⁰Pb, ²¹⁰Po exhibit relatively low transfer factor values from soil to plant systems. Transfer factor values vary significantly between Th/Po and U/Ra/Pb, with Th and Po being 10 times lower. Bioavailability of ²³⁸U and ²³²Th is influenced by soil pH, soil texture with soil exchange capacity governing ²²⁶Ra uptake. Transfer factor values differ by 10–100 times between plant species and up to 10 times among different soil types [23, 31].

Transuranic elements (Am, Cm, Pu, Np) exhibit very complex soil chemistry because of various oxidation states, absence of stable carriers, and high tendency for complexation and hydrolysis. Transfer factor values for transuranic elements varied from about 1 to 10^–6. Concentrations of these radionuclides in fruits and grains are 10–1000 times lower than in the vegetative parts of plants. Accumulation of elements decreases in the order Np > Am > Cm > Pu. Among the higher actinides, Np appears to be the most environmentally mobile and plant available. Hydrolysis is a major factor influencing the behavior of Am and Cm in soils. The mobility of Pu depends on its form and decreases according to Pu (V) > Pu (VI) > Pu (III) > Pu (IV) [28].

Fission products (^89,90 Sr, ^134,137 Cs, ^129,131 I, ⁹¹Y, ⁹⁵Zr, ⁹⁵Nb, ^103,106 Ru, ^141,144 Ce, etc.) form a diverse group in terms of the environmental mobility. ⁹¹Y, ⁹⁵Zr, ⁹⁵Nb, ^103,106 Ru, ^141,144 Ce are poorly accumulated by agricultural plants because of their strong sorption in soil. Soil acidity and organic matter content influence the transfer of these radionuclides. Up to 99% of the plant uptake of radionuclides is retained in the roots, and very little transfer to above-ground parts occurs.

Most of the activation products (⁵¹Cr, ⁵⁴Mn, ^55,59 Fe, ⁶⁰ Co, ⁶⁵ Zn, ¹¹⁵ Cd) are radioisotopes of biologically important micronutrients and hence are highly mobile in soil–plant systems [29]. The behavior of these radionuclides depends strongly on the oxidation–reduction potential of the soil, the acidity of the soil solution, and the organic matter content. About 6.5% to 37% of Cd is known to be in mobile form followed by Zn and Cu in all soil types.

5.5 Soil Properties

Soil is an important component of terrestrial environment of biosphere and provides nutrients to most of the vegetation including some biota. Its critical functions include nutrient cycling, vegetative growth, exchange of gases, storage of carbon and contribute to the decomposition of organic matter, serving as a natural system for waste disposal. Soil comprises of a mixture of sand silt and clay in varying proportions, which give each soil type, distinctive characteristics and structure. Interrelation between physical, chemical, and biological properties of soils makes it an effective medium for plant growth. Nutrient availability in soil and its transformations are controlled by chemical properties of soil whereas soil structure, development, and productivity are controlled by biological properties [30].

It plays an intermediary role in the transport of radionuclides deposited from atmospheric releases or directly input into the soil to plants and animals and ultimately to human dependent on these plants and animals for food. All the food (crops, vegetables, fruits) consumed by humans are cultivated on land. The radionuclide further enters food chain via various pathways like soil to crop to humans, soil to grass to animal to milk to human, soil to grass to animals to meat, etc.

The mobility and bio-availability of radionuclides are mostly governed by the soil’s physico-chemical characteristics such as clay mineral type and content, organic matter content, complexation mechanisms of radionuclides, leaching, and sorption and desorption processes, synergistic and antagonistic effects of other elements, agroclimatic conditions, etc. in addition to the chemical forms of the released radionuclides. The effect of all these parameters can be encompassed and described by distribution coefficients (K_d). But most of these parameters display wide spatial variability, consequently, K_d employed in the estimation of transfer and uptake of radionuclide by plants or other biota has to be site-specific to minimize uncertainties.

Soil-to-plant transfer factor (TF), via plant root uptake of radionuclides to an edible part/compartment of a plant, is an important parameter in the dose assessment model. Generation of data on all types of edible plants, under different climatic conditions, soil type, soil remediation, radionuclides etc. is recommended for holistic assessment of impact and applications during an emergency especially, when data is not available for a site.

The impact of soil–plant interactions is crucial in understanding the behavior of radionuclides in the environment. The experiences gained from environmental radiological impact assessments following accidents like Chernobyl in 1986 and Fukushima in 2011 have significantly contributed to the scientific understanding of these interactions and had helped in the formulation of remediation strategies to reduce the radiocesium uptake from soil [4]. Transfer of radionuclides in farm crops exhibits considerable variation that can span up to three orders of magnitude due to variation in soil properties which affect the bioavailability of radionuclides to plants. Mineralogical and granulometric composition (sand, silt, clay), affecting water retention and nutrient availability, organic matter affecting the retention or release of radionuclides, acidity or alkalinity of soil, fertility of the soil and cation exchange capacity (CEC) are some of the important soil properties, affecting bioavailability, which are discussed below.

5.5.1 Soil Formation, Classification, and Soil Chemistry

As per USDA, NRCS, soil is a natural system consisting of minerals and organic matter as solids, liquids, and gases including oxygen and carbon dioxide, in the pore spaces, which occupy space. Soils characterized by distinct horizons or layers are different from its parent material from which it is formed due to the addition of organic material, loss of minerals through leaching, transfer of nutrients, and transformations of energy and matter. An important function of soil is its ability to support the growth of rooted plants in a natural environment [6].

Temporal properties of soil vary hourly, daily, and seasonally. It may be alternately cold, warm, dry, or moist. Biological activities in the soil are influenced by temperature and moisture fluctuations and extreme conditions of coldness and dryness may slow down or halt these activities. The soil receives additions of fresh, unrecompensed organic matter during specific periods, such as when leaves fall or grasses die. This input contributes to the organic content of the soil and influences nutrient cycling. Soil pH, soluble salts, organic matter content, carbon–nitrogen ratio, number of microorganisms, soil fauna along with temperature and moisture are some properties that are subjected to seasonal [6].

The parent material, derived from weathered mineral rock, is a critical determinant of soil properties. It contributes to both the chemical and physical characteristics of the soil. Temperature and precipitation, as climatic factors, play a significant role in soil formation. Biotic species, including plants and soil microbes, contribute to soil structure and composition by changing porosity of soil facilitating movement of air, water, and solutes through it. Plant roots enhance soil porosity, improve infiltration, and bring nutrients and moisture from deeper soil layers. Soil microbes are highlighted for their role in biological transformations and to drive the development of stable and labile pools of carbon (C), nitrogen(N), and other nutrients, influencing soil fertility and supporting plant communities [11]. Topographical features such as the slope aspect (e.g., south-facing vs. north-facing slopes), affects soil characteristics, for example south facing slopes receive more intense solar heat than north facing slopes which ultimately affect soil temperature and moisture conditions. The time period during which soil undergoes weathering processes is also a key factor [9]. Regional variations, between tropical and temperate environments, contribute to differences in soil characteristics, for example, in tropical zones with high rates of mineral weathering, clays may have low exchange capacity, thereby increasing the bioavailability of contaminants than temperate region soil. This understanding is significant to predict soil-to-plant transfer in different environments [2, 3].

5.5.2 Soil Profile

A soil profile is a vertical section of soil that reveals distinct horizontal layers called horizons (Fig. 3). Horizon boundaries, which separate different layers in the soil profile, are established based on several key measurements that include soil color, texture, structure, presence of rock fragments, and is a characteristic feature of topography of a given place. The uppermost layer the O horizon, primarily consists of organic material. Forested areas typically exhibit a distinct O horizon, while it may be absent in grasslands or cultivated fields. The A horizon is a mineral horizon formed at or just below the soil surface. It may include the accumulation of organic matter, with darkness attributed to the movement of organic material from the overlying O horizon. Common in forest soils, the E horizon (or eluvial layer) lacks clay, iron (Fe), or aluminum (Al) due to their removal by leaching, resulting in a lighter color compared to adjacent horizons. Located below O, A, and/or E horizons, the B horizon (zone of accumulation) receives deposits of illuviated materials that include clay particles, Fe and Al oxides, humus, carbonates, gypsum, and silicates leached from overlying horizons. The common presence of Fe and Al oxide coatings imparts a redder or darker color to the B horizon. C Horizon comprised of partially weathered parent material and represents a transition between soil and bedrock. The R horizon, or bedrock, is found beneath the C horizon. Bedrock depth varies based on geographic location, environmental conditions and can range from a few centimeters to 30 m from the soil surface. Soil profiles can differ, with some containing O, A, E, B, C, and R horizons, while others may only have C and R horizons [12].

Fig. 3

Adapted from USDA NRCS

Master horizons identified in a typical soil profile.

5.5.3 Soil Classification

Soil texture is classified based on the fractions of sand, silt, and clay present in a soil sample. Different countries may use various classification systems, often naming the soil based on the primary constituent particle size or a combination of sizes (e.g., sandy clay, silty clay). Loam soils are characterized by roughly equal proportions of sand, silt, and/or clay. The soil texture triangle is a graphical representation used to classify soil based on the distribution of sand, silt, and clay. It helps in visualizing the soil texture and understanding its composition. Table 1 also shows the soil particle sizes. Clay particles are the smallest and are classified as having diameters of less than 0.002 mm [13]. Silt particles have diameters between 0.002 mm and 0.05 mm as per USDA soil taxonomy. Sand particles are the largest, with diameters larger than 0.05 mm. They can be further classified as coarse, intermediate, medium, and fine. Texture has a profound impact on various properties of soil, including infiltration, structure, porosity, water holding capacity, and chemistry [6].

Table 1 USDA: United States Department of Agriculture WRB: The World Reference Base for Soil Resources (WRB). (Fourth edition 2022)

5.5.4 Soil Formation in Different Agroclimatic Zones

The process of soil formation in tropical environments is influenced by the climatic conditions, and higher temperatures and abundant rainfall contribute to the development of deep soils often nutrient-depleted due to the rapid leaching of nutrients. In some tropical areas, the surface soils may consist of hardened layers rich in iron and/or aluminum oxides, known as laterites. The clay mineral illite (2:1 clay type) may be absent in tropical environments because it is primarily formed at lower temperatures [14, 15]. Kaolinite or secondary mineral clays (such as goethite and gibbsite) are more common in tropical regions. These are 1:1 clay types and have lower cation binding abilities. Rapid decomposition of organic material on the soil surface leads to a quick recycling of nutrients and contaminants into the vegetation. This cycling is influenced by the warm and humid conditions that enhance microbial activity. In arid and semi-arid regions, soil formation is dominated by physical processes and the availability of water. Entry and exit of moisture at soil surface lead to incomplete leaching and the precipitation of secondary minerals. Fertile soils can still develop in these regions if there is sufficient water to reach base saturation at 100% to support plant growth, allowing for the accumulation of nutrients [16].

Grassland soils are strongly influenced by the prevailing climatic conditions where grassland vegetation predominates along with distinctive features of grassland ecosystem. Frequent soil-moisture deficits are common in grasslands, limiting the rate of mineral weathering. Due to frequent moisture deficits, grassland soils often experience secondary carbonate mineral accumulation in lower soil horizons. Grassland ecosystems exhibit a characteristic pattern where the relative abundance of below-ground biomass and active bioturbation (the mixing of soil by organisms) contribute to the development of thick, dark, organic matter-rich A horizons. In well-drained grassland soils, the organic matter content tends to increase with higher effective moisture levels and decreasing mean annual soil temperatures. Conversely, the depth to secondary carbonates decreases with decreasing effective moisture [17].

Arctic regions, characterized by glacial and Arctic deserts, soil-building processes are limited and occur only in rudimentary forms. The unique environmental conditions give rise to specific types of soils, particularly in areas with tundra vegetation. Soils in the Arctic are skeletal, meaning they have a high proportion of mineral components and are low in humus (organic matter). The subarctic north of Asia is characterized by a timber less zone with tundra vegetation. Tundra-type soils in these regions are specifically influenced by the subarctic climate and tundra vegetation. Tundra soils experience poor drainage, primarily due to permafrost (permanently frozen ground), which limits water movement through the soil. The short period available for organic substances to decompose, along with poor drainage, leads to the accumulation of undecomposed organic residues, often in the form of peat particles. Undecomposed organic residues and poor drainage results in the formation of peaty-gley soils in the tundra. Gley is a bluish substance formed in oxygen-deprived (anaerobic) conditions, characteristic of poorly drained soils. Tundra soils alternate with soils of the taiga (boreal forest), a cold, swampy, forested region. Soils below the frozen taiga are referred to as cryogenic soils, influenced by frost action. In mountainous regions, peaty-gley soils are replaced by mountain tundra soils, and the soils are often weakly developed, consisting of detritus and stony fragments [13].

5.5.5 Chemistry of Soil Affecting Radionuclide Migration

Soil chemistry is a critical factor influencing vegetative productivity, species composition, and is significantly influenced by several key factors such as soil pH and CEC, i.e., ability of soil to retain and supply cations to plant roots and depends on soil texture and organic matter content. Vertical migration in soil is influenced by various factors, including composition of the soil solution (pH, inorganic ion concentration, redox potential, organic substance concentration), physical and chemical soil properties (clay minerals, oxides, organic matter, surface characteristics, particle size distribution), temperature, microorganisms and fungi, including mycorrhiza. Speciation of most radionuclides in soil is in cationic form, except for iodine. The chemical species and concentration of other cations in the soil influence the transport, availability, and uptake of radionuclides. Other elements in the soil compete for binding sites with radionuclides. The presence of other cations can have either a synergistic or antagonistic effect on radionuclide uptake. Low pH values and low clay content, leading to lower CEC, can increase the mobility of radionuclides within the soil profile [18]. Soil texture, including particle size distribution, plays a role in migration rates. Migration rates of radionuclides decrease in the following order: Iodine (I) > Strontium (Sr) > Ruthenium (Ru), Cerium (Ce), Cobalt (Co) > Cesium (Cs) > Plutonium (Pu), Americium (Am).

The sorption of dissolved radionuclide ions to solid surfaces is an important process that affects their mobility and bioavailability in the environment. Models describing radionuclide sorption often rely on the concept of solid–liquid distribution coefficients (K_d values). K_d is the ratio of the concentration of radionuclide sorbed on a specified solid to the radionuclide concentration in a specified liquid phase at equilibrium (K_d, L/kg). The simplest K_d model is based on hypothesis that radionuclide sorption on the solid phase is in equilibrium with the radionuclide in solution and can exchange with it. Elapsed time since the incorporation of radionuclides can affect the quantification of K_d. A fraction of the incorporated radionuclide may become fixed by the solid phase over time, leading to an aging effect related to sorption dynamics. Labile or exchangeable K_d refers to the initial sorption process where the radionuclide is reversibly sorbed. This initial sorption is often due to ion-exchange mechanisms, and the sorbed radionuclide can be exchanged back into the solution. K_d can be predicted based on soil characteristics and the composition of the soil solution. Soil properties, including the availability of sorption sites and the presence of competitive ions, play a crucial role in predicting K_d values.

K_d values are categorized based on the organic matter content, and the percentages of sand and clay in the mineral matter. This grouping helps understand how different soil textures and organic matter influence the sorption behavior of radionuclides. For specific radionuclides (e.g., radiostrontium and radiocesium), the cofactor approach is employed, considering specific soil properties. Studies in podzolic and peat soils highlight that Ca⁺² and Mg⁺² ions in the soil solution and soil properties of CEC are key factors influencing the transport of strontium. Further, CEC and ratios of Ca⁺² and Mg⁺² ions are important parameters affecting the K_d of strontium. The cofactor approach includes considerations of soil pH for uranium and heavy metal radionuclides, with a focus on observations with a sufficiently large dataset. Speciation data and water regime are taken into account, particularly for radioiodine, where factors like the chemical forms of iodine and water conditions influence sorption.

The sorption of radiocesium in soils is influenced by ionic exchange processes occurring at two types of sites, frayed edge sites (FES) with high affinity and regular exchange sites (RES) with low affinity. These sites are associated with end of expanded clay layers and organic matter phases and external position in clay minerals respectively. The dominance of FES in radiocesium sorption is observed in soils, except for those with high organic matter content or low 2:1 phyllosilicate content. The irreversibility of Cs binding to micaceous minerals interlayer is noted due to interlayer collapse [17, 20, 21]. To estimate the soil's capacity to sorb Cs, the concept of Radiocesium Interception Potential (RIP) is employed [22, 23]. RIP is linked to the content and selectivity of expandable clays, particularly illite and other 2:1 phyllosilicates containing FES. RIP serves as a crucial parameter for predicting radiocesium interactions and short-term mobility after a radioactive release. Radiocesium sorption is further influenced by the NH4⁺ and K status in the solid and solution phases of soil. Potassium (K) is essential for plant growth and is often added to agricultural soil as fertilizer, resulting in high dissolved K concentrations. The availability of NH4⁺ in agricultural soils can be affected by nitrogen fertilizer application, redox potential, organic nitrogen decomposition, crop nitrogen uptake, nitrification, and the types of expandable clay minerals present [19]. K_d (Cs) values, which represent the ratio of radiocesium concentration in the solid phase to that in the liquid phase, increase with clay content and are lowest in organic soils. Time-dependent migration studies in sandy soil in Kuwait using lysimetric studies revealed that 90% of ⁸⁵Sr and ¹³⁴Cs were situated at 10.38 and 5.73 cm in the top soil layer, respectively. The average vertical migration rate of ⁸⁵Sr varied from 2.2 to 4.4 cm per year, while for ¹³⁴Cs, it ranged from 0.3 to 0.9 cm per year [24].

There is a strong dependence of K_d for uranium, particularly within the pH range of 5–7, where K_d (U) is observed to be 10 times higher. The sorption behavior (K_d Iodine) of iodine is more complex and involves factors such as soil redox potential, drying temperature, sorption contact time, organic matter content, water content, microbial activity, and oxidizing-reducing conditions.

6 Ecological Transport and Trophic Food Chain

Plants, the primary producers, uptake the radioactive elements from non-living portions of the environments like soil or air and put these substances into the ecological food chain. Besides direct consumption of food, there are other pathways the radionuclides can reach human. The plants or vegetation are consumed by herbivores or the primary consumer and the radionuclides along with other nutrients are transferred to animals. These are successively eaten by carnivorous animals and such a typical food chain may have three to five successive links or trophic levels [31]. Human beings are at the top of the trophic food chain. As radioactivity gets concentrated at each trophic level in the chain, the maximum activity will be present in human beings and hence can be considered as indicator organism for the presence of radioactivity in the environment or the ecosystem. This concept gave rise to the principle ‘If human is adequately protected, the environment, other biota and ecosystem is also protected’. Transfer of radionuclides between different compartments of the ecosystem and different trophic levels are described by ‘Concentration factor’ or ‘Bioaccumulation factor’.

The contamination of animals by radionuclides can be through skin absorption and inhalation, but the transfer of radionuclides through ingestion of contaminated feed and soil is considered the most significant pathway. Animals can ingest radionuclides present in vegetation and associated soil as part of their feed but radionuclides associated with soil may not be easily available for absorption. Animals can inhale radionuclides present in the air, but this pathway is often of lesser importance in terrestrial ecosystems. The transfer of radionuclides from soil to plants and subsequently to animals is expressed by transfer coefficients. Transfer coefficient quantifies and considers all the processes from the ingestion of a radionuclide to its incorporation into specific animal tissues. Transfer coefficients adopted for expressing transfer of radionuclide to milk (Fm, dL⁻¹, d kg⁻¹) and meat (Ff, dkg⁻¹), represent the equilibrium ratio of radionuclide activity concentration in milk/meat to the daily dietary radionuclide intake. Radionuclides transferred to animals can ultimately reach humans through the consumption of animal products such as milk and meat.

Quantifying the animal's diet and understanding its composition are important in estimating transfer coefficients, especially for agricultural animals. Several factors influence these coefficients, and variations occur based on feeding strategies (whether animals graze outdoors or are kept indoors), agricultural practices, dry matter digestibility, and diet composition, etc. Accurate assessments require site-specific data, considering variations in countries, seasons, specific vegetation (grass or bush), dietary constituents, farming system, etc. Transfer coefficients need to be in equilibrium with the dietary intake of radionuclides. Equilibrium is often rapidly reached for many radionuclides in milk but may vary for others. Temporal variations in an animal's radionuclide intake can result in changing tissue concentrations. Experimental observations may not cover a sufficiently long period for equilibrium, especially for radionuclides with long radiological and biological half-lives in tissues (plutonium). Dynamic models are employed to predict radionuclide activity concentrations in different tissues over time, considering continuous, single, or varying intakes [14].

The use of dry matter intake in calculating transfer coefficients can introduce anomalies, especially as dry matter intake tends to increase with the size of the animal. To address this issue, transfer coefficients are modified using concentration ratio (CR). CR is defined as the equilibrium ratio of the radionuclide activity concentration in the food product (measured on a fresh weight basis) divided by the radionuclide concentration in the feed (on a dry matter basis). CR provides an advantage in field studies as it eliminates the need to calculate or assume values for dietary dry matter intake. In cases where the animal's diet consists of multiple foodstuffs, knowledge of the relative proportions of all dietary components is essential to apply CR values accurately.

In terms of transfer to meat, gastrointestinal absorption in ruminants can be used for calculating the transfer coefficient. However, Technical Report Series No. 472 [14] proposes that the concentration ratio is a more robust and generic parameter compared to the transfer coefficient. For most nuclides, concentration ratio data show little variation between different species of ruminants such as sheep, goats, cattle, and horses. Hence, concentration ratios derived for one species of ruminant can be applied to other species within the same category. In many cases, stable elements are used to determine the concentration ratio. However, it's important to note that stable elements may be under homeostatic control, and their transfer may not be linear with the intake rate of radionuclides.

7 Transport Mechanism of Typical Radionuclides of Natural Origin

Radionuclide characteristics are deciding factor in their transport and uptake by plants. Their residence in the environment is determined by their radiological half-life and ecological half-life. Radiological assessments of specific radionuclides in the environment require prior knowledge of their behavior and transport patterns. This information helps in evaluating potential risks and implementing appropriate measures for radiation protection. Ecological half-life of a radionuclide is the time required for the removal of half of its concentration from the environmental media and depends on the characteristics of the environmental compartment as well as the radionuclide chemical properties. In this discussion, we will focus on radionuclides such as U, Th, Ra, Pb, and Po which are of natural origin. These radionuclides make significant contributions to the radiation dose received by humans and other organisms. The most abundant isotopes of Th and U are ²³²Th and ²³⁸U, respectively. They serve as the starting elements of the 4n and 4n + 2 natural radioactive series, which have extremely long half-lives. Many of their daughter products also have long half-lives and different decay modes.

7.1 Radium

All isotopes of Radium are radioactive, and within the three-natural series, these include ²²⁶Ra from the U-series, ²²⁸Ra and ²²⁴Ra from the Th series, and ²²⁵Ra from the actinium series. ²²⁶Ra and ²²⁸Ra are particularly noteworthy among the Radium isotopes because they have long half-lives for environmental mobility but considerably shorter than their parent. This characteristic results in a significant contribution to the radiation doses received by humans from natural sources. Human activities such as mineral extraction, phosphate industry, and other practices have led to an increase in the concentration of naturally occurring radioactive materials (NORM) in the accessible environment. Once these materials enter the food chain, they tend to accumulate in bones, similar to calcium, which is also an alkaline earth metal.

7.1.1 Radium in Soil

Radium naturally coexists with uranium and thorium in ores, and when undisturbed, it remains in equilibrium with its precursor thorium isotopes. However, natural processes can disrupt this equilibrium due to the distinct chemical and physical properties of these elements [32,33,34,35,36]. Shale rock, a type of sedimentary rock composed of clay and other minerals, tends to have higher concentrations of radium. Radium can also be found in bitumen slate, volcanic rocks, and phosphate rocks. When these rocks weather and break down into fine soil, the properties inherited from the parent rock are retained. In this soil form, radium becomes mobile in the environment. Within the soil matrix, radium behaves similarly to the alkaline earth elements calcium, strontium, and barium. Radium's distribution in the environment is influenced by the ion exchange properties of the soil depending on soil characteristics. Clay minerals show good affinity for Ra compared to other alkaline metals. Studies have shown that clay mineral Kaolinite can adsorb Ra at higher pH when calcium concentration is low [37]. However, a change in the surrounding chemistry with the presence of other competing ions can enhance the Ra desorption, and the exchange of radium is influenced by pH and competing ions [14, 31].

The high cation exchange capacity of Fe and Mn oxyhydroxides, as well as Mn oxide, significantly influences the cation migration in the environment. Manganese oxide, which occurs as fine aggregates or as sediment coating, in particular, strongly scavenges radium. Organic matter present in soil also exhibits a good cation exchange capacity and can absorb radium up to ten times more effectively compared to clay alone [38].

7.1.2 Radium in Groundwater

The interaction of groundwater with Ra-rich materials, such as rocks and minerals, can result in the release of radium into the aqueous medium. Several factors influence this process, including the acidity of the medium, the solubility of the material, and the concentrations of both the parent isotope (usually uranium or thorium) and radium present.

Radium is released to environment due to mining activity of minerals. Due to in situ leaching process, sometimes contamination level exceeds the prescribed limits [36, 39]. Groundwater tends to contain higher concentrations of uranium (U) and its decay products compared to thorium (Th) and its decay products due to the greater solubility of uranium [37, 40]. But quartzose sandstones with high total dissolved solids (TDS) may exhibit elevated concentrations of ²²⁴Ra and ²²⁸Ra in the aqueous phase due to the limited adsorption sites available on the aquifer solids [41].

7.1.3 Radium in Freshwater

Radium can be found in freshwater from various sources, including groundwater, the re-suspension and re-solubilization of radionuclides bound to sediment, and its deposition through air via precipitation and particle deposition. In rivers, lakes, and similar water bodies, concentrations of ²²⁶Ra typically fall within a narrow range, although there may be exceptions. Limited available data suggests that ²²⁸Ra also exhibits a similar range of concentrations [42]. High concentrations of ²²⁶Ra in surface water can be attributed to the presence of uranium-bearing minerals in the surrounding area [43]. Exposure of sedimentary rocks containing sulfide minerals to air results in the generation of sulfuric acid. This acidic environment facilitates, the dissolution and leaching of radionuclides, as their solubility tends to be higher under low pH conditions [44, 45].

Human activities can also contribute to the transfer of radium from groundwater to surface water and then to sediment. ²²⁶Ra can be introduced into surface water through the discharge of wastewater from coal mines, where it is predominantly present as Ra²⁺ ions. In surface water, Ra²⁺ can react with sulfate ions (SO₄²⁻) to form radium sulfate (RaSO₄), which then co-precipitates with other alkaline metal sulphates and accumulates at the bottom sediment [46].

7.1.4 Radium in Air

The concentration of radium in the air can vary depending on the climate, and it tends to be higher in dry regions due to lower levels of uranium (U) in sea air of coastal locations compared to industrialized areas [46]. Radium in the air is predominantly associated with re-suspended soil particles. In most areas, the observed radium concentration in the air is relatively low, often reported as low as 1.5 μBq/m³ which agrees with UNSCEAR data [47].

Coal fly ash is another source of radium in the atmosphere, which can subsequently deposit onto the ground, vegetation, and surface water. The concentration of radium in fly ash, a by-product of coal combustion, is higher than in coal itself. Mean values for ²²⁶Ra concentration in escaping fly ash range from 44.3 to 2400 Bq/kg [44].

7.1.5 Radium in Plants

Radionuclides can enter the plant system primarily through two pathways: root uptake from soil and foliar uptake from interception. The dominance of each pathway depends on factors such as the type of plant, the concentration of radionuclides in the soil and atmosphere. When compared to radium (Ra) with its large ionic radius, plants readily take up calcium (Ca) and magnesium (Mg), which are alkaline earth metals essential for plant growth [48]. Ra uptake is also influenced by soil exchange capacity since it tends to form strong complexes with soil, reducing its availability to plants [49]. Various factors can decrease Ra uptake in plants, including higher concentrations of Ca, elevated pH levels, and increased soil sulfate content [49, 50]. Studies have indicated that soils containing 14–50% organic matter, effectively complex Ra, more than its parent element Uranium. This is evident from the activity ratios of ²²⁶Ra/²³⁸U, which can increase up to nine from their isotopic equilibrium value of one [42]. Furthermore, these studies suggest that uranium ions are more prone to leaching and oxidation by percolating soil water, while Ra remains relatively immobile under such conditions [42].

7.2 Uranium

Uranium, a radioisotope of primordial origin having long half-life and exists in significant quantities, is contributing to the natural background radiation. Its distribution across the planet is uneven, with varying concentrations in different substances such as soil, rocks, water, and air. Uranium can be found in various geological environments, including igneous, metamorphic, and sedimentary settings. In the presence of oxygen, uranium is highly soluble and readily attaches to mineral surfaces, although to a lesser degree than thorium and radium. It can also form complexes with organic compounds.

7.2.1 Uranium in Soil

The average concentration of ²³⁸U in the continental crust is estimated to be 32.9 Bq kg⁻¹ [51]. However, in areas with normal natural radiation levels, the concentration of ²³⁸U in soils can vary widely, ranging from 0.4 to 20 mg/kg [52,53,54]. This variability can be attributed to factors such as the diversity of soil types studied, the heterogeneity of the soils themselves, and the varying concentrations of uranium in the rocks from which the soils originate. These factors contribute to the accumulation of uranium in the topsoil. In certain regions of the world, however, radiation levels in soils or waters greatly exceed the normal range of naturally occurring radionuclides.

Black shale and phosphate rocks are known for their high concentrations of uranium. In black shale, which primarily consists of clay and organic matter, uranium exists in complex forms. Conversely, in phosphate rock, which forms through sedimentary processes, uranium is found in mineral deposits. The concentrations of uranium in soil are strongly influenced by the uranium content present in the parent rock. The composition of parent rocks is influenced by various environmental factors that ultimately shape the formation of soil.

7.2.2 Uranium in Water

The process of weathering and groundwater interaction facilitates the release of uranium from its mineral sources into the surrounding environment. Groundwater has the ability to dissolve uranium from sandstone and permeable sedimentary rocks, but the presence of layered structures and fractures creates pathways for groundwater to penetrate less permeable rock formations [55].

Uranium exists in various minerals with different oxidation states [56]. Minerals such as uraninite (UO₂^+x) and coffinite (USiO⁴·nH₂O) contain uranium in its +4 oxidation state, which is relatively less mobile. However, when these minerals come into contact with an oxidizing environment, the uranium undergoes a conversion to its highly mobile +6 oxidation state. U⁺⁶ ions always combine with two oxygen atoms to form UO₂²⁺, which is a stable species in an aqueous medium. This UO₂²⁺ species serves as the foundation for several aqueous uranyl species and uranyl minerals [57].

The chemistry of groundwater plays a crucial role in the creation of uranyl minerals. When groundwater comes into contact with a primary ore body such as uraninite, a process called in situ oxidation occurs, converting +4 uranium to its +6 form and transforming its chemical composition into a uranyl silicate mineral [58]. This transformation leads to the mobilization of uranium, making it more soluble and allowing it to be transported away from the primary ore body, primarily in the form of uranyl carbonate complexes. In cases where the groundwater contains sufficient dissolved phosphate and magnesium [59], a secondary ore body composed of uranium (VI) phosphate can form, known as saleeite (Mg(UO₂PO₄)₂·10H₂O). Additionally, iron oxides scavenge dissolved UO₂²⁺, copper, and phosphate, resulting in the formation of metatorbernite (Cu(UO₂)²(PO₄)²·(H₂O)8) [60].

At low pH, uranium is present as UO₂²⁺ and at higher pH hydrolysis species like UO₂OH + , (UO₂)3(OH)⁵⁺ and UO₂(OH)20 are formed. With increasing pH, uranyl forms carbonate complexes [61] and is commonly present in soil and groundwater environments. If Ca ions are present, a neutral species, Ca₂UO₂(CO₃)³ is formed in both, natural and seawater.

7.2.3 Uranium in Plants

Uranium exhibits strong binding affinity to cell walls by precipitation, which can impede its absorption by roots and hinder its translocation within plants [62]. On the other hand, researchers also noted that the presence of organic complexes, such as carbonate or citrate, reduces uranium retention in roots while increasing its translocation to the shoots. Conversely, uranium complexation with phosphate was found to decrease uranium accumulation in all plant tissues by retaining it on the root epidermal cells.

7.3 Thorium

In nature, thorium (Th) consists of four isotopes: ²²⁸Th, ²³⁰Th, ²³²Th, and ²³⁴Th. Among these isotopes, ²²⁸Th is produced through the radioactive decay of naturally occurring ²³²Th, while both ²³⁴Th and ²³⁰Th are decay products of natural ²³⁸U.

7.3.1 Thorium in Soil and Rocks

The redistribution of uranium and thorium within soil is a complex phenomenon due to their radioactive decay, leading to the formation of daughter products where the interchange of both elements occurs [63]. Thorium being present as an insoluble salt is abundant in rock and soil compared to sea water. More than 95% of world total thorium resources are attached to four major deposit types, carbonatites, veins, alkaline/peralkaline rocks and placers [64]. In monazite mineral [(Ce, LaNdThU) PO₄], it is present as phosphate with an average of 6–7%. A good percentage of it is found as heavy mineral sand deposit on the south and east coasts of India [65]. Other minerals are thorite (ThSiO₄), thorianite (ThO₂), and bastnaesite [(Ce, La)CO₃F] [66].

Thorium (Th) due to its high insolubility tends to strongly adsorb onto mineral surfaces. It also demonstrates a strong affinity for humic acids and other organic ligands. As a result, thorium can become concentrated in organic deposits or be transported in the form of organic colloids [67]. In natural conditions, thorium exists as tetravalent ions. Unlike Uranium, under geological conditions, thorium is unable to undergo oxidation to a hexavalent state and form an analog of the uranyl ion [68]. The typical concentration range of naturally-occurring thorium in soil is 2–12 μg/g, with an average value of 6 μg/g [69].

Human activities contribute to thorium contamination in soil through various sources such as mining, milling, processing operations and the production of uranium, thorium, tin, and phosphate fertilizers. There are two primary processes associated with these industries that can lead to soil contamination: the precipitation of airborne dust particles and the disposal of waste materials containing uranium or thorium on land.

7.3.2 Thorium in Water

In water, Th can be easily hydrated and precipitated. If adsorbed on colloidal particles, it can get transported also [62]. Thorium migration is influenced by the pH conditions of water. In slightly acidic and neutral water, the dominant Th species is negatively charged complex (ThCO₃(OH)₃)⁻. At pH 5.9–6.2, positively charge thorium hydroxo complexes ((Th (OH)₃)⁺ and (Th(OH)₂)²⁺) are formed. Thorium accumulation in water is more in reducing environment than in oxidizing environment [62].

7.3.3 Thorium in Plants

The uptake of thorium by plants is influenced by various factors, including the type of plant species, the characteristics of the soil, and the thorium content present in the soil [70]. Thorium uptake by plants has reduced significantly in the presence of phosphate and nitrogenous fertilizers as it will form insoluble phosphates and nitrates [70]. The concentration of thorium in plants can vary depending on the specific plant part, but it consistently remains lower compared to the thorium concentration in the soil. Specifically, the roots of plants tend to have significantly higher thorium concentrations compared to leaves and seeds [63]. When wheat seedlings were subjected to thorium nitrate solution treatment, increase in thorium concentration was observed both the roots and seeds. However, the concentration of thorium in the leaves remained relatively low. Furthermore, the presence of thorium-enriched medium had a significant impact on the uptake of calcium by the seedlings, resulting in a decrease in calcium levels across all plant parts [63].

7.3.4 Thorium in Air

The transportation of thorium to the atmosphere can occur through mining, milling, and processing activities involving thorium and uranium ores. Due to its smaller aerodynamic diameter compared to ²³⁰Th and ²³²Th, ²²⁸Th is capable of traveling greater distances [72]. While there is currently no strong evidence available, it is assumed that the most probable form of thorium in the atmosphere is thorium dioxide (ThO₂) [71]. Thorium will be transported from the atmosphere to soil and water by wet and dry depositions, like other particulate matter in the atmosphere.

7.4 Polonium and Lead in the Terrestrial Environment

²¹⁰Po an alpha emitter is widespread in the terrestrial environment. The presence in deep soils and minerals can be traced to the decay of radionuclides in the ²³⁸U decay chain. Atmospheric fallout of ²¹⁰Po is generally assumed to be constant at a specific site over timescales of a year or more. However, the ²¹⁰Po flux may vary spatially by an order of magnitude, influenced by factors like rainfall and geographical location. Reported atmospheric residence time of ²¹⁰Po varies between 15 and 75 days. The mean atmospheric residence time is in the order of 26 ± 3 days. Airborne particles carrying attached ²¹⁰Pb and ²¹⁰Po return to the earth's surface through fallout. This leads to the deposition and accumulation of the final long-lived ²¹⁰Pb (22.3 y), which decays through various stages (²¹⁰ Bi, 5d; ²¹⁰ Po 140d) ultimately forming stable Pb in plants or topsoil. ²¹⁰Po in soils may originate either from the radioactive decay of radionuclides in the ²³⁸U series present in the soil (supported) or from the precipitation of radon decay products from the atmosphere( unsupported). Levels of ²¹⁰Pb and ²¹⁰Po in the top layer of soil can be correlated with the amount of atmospheric precipitation. In soils, ²¹⁰Po is in equilibrium with ²¹⁰Pb, indicating that ²¹⁰Pb in the soil is the primary source of ²¹⁰Po irreversibly adsorbed on clay and organic colloids in the soil.

Vertical distribution of ²²⁶Ra and ²¹⁰Po in cultivated soils of Buyuk Menderes Basin, Turkey had been investigated [73]. Concentrations of ²²⁶Ra ranged from 80 to 1170 Bq kg⁻¹ and ²¹⁰Po ranged from 10 to 870 Bq kg⁻¹, respectively in soil cores with depth. Study reveals that concentrations of ²²⁶Ra and ²¹⁰Po in different soil strata do not show significant variations with depth.

²¹⁰Pb and ²¹⁰Po contaminate vegetation through direct deposition. In plants, ²¹⁰Po is the predominant natural radioisotope, primarily resulting from direct deposition of daughters of ²²²Rn from atmospheric precipitation. Plants absorb radioactive nuclides from the soil (supported Po) and through direct deposition as radioactive fallout (unsupported Po) depositing on plants. The presence of ²¹⁰Po in soil adds uncertainty when using plants with root systems as monitors for ²¹⁰Po deposition study. Understanding the contribution of soil-derived ²¹⁰Po to plant contamination is essential for accurate monitoring. High concentrations of ²¹⁰Po and ²¹⁰Pb are found in tobacco and tobacco products. Plant roots absorb ²¹⁰Po and ²¹⁰Pb from the soil, with a minor contribution from fallout and hence soil is identified as primary source for ²¹⁰Pb, and consequently ²¹⁰Po, in tobacco.

Lichens are slow-growing perennials that contain significantly higher concentrations of ²¹⁰Po and ²¹⁰Pb compared to vascular plants and fungi. The ²¹⁰Po/²¹⁰Pb activity ratio in lichens is typically equal to one. This ratio signifies that ²¹⁰Po is approaching secular equilibrium with ²¹⁰Pb. Study of reindeer meat samples from Finnish Lapland showed seasonal variations in ²¹⁰Po activity concentrations (about 3 Bq kg⁻¹ wet weights (w.w.) in autumn, 5 Bq kg⁻¹ w.w in winters and 12 Bq kg⁻¹ w.w during spring). Further study reveals that annual average activity concentrations of ²¹⁰Pb in reindeer meat samples are 0.22 ± 0.04 Bq kg⁻¹ w.w., 10 times lower than ²¹⁰ Po and with less seasonal fluctuations.

8 Transport Mechanism of Radionuclides of Anthropogenic Origin

8.1 Cesium

Naturally occurring cesium is found in soil and water as a stable isotope, originating from soil minerals. However, radioactive cesium is primarily generated through three main sources: discharges from nuclear facilities, global fallout resulting from nuclear weapons testing, and fallout following nuclear accidents.

8.1.1 Cesium in Soil

Cesium, the first element in its group and sharing similar properties with Sodium and Potassium, is typically found in soil at very low levels. While Potassium is essential for plant growth and both Sodium and Potassium are important for human health, cesium levels in the environment can increase significantly due to nuclear accidents, particularly through radioactive isotopes like ¹³⁴Cs and ¹³⁷Cs, which are beta and gamma emitters. This radiation can pose risks to living organisms through both external and internal exposure. Hence, it is crucial to monitor the entry of cesium into the food chain.

Cesium can bind to soil through mechanisms like sorption, ion exchange, or complex formation, depending on the type of organic matter present. Smaller soil particles exhibit more pronounced sorption due to their larger surface area [74]. Once cesium becomes complexed with soil particles, it becomes difficult for it to exchange with other ions, resulting in limited biological uptake and environmental transport [71, 75, 76, 77]. Once cesium is introduced into the soil, it can be taken up by plants and enter the food chain.

8.1.2 Cesium in Plant

Plants have the capability to uptake both potassium and cesium, as they belong to the same group. However, the efficiency of cesium uptake by plants is lower compared to potassium. This is primarily due to cesium's higher atomic weight and larger ionic radius.

The presence of potassium and other ions in the soil can affect the uptake of cesium by plants. Higher concentrations of potassium tend to reduce the uptake of cesium, while lower concentrations of potassium can increase cesium uptake. Radiocesium can be found not only in soil but also in water and air, particularly after nuclear accidents. During the initial fallout period, airborne cesium can enter plants through their leaf surfaces, while cesium in water can be absorbed by the entire exposed surface of the plant. After the Chernobyl accident, a significant amount of cesium was intercepted by plants, making it the primary contamination route for terrestrial plants [78]. The uptake of cesium by plants is influenced by the morphology and physiology of the plant, leading to variations in uptake levels among different plant species growing on the same soil. Factors such as cesium speciation, increased organic matter content in the soil, temperature affecting microbial activity, and soil moisture content also play significant roles in determining plant uptake. These factors collectively contribute to the variability observed in cesium uptake among different plant species.

8.1.3 Cesium in Air

During nuclear accidents, cesium is released as a volatile cloud. As the cloud cools down, cesium isotopes become less volatile and can be found in two forms: condensed on particles present in air as well as fuel particles in the size range of 20–50 μm [65]. These particles deposit on soil and plant surfaces. Due to their size and weight, larger particles of cesium tend to settle closer to the site of the nuclear accident. On the other hand, smaller particles are lighter and more easily transported by air currents. As a result, they can be carried over longer distances, potentially reaching areas located farther away from the source of the accident.

8.2 Strontium

Strontium is a naturally occurring element having four stable isotopes ⁸⁴Sr, ⁸⁶Sr, ⁸⁷Sr, and ⁸⁸Sr. Its radioactive isotopes ⁸⁹Sr, ⁹⁰Sr are of anthropogenic origin and are high yield fission products. Due to atmospheric nuclear tests, accidents, and nuclear facilities ⁹⁰Sr and ⁸⁹Sr are released into the environment since middle of twentieth century. Being a second group element, it behaves like calcium and after entering into the biological system gets deposited on bone.

8.2.1 Strontium in Soil

The behavior of strontium in soil is influenced by various factors such as the physicochemical properties of both the soil and strontium, soil type, agricultural practices, climate, and hydrological conditions [79]. Strontium exists in both soluble and insoluble salt forms. It primarily attaches to the soil through ion exchange processes. Strontium tends to accumulate in the humus-rich horizon where organic matter content is high. Radiostrontium forms humates, a complex with humic acid from humus. Interestingly, increased soil acidity can also enhance the migration of strontium[80]. Studies conducted after the Chernobyl accident revealed that a significant portion of the radioactive isotope ⁹⁰Sr remained concentrated in the upper layer of soil (approximately 10–20 cm) even after a span of 30 years [81].

In saline soil, strontium is present as less soluble strontium sulfate and hence it is less mobile.

Increased calcium content has enhanced the strontium migration in the soil. It was found that the effective half-life of ⁹⁰Sr in 0–5 cm layer of fallow automorphous sod-podzolic soils is 14.3–15.0 years. But due to vertical migration, in sod-podzolic clayey sandy-loam soil, the effective half-life of radionuclides is reduced to 10.5 years [82].

8.2.2 Strontium in Plant

The accumulation of strontium in plants differs significantly depending on the type of plant even within classes, families, and species. For same radionuclide, inventory on the soil dry biomass content of radio strontium of individual crops differs significantly. Legumes and root crops have been observed to have the highest levels of strontium accumulation, while cereals exhibit significantly lower uptake. Strontium behaves similarly to calcium in terms of its absorption and distribution within plant system. However, research has demonstrated that the introduction of calcium into soil can hinder the uptake of ⁹⁰Sr by plants. When plants were irrigated with solutions containing calcium thiosulfate and calcium nitrate, there was a notable decrease in the activity of strontium compared to situations where no calcium was present. According to Kabata and Mukherjee, higher solubility of Strontium bicarbonate to calcium bicarbonate makes strontium more mobile in soil than calcium [80]. Strontium is characterized by a significant presence of readily available forms that can be easily accessed by plants. These accessible forms make up approximately 53–87% of the total content of the radionuclide and tend to increase over time. At neutral pH, strontium uptake is maximum. Strontium accumulation observed in descending order in different plant species in different studies are as follows:

• (a) lichens > mosses > ferns > gymnosperms > angiosperms.

• (b) root-crops > beans > potatoes > groats > cereals and vegetable crops.

• (c) clover > lupine > pea > perennial grasses in floodplain lands > perennial cereal–bean mixtures > vetch > spring rape > pea–oat and vetch–oat mixtures > herbs in natural hayfields > grasses on drained lands > grasses on arable lands > corn.

8.2.3 Strontium in Air

Limited research has been conducted on the distribution of stable strontium in the air. The reported median strontium concentrations in remote regions around the world are approximately 0.81 ng m^–3. However, in polluted areas, these concentrations can increase significantly, reaching up to 50 ng m^–3 [83].

During thermal processes, strontium is emitted into the atmosphere in the form of strontium oxide (SrO). SrO is not stable and readily reacts with moisture or carbon dioxide present in the air. These reactions result in the formation of strontium hydroxide (Sr[OH]₂) or strontium carbonate (SrCO₃), respectively. When Sr (OH)₂ comes into contact with water in clouds or is washed out by rain, it undergoes ionization, leading to the formation of Sr²⁺ and SrOH⁻ ions. However, there is currently no evidence in the literature regarding the interaction of SrO with other compounds in the atmosphere [84].

8.3 Radioiodine

8.3.1 Soil to Plant

Geochemistry of iodine is governed by its volatile nature. Major component of global cycle of iodine is from volatilization of organo-iodine compounds and elemental iodine from both biological and non-biological sources in oceans. The dominant species in the aerobic soil environment include I⁻, IO⁻, and I. Stable ¹²⁷I is typically present in soils at an average concentration of 5 mg/kg. Terrestrial plants and food crops contain stable I (¹²⁷I) in the range of 0.07–10 mg/kg dry weight. Another natural isotope, ¹²⁹I, is less abundant and can be released during certain nuclear activities, but it has a lower radiological impact compared to ¹³¹I. Radioiodine dissolves in water and easily moves from the atmosphere to different environmental components. It readily absorbs to various soil components, including organic matter and soil minerals. This absorption limits the uptake of iodine through the plant root system. Naturally occurring isotopes, ¹²⁷I and ¹²⁹I, usually behave similarly in the environment. However, their soil-to-plant uptake rates may differ in some soils [3].

Soil-to-plant transfer for short-lived radioiodine isotopes, especially ¹³¹I, is generally thought to be of negligible importance because of the short physical half-life. In the emergency and transition phases, interception by plants of short-lived radionuclide ¹³¹I becomes significant. However, accumulation of iodine in plants becomes relevant for the long-lived isotope ¹²⁹I only. Transfer of radioiodine from soil to plants in the emergency phase has received limited attention in research and radiation protection. Few compiled data exist for iodine transfer to plants, especially for transfer factor values, which range from 0.1 to 5.0 for vegetative plant mass. TRS 472 does not mention such value for transfer of soil to grass [14]. Transfer factor values for iodine are low in soils with high cation exchange capacity and organic matter content. For grains such as rye and wheat (components of animal diets), iodine concentration ratio (CR) values vary from 5 × 10⁻⁴ to 8 × 10⁻³

8.3.2 Plant to Animal to Milk to Human

Raw milk is prone to radioiodine contamination as livestock, particularly grazing animals like dairy cows, goats, and sheep, feed on grass contaminated with deposited radioiodine. Radioiodine isotopes intercepted by pasture vegetation and ingested by grazing animals are quickly and completely absorbed through the gut. Different physico-chemical forms of iodine do not alter the extent of true absorption, which remains consistently complete (Fa = 1) [85, 86, 87]. Enhanced intake of stable iodine does not result in a reduction in gut absorption of radioiodine isotopes. Iodine is rapidly absorbed into the blood plasma, circulating as iodide and subsequently accumulating in the thyroid. Radioiodine is transferred into the mammary gland and is excreted via milk and urine.

Thyroid gland has a high affinity to concentrate iodine, magnifying the hazard caused by accumulation of ¹³¹I similarly to stable iodine. ¹³¹I accumulates in the thyroid and rapidly transfers into the milk within 30 min of introduction into the body [85]. Peak radioiodine activity concentrations are reached in 6–12 h after introduction into the body. Radioactive iodine can also be absorbed via the lungs into the plasma. Goat’s milk and sheep’s milk contain approximately ten-fold higher radioiodine activity concentrations than cow’s milk. For cows, the milk/plasma ratio has been reported as 0.6–5.5, whereas, for sheep and goats, it was 2–24 [85].

A controlled feeding experiment was conducted using herbage recently contaminated by fallout from the Chernobyl accident. The transfer coefficient of ¹³¹I to sheep milk was determined to be 0.3 ± 0.017 d/L [87]. These values are in agreement with reported value of 0.03 to 0.9 d/L, (geometric mean of 0.23 d/L) in TRS 472 for iodine in sheep milk [14]. Values reported for stable iodine in dairy cows after the Chernobyl accident are in the range of 0.015–0.020 d/L. The daily proportion of ¹³¹I intake secreted in sheep milk was found to be 5.6 ± 0.035%, which is an order of magnitude higher than for cattle. This higher transfer agrees with the higher transfer of stable iodine from plasma to milk observed in sheep and goats. The lactation phase does not seem to have a significant effect on iodine transfer to milk in sheep.

In controlled experiments, researchers investigated the effect of stable iodine intake on the transfer of radioiodine to milk in dairy animals. Different levels of stable iodine intake were considered: low, moderate, and high. Mean Fm value for oral radioiodine to milk was reported as 0.020 d/L for a low stable iodine intake. The mean Fm value increased to 0.024 d/L for a moderate stable iodine intake rate. However, a significant decrease in the transfer to milk was observed for a high stable dietary iodine intake rate, with a mean Fm of 0.018 d/L compared to the moderate treatment. The observed differences in Fm values for the three stable iodine treatments were attributed to the differential affinities and saturation levels of the thyroid and milk pathways, which compete for the available iodine. The competition for available iodine between the thyroid and milk pathways influenced the transfer of radioiodine to milk.

Modeling studies confirmed that stable iodine intake influences the partitioning of iodine among the thyroid, milk, and excreta in dairy animals [86]. The model was utilized to predict the effects of varying stable iodine intake on the chemical contamination of milk by stable iodine. Predicted outcomes suggested a significant variation in the time taken for radioiodine to reach peak concentrations in milk after a deposition event across different stable iodine intake levels. Administering low amounts of stable iodine, specifically less than 100 mg/d, to dairy animals was predicted to increase the transfer coefficient (Fm). However, higher stable iodine intake levels, exceeding 150 mg/d, were anticipated to reduce the transfer of radioiodine to milk.

8.4 Cobalt

⁶⁰Co and ⁵⁸Co are activation isotopes that are generated during the routine operation of nuclear reactors. Neutrons interacting with structural materials within the reactor lead to the production of these radionuclides. Small quantities of ⁶⁰Co and ⁵⁸Co may be released into the environment as liquid discharges from nuclear reactors.

8.4.1 Environmental Behavior and Mobility

Radioactive isotopes of cobalt, including ⁶⁰Co and ⁵⁸Co, will exhibit chemical behavior similar to stable cobalt in the environment. However, ⁶⁰Co and ⁵⁸Co undergo radioactive decay with half-lives of 5.3 years and 71 days and monitoring for these isotopes is restricted by their decay characteristics [88, 89].

Cobalt distribution in soils is influenced by clay content, presence and distribution of iron and manganese oxides [90, 91]. In clay-rich soils, cobalt adsorption may occur through ion exchange at cationic sites on clay. Adsorption can involve simple ionic cobalt (Co) or hydrolyzed ionic species like CoOH²⁺ [91, 92]. The presence of organic matter, moisture content, and complexing ions influences cobalt bioavailability whereas redox potential (Eh) and pH conditions impact cobalt mobility and uptake. The content of clay minerals plays an important role in cobalt translocation via root systems and foliage into plants. In the terrestrial environment, pH is a dominant parameter regulating cobalt availability. At lower pH, simple divalent cobalt dominates in most situations. Adsorption of Co²⁺ on soil colloids is high between pH 6 and 7. At lower pH, cobalt is absorbed less strongly enhancing leaching and uptake by plants [93, 82]. The partition between solid and solution phases, indicated by the distribution coefficient (K_d) at equilibrium, serves as an indicator of potential bioavailability.

8.4.2 Uptake

Plants can accumulate small amounts of cobalt from the soil, particularly in plant parts regularly consumed, such as fruits, grains, and seeds [93]. Cobalt availability to plants is influenced by the presence of manganese dioxide (MnO₂) in soil. Cobalt is strongly sorbed and co-precipitated with MnO₂. Soil pH is a major factor controlling the availability of cobalt to plants [94]. More acidic soils may sorb cobalt less strongly, potentially enhancing plant uptake. The presence of humus and high concentrations of manganese in soil can limit cobalt uptake by plants. Cobalt is taken up by plants as a divalent cation, the common oxidation state found in soil minerals. The differences in the uptake pattern of cobalt by plants, especially hyper accumulators like Alyssum, from different soils and pH may be related to variations in organic matter and iron contents of the soils [94]. Soil water status significantly influences the amount of cobalt available for plant uptake. In poorly drained soils, the extractable cobalt amount may be greater compared to well-drained areas, leading to increased plant uptake [94, 95].

Uptake of cobalt by plants depends on its concentration in the ionic form in the soil solution and the concentration on the exchange sites of cation-exchangeable soil surfaces. The highest accumulation of cobalt in plants typically occurs in the roots. Other parts of the plant also accumulate cobalt to varying degrees. The accumulation of cobalt in plant roots can lead to potential biomagnifications, resulting in concentrations in other plant-eating organisms like predatory birds or mammals. High cobalt concentrations in these secondary consumers may cause toxic effects, a phenomenon known as secondary poisoning. Variation in bioconcentration factors for ⁶⁰Co and stable Co fertilizer for different tropical crops had been demonstrated by spiking experiments. These factors may vary between 0.087 and 1.03 in different crops and were found to diminish in crops over subsequent years.

9 Conclusion

In conclusion, the terrestrial environment has long been marked by the influence of radioactivity, arising from natural radionuclides like ²³⁸U and ²³²Th, alongside artificial counterparts such as ¹³⁷Cs and ⁹⁰Sr. As a crucial source of human sustenance, the majority of consumed food originates in this environment, exposing both humans and non-human biota to radiation through diverse pathways. Soil, with its distinct textures and properties derived from the interplay of sand, silt, and clay, plays an important role in nurturing plants. Physicochemical processes, primarily governed by the clay fraction, dictate the nourishment plants receive. Agroclimatic zones, shaped by precipitation and temperature, influence soil formation. Assessing the terrestrial impact necessitates consideration of pathways like foliar interception, translocation, and root zone uptake by plants. Radiological Environment Impact Assessment (REIA) models depend on crucial parameters like soil-to-plant transfer factor (TF) or concentration ratio (CR). TF is markedly influenced by soil characteristics, agricultural practices, crop types, rainfall frequency, and dietary habits. Studies also highlight the impact of ionic exchange processes at frayed edge sites (FES) in influencing radiocesium sorption in soils. In unveiling the dynamics of radioactivity in the terrestrial realm, this chapter sheds light on the different factors shaping its environmental impact.