Evaluation of Long-Term Environmental Impact and Radiological Risks at a Former Thorium and Rare Earth Site in North-Eastern Kazakhstan

Abstract

Kazakhstan holds the global leadership position in natural uranium mining. Nonetheless, the extraction and processing of radioactive ores has the potential to induce instances of radiological contamination. This study aimed to evaluate the radiological soil contamination at a former monazite, tin, and radioactive ore processing facility located in Ust-Kamenogorsk city. Pedestrian gamma–ray measurements revealed dose rates up to 1.00 µSv/h, significantly exceeding the natural background (0.16–0.18 µSv/h). The analysis of the 28 soil profiles demonstrated that deeper soil layers (below 60 cm) were significantly contaminated with radionuclides constituting production waste. Furthermore, the total activity in the superficial soil layer is in the range of 583–5275 Bq/kg (alpha emitters) and 641–1749 Bq/kg (beta radionuclides). The maximum of total radioactivity in the samples collected at the 80–100 cm layer was at the level of 22,482 Bq/kg (α-emitters) and 6845 Bq/kg for gross beta radiation. In consideration of the site’s proximity to public buildings, the calculated radiological hazard indices were calculated, revealing the potential danger for human health. The elevated excess lifetime cancer risk and annual gonadal dose equivalent obtained for the topsoil layer indicate a high level of radiological risk to the local population. The obtained results emphasise the necessity of developing rehabilitation strategies and long-term monitoring of the contaminated site, which is consistent with the global objectives of sustainable development in the field of environmental protection and public health.

1. Introduction

The legacy of industrial activities involving naturally occurring radioactive materials (NORM) presents significant environmental and public health challenges, particularly in regions with a history of rare earth mineral processing. In north-eastern Kazakhstan, one such site, formerly dedicated to the extraction and treatment of thorium ore and rare earth elements, has become the subject of increasing concern due to its potential radioactive contamination. This site not only poses a localised hazard but also serves as a case study in the broader field of radioecology, where understanding the environmental behaviour and migration of radionuclides is essential. Conducting a comprehensive radiological risk assessment is crucial to evaluate the extent of contamination, potential exposure pathways, and the associated health risks for both the environment and nearby populations. This study aims to integrate field data with established radiological models to provide a thorough evaluation of the site’s current radiological status and to inform future remediation strategies.

Radiological assessment plays a crucial role in understanding and managing the environmental and health impacts associated with radioactive substances. Within the field of radioecology, particular attention is given to the behaviour, distribution, and long-term effects of radioactive contamination in various ecosystems. NORMs are of increasing concern, especially in relation to the exploration and processing of rare earth minerals, which are essential for modern technologies. These minerals often contain significant amounts of thorium, a naturally radioactive element, whose presence necessitates detailed radiological evaluations to ensure safety and compliance with environmental standards. This study aims to address the complexities of such assessments, emphasising the importance of integrating radioecological principles with modern analytical techniques. This approach not only provides risk assessment but is also directly linked to the goals of sustainable development, aimed at minimising the negative impact of past industrial activities and ensuring environmental safety for future generations.

Both thorium and uranium are naturally occurring radioactive elements found in the Earth’s crust, often in overlapping geological settings. While they share some similarities, they differ significantly in terms of isotopic composition, radioactivity, environmental behaviour, and nuclear applications. Thorium is almost exclusively present as 232Th, while uranium occurs as a mixture of ²³⁸U and ²³⁵U [1,2]. Both elements undergo alpha decay and are parents to long radioactive decay chains [3]. Thorium’s toxicity is primarily radiological through inhalation exposure, while uranium poses both radiological and chemical toxicity, particularly nephrotoxicity due to its heavy metal properties [4,5,6,7,8].

The radionuclides contribute to gross alpha radioactivity (e.g., Ra-226, U-238, Rn-222, and their progeny as Po-210), which are associated with highly adverse health effects. It has been demonstrated that chronic exposure to Ra isotopes leads to an increased risk of anaemia, cataracts (in particular bone tumours), and fractures of the teeth. Absorbing uranium isotopes through the digestive tract promotes kidney damage. Furthermore, it has also been confirmed that drinking water with elevated radon levels increases the risk of stomach cancer development [9,10]. Among the radionuclides that contribute to the gross beta radioactivity are K-40, those that come from the uranium and thorium series, in particular Ra-228 and Pb-210 [11]. It is important to bear in mind that in most cases, alpha and beta decay are also accompanied by long-range gamma radiation. As mentioned above, it is necessary to identify the radionuclides present if a high gross alpha/gross beta analysis result is obtained. These radioactive substances are often found associated with thorium and rare earth ores [12].

By order of the Council of Ministers of the Kazakh Soviet Socialist Republic of 21 May 1989, the programme “Research on the radiation ecology of cities of the Kazakh SSR” was developed to assess the impact of the Semipalatinsk test site on the population, including the Ust-Kamenogorsk city. Comprehensive research carried out in 1990–1992 included gamma–radiometric measurements using a helicopter and laboratory tests of samples of soil, snow cover, water, and bottom sediments for the content of natural and artificial radionuclides. A series of flights on a scale of 1:10,000 over the entire urban area and adjacent villages showed that in the southwestern part of the right-bank residential zone of the Ust-Kamenogorsk city, there is a territory characterised by the exposure dose rate of gamma radiation equal to 0.20 μSv/h or more, i.e., at a level of natural background radiation of 0.14–0.16 μSv/h. In the administrative and territorial plan, the location of the discovered facility was determined by the north-eastern part of the Student Campus area and adjacent areas, and its epicentre was the northern wing of the building belonging to the D. Serikbayev East Kazakhstan Technical University. The results of ground-based gamma radiation studies conducted in 2004 (as part of the programme “Conducting comprehensive geoecological studies of the territory and health status of the population of Ust-Kamenogorsk”) confirmed the previous findings.

The detected radioactive contamination began to arise in the 1940s and 1950s as a landfill for post-flotation waste from the processing of rare earth and radioactive ores (metallurgical Plant in Ulba producing monazite, tantalum, and tin concentrates). After production ended in the 1950s, the factory was demolished, and the abandoned area was used for the development of urban infrastructure, which consequently led to the creation of residential zones on waste radioactive rocks. Contaminated areas contain radionuclides of the uranium-238 and thorium-232 series in quantities significantly exceeding their average content for the Ust-Kamenogorsk. The area of radioactive contamination is represented by several regional and local areas stretching along the Serikbaev St. to over 700 m and forming an almost continuous zone of radioactive contamination, the area of which is over 80,000 m², and the level of gamma radiation was elevated, exceeding the natural background characteristic of this area by more than 20 times.

In light of the well-documented deleterious effects of high-dose radiation on living organisms, three research hypotheses were formulated. Firstly, the deeper layers of soil are significantly more contaminated with radionuclides than the surface layers, which indicates the migration of radionuclides deep into the soil profile. Secondly, radioactive isotopes can penetrate water when in contact with soil. The subsequent research hypothesis posits that the doses resulting from the presence of radionuclides in the soil exceed the permissible radiological protection standards. A multi-faceted analysis approach will be employed in order to facilitate an exhaustive examination of the salient issues. Consequently, analysis of the distribution of radionuclide content in the soil profile will facilitate the drawing of inferences regarding the availability of radiation emitters to plant roots, the impact on soil fauna, and the potential for incorporation into trophic chains and biogeochemical cycles. It should be noted that to date, the literature data regarding gross alpha and beta radioactivity levels in Kazakhstan’s soil remains limited to the quantification of water radioactivity. However, it must be emphasised that this is merely one of the fundamental elements required for a comprehensive environmental risk assessment. Taking into account the fact that Kazakhstan is one of the three largest exporters of uranium, this literature gap must therefore be filled. Furthermore, there are a number of industrial facilities engaged in the processing of radioactive ores. It can thus be concluded that the results presented herein are of significant cognitive value and can be related to other locations exhibiting a similar economic use. A component of the innovative approach pertains to the measurements taken at multiple depths, thereby enabling the analysis of soil profiles. Usually, researchers concentrate on the superficial layer of soil. However, plant roots penetrate more profoundly, and radionuclides that have accumulated in the deeper layers can readily percolate into groundwater. The surveys carried out enabled the preparation of a comprehensive and complex analysis of the survey area. The delineated approach has the potential to serve as a preliminary step in the planning and organisation of the management of radioactive waste. Considering the proximity of the residential areas and the public buildings, the results and their interpretation take on an additional significance.

2. Materials and Methods

2.1. Study Area

Samples were collected from a residential area in Ust-Kamenogorsk city (Kazakhstan). The geographical location of the study area is illustrated in Figure 1.

The coordinates of the corners of the radioactive contamination are as follows:

• 49°57′33.3″ N, 82°35′33.6″ E;
• 49°57′32.9″ N, 82°35′34.0″ E;
• 49°57′32.7″ N, 82°35′ 32.0″ E;
• 49°57′32.5″ N, 82°35′32.8″ E.

The climate in the area of sampling (Ust-Kamenogorsk) is strongly continental. The lowest temperatures are recorded in January (with an average of −16.1 °C). Summer is hot, with an average temperature in July of 20.6 °C. Annual precipitation is about 500 mm. As a rule, a stable snow cover is formed in the second decade of November. The average monthly snow depth increases gradually in winter. At the end of February, the freezing depth is about 20 cm. The average wind speed is 2.5–3.5 m/s, with gusts of up to 15 m/s. The sampling area is dominated by winds from the northwest and southeast (along the Irtysh River valley). During the winter months, when anticyclonic weather prevails, southeasterly winds are most frequent. Windy days account for up to 50–70% of the cases during the year. During the rest of the year, the almost windless weather conditions can contribute to the spread of the urban atmosphere with its harmful components. The atmospheric circulation in the studied area is one of the most important climate-forming factors. The Ust-Kamenogorsk region is one of the areas with insufficient atmospheric precipitation. The amplitude of variation in the average monthly air temperature from winter to summer is 37 °C. The absolute extremes of the air temperature can reach up to 49 °C. Air temperature inversions lead to an increase in the concentration of aerosols in the ground layer. The construction of the Ust-Kamenogorsk hydroelectric power station on the Irtysh River has significantly increased the number of days with fog (from 35 to 40 to 65–70). The duration of the fog can vary from a few hours to a few days. Simultaneously, 85% of fog is observed in calm weather and 15% in winds up to 1–3 m/s [13].

The area under investigation is of Lower Quaternary age. It is represented by sandy-gravelly sediments, which are distinctly clayey with a clay admixture in the interlayers. The general direction of groundwater flow coincides with that of surface flowing water. Exceptions are some stretches of streams “suspended” above the alluvial aquifer within the flood terraces and channels of the Ulba and Irtysh Rivers. The aeration zone of the aquifer is formed by permeable sandy and clayey clays. The vertical filtration coefficient of the cover sediments is between 0.1 and 2 m per day. There are upper groundwater levels of floodplains and the first terrace above the floodplain at a depth of up to 5 m, during floods—up to 1.5–2.5 m. Currently, the natural surface of the land is heavily disturbed (small excavations, blockages of canals, and consequences of other construction works), which contributes to an increased absorption of all types of wastewater [14].

The main surface streams are the Irtysh and Ulba Rivers, whose watersheds are formed in the surrounding areas. The Irtysh riverbed is about 1500 m to the west, the Ulba river—740 m to the south of the sampling site. These areas are prone to extreme flooding, which can cause some of the buildings and houses in the rural areas of the islands to be inundated and lead to massive run-off of manure, domestic and household waste, which can be a major source of pollution. The sampling site is confined to the first terrace above the floodplain, from an absolute height of 285 m. The area of sampling was 480 m², with samples being obtained from a depth of up to 100 cm from the ground surface. The samples consist of a gravel mixture (up to a depth of 100 mm) and sand-clay filler (40–50%). An analysis of the study area reveals the presence of a thin layer of poorly developed turf. The area is represented by a typical floodplain vegetation flora such as willow, aspen, poplar, cattail, and other species of trees and shrubs, as well as ruderal plants. As a result of human economic activity, the fauna is limited and represented by birds (sparrow, magpie, crow, tit, pigeon) and rodents (field mouse, house vole). There is a notable absence of rare and endangered species of flora and fauna, as well as natural sources of food and medicinal plants.

2.2. Pedestrian Gamma Surveys

A SRP 68-01 portable geological radiometer (Elektro-mag, Rovno, Ukraine) and a DKS-96 radiometer-dosimeter (manufactured by Scientific Production Company “Doza”; Moscow, Russia) were used for the field investigations. A total area of at least 1500 m² was measured. Fixed point measurements were made by placing the tip of the SRP 68-01 radiometer sleeve on the ground and reading the radiometer readings collected over 5–10 s. Using the geographic information system MapInfo Professional (MAPA; MapInfo Corporation, New York, NY, USA), gamma radiation exposure doses were plotted on a topographic base of the study area. Subsequently, gamma dose rate measurements were made in boreholes up to 1 m deep along the borehole wall using an SRP-88 radiometer (Diapazon; Kaliningrad, Russia) (taking into account background radiation in layers 0–100 cm). Gamma-ray background measurements from 0 to 100 cm were also performed.

2.3. Radioactivity Measurements

The main objective of soil sampling is to establish correlation coefficients between the gamma radiation dose level and the specific activity of the radiochemical constituents. Firstly, along the designated survey point, a 120–130 cm deep pit was drilled. Then samples were taken sequentially at intervals of every 20 cm from the surface, along the trench wall, and following a 20 × 20 cm square in the horizontal plane from the edge of the trench wall. The collected material was cleaned of contaminants (plant roots, insects, stones, glass, etc.), sieved through a 5 mm mesh sieve, and then homogenised. A sample of at least 1 kg selected for further testing was placed in a plastic, labelled bag. A total of 28 pits were excavated, from which 140 samples were collected as a result of sampling at the former ore processing plant. Samples were prepared in accordance with GOST 17.4.4.02-84 Standard [15] procedure.

Prior to radiochemical analysis, samples were oven dried at a temperature not exceeding 105 °C and sieved through a 1 mm mesh size (yield of the 1 mm fraction at least 95%). A representative sample of 50 g was ground in a laboratory planetary mill to a particle size of less than 71 microns. The resultant material was divided into 2 parts. One was used to study the total alpha and beta radioactivity. The remaining part was used for the determination of the specific activity of the gamma-ray emitters (K-40, Ra-226, Th-232). The calibration of the spectrometers was conducted in accordance with the ISO 4037-3:2015 (ISO 4037) Standard [16].

The total activity of alpha and beta emitters was determined using a UMF-2000 low background alpha-beta radiometer (manufactured by Scientific Production Company “Doza”; Moscow, Russia). The accuracy of the method has been confirmed by measuring a certified reference sample containing ²³⁸U equilibrated with decay products (State Standard Sample SG-1A, the radionuclide content and matrix composition were close to the typical composition of the analysed soil samples). Meanwhile, for measuring total beta activity, potassium sulphate is widely used as a reference material. In the first step, the sample is calcined in a muffle oven at 350 °C for 1 h and then left to stand for 3 h in order to minimise the interfering effects of radon-222 and its decay products. The resulting sample was then homogenised by grinding with a porcelain pestle. 1.0 g was collected and transferred to a measuring cell previously cleaned with ethanol. Measurement of the alpha or beta particle count rate is made between 3 and 15 h after sample calcination (during this time, the effect of Rn-222 and its decay products on the results is minimal). Each sample was measured at least 3 times for 1000 s. The total specific activity of the alpha- or beta-emitting radionuclides in the sample was calculated using the relative method with correction for radioactive background. The final result was taken as the mean of 3 parallel determinations of the sample.

The activity of the gamma-emitting radionuclides (Ra-226, Th-232, K-40) was determined by gamma-ray spectrometry (InSpector 1000, Canberra, Smyrna, GA, USA). The 2″ × 2″ high-performance NaI scintillation detector equipped with a multichannel analyzer was working within the energy range of 50 keV–3 MeV. The previously prepared homogenised samples were transferred to a 4π measurement geometry vessel. The container was closely sealed and stored for at least 21 days to achieve the radioactive equilibrium between radium (Ra-226) and radon gas (Rn-222). The sample was then placed in the lead shielding of a pre-calibrated spectrometer. K-40 was determined using a single line at 1460 keV, while Ra-226 and Th-232 were detected by the progeny. In the case of Ra-226, the activity was calculated based on peaks originating from Pb-214 and Bi-214. For Th-232, full energy peaks of Ac-228 (911 and 969 keV) were used. Regarding U-235, the instrumentation cannot provide a quantitative evaluation of its activity due to its low environmental abundance (~0.7%) and the insufficient detector resolution. The content of Cs-137 was not considered in the presented paper–its concentration was negligibly low compared to the activity of Ra-226, Th-232, and K-40, resulting in a large measurement error. At least three parallel measurements were made for each of the 140 samples. The collected gamma spectra were analysed using GENIE software 1.1 version. The radionuclide activity is given as follows:

$$C={\displaystyle \frac{A}{M\epsilon IT{K}_c{K}_w}}$$

(1)

where

• A is the net peak area,
• M is sample mass,
• ε is detector efficiency,
• I is branching factor,
• T is measurement live-time,
• K_c and K_w are factors for decay and corrections, respectively.

Assessment of radionuclide migration capacity measurements was carried out in the following way. First, they were dried at a maximum temperature of 40 °C. Next, the sample was spread out on a polyethylene film, and any large lumps were crushed using a spatula or pestle in order to remove any contaminants, such as plant roots, insects, stones, glass, coal, or animal bones. The purified sample was sieved through a 1 mm sieve and homogenised. A sample weighing 300 g was transferred to a 2 dm³ flask, to which 1.5 dm³ of distilled water was added. As a result, suspension with the solid-to-liquid ratio of 1:5 was obtained. In the next step, the suspension was shaken for 2 h, incubated for 22 h at room temperature, and then filtered through a 0.45 μm filter. 1 dm³ of the filtrate was collected and evaporated to dry, then calcined in a muffle furnace for 1 h at 350 °C (in the presence of ethyl alcohol). The resulting suspension was poured into a pre-weighed measuring cuvette and dried until a uniform, homogeneous layer was obtained. The sample prepared in this way was weighed and subjected to gamma spectrometric measurement.

2.4. Calculation of Radiological Hazards

One of the primary objectives of the present study is to ascertain the extent of the radiological risk posed by the disposal of radioactive ore and rare earth processing waste. Therefore, a number of parameters were calculated in order to provide a comprehensive assessment of the potential health implications and the associated carcinogenic risk for the local population.

Radium equivalent activity (Raeq; Bq/kg) is a widely used parameter for the identification of the radiation exposure associated with the occurrence of gamma emitters in soil [17].

$${Ra}_{eq}=A_{Ra-226}+1.43A_{Th-232}+0.077A_{K-40}$$

(2)

A_Ra-226, A_Th-232 and A_K-40 are the activity [Bq/kg] of Ra-226, Th-232, and K-40, respectively.

It is assumed that the Raeq value should not exceed 370 Bq/kg. This threshold is equivalent to an effective dose of 1.5 mGy/y received from gamma radiation [18]. The coefficients at the activities of the individual isotopes are the result of the assumption that 1 Bq/kg of Ra-226, 0.7 Bq/kg of Th-232, and 13 Bq/kg of K-40 produce the same gamma dose rate [19].

In order to ascertain the external terrestrial gamma radiation characteristic, it is necessary to calculate the absorbed dose rate (ADR; nGy/h). It corresponds to terrestrial gamma radiation measured at a height of 1 m above the soil surface. It is noteworthy that all decay products of Ra-226 and Th-232 should be in radioactive equilibrium with the parent nuclei [17]. The global average is between 55 and 57 nGy/h. The safe threshold recommended by ICRP is 1000 nGy/h [20], while UNSEAR recommends the level of 1500 nGy/h [21].

$$ADR=0.462A_{Ra-226}+0.604A_{Th-232}+0.041A_{K-40}$$

(3)

A_Ra-226, A_Th-232 and A_K-40 are the activity [Bq/kg] of Ra-226, Th-232, and K-40, respectively.

The annual effective dose equivalent (AEDE; mSv/y) constitutes an evaluation of potential biological effects experienced by populations exposed to ionising radiation.

$$AEDE=ADR ·CF ·OF ·T · {10}^{-6}$$

(4)

where ADR refers to the absorbed dose rate (nGy/h). CF—conversion factor (0.7 Sv/Gy), used to convert the dose absorbed at 1 m to the effective dose that is received by an adult. OF—occupancy factor, for outdoor occupancy equal to 0.2 (estimation that the average person spends approximately 20% of the time in outdoor environments). T—time of one year expressed in hours (8760 h). 10⁻⁶—coefficient used to convert dose expressed in nano- to miliSv [22]. The worldwide mean AEDE value is 0.07 mSv/y [17].

Based on the AEDE, the excess lifetime cancer risk (ELCR) can be calculated. This unitless parameter is used to estimate the probability of cancer in the human population during the lifetime of an individual exposed to NORM [23].

$$ELCR=AEDE ·DL ·RF$$

(5)

DL—life expectancy (70 years as specified by WHO in 2014). RF—fatal cancer risk factor (0.057 per Sv for stochastic effects on ionising radiation proposed by ICRP). Worldwide mean value is 0.29 × 10⁻³ [17].

The annual gonadal dose equivalent is defined as the measure of the genetic significance of the yearly dose received by the population’s reproductive organs:

$$AGDE=3.09A_{Ra-226}+4.18A_{Th-232}+0.314A_{K-40}$$

(6)

A_Ra-226, A_Th-232 and A_K-40 are the activity [Bq/kg] of Ra-226, Th-232, and K-40, respectively. Worldwide average is 300 μSv/y [17].

The gamma level index (Iγ) is used to monitor the annual dose rate received by exposure to external gamma radiation emitted from the superficial soil layer. It is calculated using the following formula:

$$I_{\gamma }={\displaystyle \frac{A_{Ra-226}}{300}}+{\displaystyle \frac{A_{Th-232}}{200}}+{\displaystyle \frac{A_{K-40}}{3000}}$$

(7)

A_Ra-226, A_Th-232 and A_K-40 are the activity [Bq/kg] of Ra-226, Th-232 and K-40, respectively [17]. If the calculated values are greater than or equal to 1, the soil may be deemed contaminated [24].

3. Results and Discussion

3.1. Gamma Radiation Dose

The main objective of the pedestrian surveys carried out in the area of the potentially contaminated site was to produce a detailed radiation map of its surface and surrounding area (Figure 1). The background radiation levels measured for the analysed area were in the range of 0.16 to 0.18 µSv/h. As can be seen, the gamma radiation dose (GRD) rate values vary widely from 0.20 to 1.00 μSv/h. Based on the 2σ criterion, GRD values of 0.20 µSv/h and above were classified as abnormal. The results of the gamma survey are reported in Table S1. Consequently, the area was divided into smaller sectors (survey points are marked A-1 to A-28). As demonstrated, the radiation dose increases with the increase in sampling depth. This may be a consequence of the fact that more radioactive material was stored at a given depth. The gradual migration of radioactive substances into soil and groundwater reservoirs may explain the slight decrease in dose at depths above 80 cm. Furthermore, the distribution of values recorded above 20 cm is not uniform, with hot spots occurring (e.g., A-1, A-11, A-21). These sites are likely to have the highest concentration of residues from ore processing operations. One potential explanation for this phenomenon is the burial of post-processing waste in the dug holes.

3.2. Radioactivity of the Analysed Area

3.2.1. Gross Alpha/Gross Beta

Gross alpha and beta activity for soil samples are shown in Figure 2. The geological formation of the area, the concentration of the mineral component, and the nature of human activities in the area are the three main factors that determine the gross alpha/beta activity value [25]. Analysis of the collected data (Figure 2, Tables S2 and S3) indicates that the total alpha radioactivity is significantly higher in comparison to the gross beta. This is explained by the fact that the gross alpha measurement applies to all alpha emitters present in a given sample, not just Ra-226 and Th-232, but also their progeny. It is related to the composition of the uranium and thorium decay series. In the series starting with U-238, there are 8 alpha emitters (including the parent nuclei, U-234, Th-230, Ra-226, Rn-222, Po-218, Po-214 and Po-210) and 6 electron emitters (Th-234, Pa-234m, Pb-214, Bi-214, Tl-210, Bi-210), while in the thorium-232 decay chain we can distinguish another 6 radionuclides emitting the alpha particles and 5 beta emitters. Furthermore, the total α/β activity distribution is not uniform. Regardless of the sampling depth, there are ‘hot spots’ characterised by elevated gross alpha and beta radioactivity. This phenomenon is associated with the initial distribution of post-production waste. Elevated levels of total soil radioactivity are an indication of a potential threat to the entire local ecosystem. Soil acts as a reservoir for radionuclides, which can migrate through water supplies and, depending on their chemical properties, can subsequently enter the food chain via uptake processes by plant roots and animals. Soil can therefore be considered as a radiological indicator of the local environment. In conclusion, radiation monitoring of soil on post-industrial sites is extremely important for the classification and characterisation of contaminated areas. In the topsoil layer (0–20 cm), the highest gross alpha value was obtained for samples A-1 (5275 Bq/kg), A-12 (4200 Bq/kg) and A-10 (3487 Bq/kg), while the lowest concentrations of radioactive particles were recorded for A-9 (583 Bq/kg), A-27 (774 Bq/kg) and A-28 (732 Bq/kg). In the case of gross beta tests, the maximum value was obtained for sample A-12 (1749 Bq/kg), while sample A-8 was characterised by the lowest beta emitter activity (641 Bq/kg). As can be seen, gross alpha/beta values increase dramatically with sampling depth. It can therefore be assumed that tailings deposition was followed by slow migration of radioactive series elements deep into the soil. The deeper the layer of soil is, the less likely it is that the radon isotopes will be able to escape to the surface. Thus, the nuclei resulting from their decay become trapped between the soil grains, increasing its radioactivity. The mean total alpha radioactivity values from the uppermost surface to the deepest soil layer were as follows: 2134, 5791, 9270, 11,115, and 12,627 Bq/kg, respectively. In the case of gross beta, the following values were recorded: 969, 1852, 2620, 3215, and 3664 Bq/kg. As depicted in Table S2, for the topsoil layer, the skewness is slightly positive with a value of 1.06. In the subsequent layer (20–40 cm), the distribution is significantly shifted towards elevated values (skewness of 2.04), as evidenced by the kurtosis value of 5.68. It has been demonstrated that, upon exceeding 40 cm, a decline in skewness values becomes observable, attributable to an increase in the number of radioactive points. The value calculated for samples collected from 80 to 100 cm indicates a symmetrical distribution. This distribution is not normal, as evidenced by the kurtosis value of −0.96 (for a normal distribution, kurtosis = 0). In this case, the values for points exhibiting high α-emitter activity are offset by the presence of less radioactive areas (the histogram for the aforementioned samples is presented in Figure S1a). The coefficient of variation (COV) values initially increase in the sampling depth, reaching a maximum of 72.0% for 40–60 cm. High COV values are indicative of the anthropogenic origin of radionuclides. It can be concluded that, under the analysed conditions, the local soil can retain larger quantities of radionuclides through processes such as adsorption, ion exchange, and microprecipitation.

Analogous relationships are observed for gross beta results (Table S3). The distribution of the topsoil layer demonstrates a slight positive skew. As the sampling depth increases, so does the number of points characterised by elevated β-emitter radioactivity (demonstrated as a highly skewed distribution for 20–40 cm and 40–60 cm layers with the kurtosis rising from 5.85 to 3.85). In deeper soil layers, the distribution tends to be symmetrical, but it is not a normal distribution (for the 80–100 cm layer, the skewness is −0.03, kurtosis = −1.02; Figure S1b). The COV analysis demonstrated that only the value for the 0–20 cm layer (COV = 28.4%) can be considered indicative of the natural origin of beta radiation emitters. The remaining values (ranging from 46.1% to 66.6%) unequivocally signify human activity. A comparison of the obtained data with the relevant literature was included in the Supplementary Material. The high coefficients of variation calculated for the gross alpha and gross beta (over 50% for Th-232 and Ra-226 at various depths) indicate an extremely non-uniform distribution of contamination, which justifies the need for targeted remediation measures.

It should be noted that the World Health Organisation recommends a maximum gross alpha concentration limit in water of 0.555 Bq/L and 1.0 Bq/L for gross beta activities [26,27]. The radionuclides contribute to gross alpha radioactivity (e.g., Ra-226, U-238, Rn-222, and their progeny as Po-210), which are associated with highly adverse health effects. It has been demonstrated that chronic exposure to Ra isotopes leads to an increased risk of anaemia, cataracts (in particular bone tumours), and fractures of the teeth. Absorbing uranium isotopes through the digestive tract promotes kidney damage. Furthermore, it has also been confirmed that drinking water with elevated radon levels increases the risk of stomach cancer development [9]. Among the radionuclides that contribute to the gross beta radioactivity are K-40, those that come from the uranium and thorium series, in particular Ra-228 and Pb-210 [11]. It is important to bear in mind that in most cases, alpha and beta decay are also accompanied by long-range gamma radiation. As mentioned above, it is necessary to identify the radionuclides present if a high gross alpha/gross beta analysis result is obtained. These radioactive substances are often found associated with thorium and rare earth ores [12]. Large quantities of radioactive materials located at greater depths pose a high radiological risk. Firstly, radionuclides can be continuously incorporated into the food chain through the presence of plant roots (including perennial plants). It is worth noting that the main part of the root system of some plants (including those of economic importance) is located at a depth of 100 cm [28]. Soil is also the habitat of many animal species, which can be endangered by high doses of radiation. This would lead to soil sterilisation due to the lack of suitable organisms. Furthermore, the radionuclides’ leakage into the groundwater also poses a real risk to the region’s inhabitants.

3.2.2. Radioactivity Distribution

Figure 3a,b the average activity of gross alpha and beta isotopes.

A thorough analysis of radioactivity and evaluation of associated risk factors was undertaken for two sampling depths (0–20 cm and 40–60 cm). The first of these is of crucial importance with regard to population exposure and is also susceptible to erosion, surface runoff, and weathering. In turn, the root systems of certain plants (including those of economic importance) are located at greater depth [28]. As depicted in Figure 4a, the average activity of Ra-226, Th-232, and K-40 for samples from the 0 to 20 cm layer was 43, 184, and 537 Bq/kg, respectively. The Ra-226 minimum activity was detected in A-28 (23 Bq/kg). In the case of Th-232, the minimum activity was noted for A-9 (55 Bq/kg), whereas the lowest amount of K-40 was measured for the A-27 sample (445 Bq/kg). At the 0–20 cm layer, the maximum of all analysed radionuclides was determined for A-1 (Ra-226: 75 Bq/kg; Th-232: 450 Bq/kg; K-40: 616 Bq/kg). In general, the highest activity levels were recorded for K-40, with the lowest values observed for Ra-226. It can be concluded that the markedly lower activity values of the labelled radionuclides (compared to higher sampling depths) may be attributed to the combined effects of erosion, weathering, and flushing processes. Similar to the thorium, radium isotopes tend to adsorb on solid surfaces, so it does not migrate far from their storage or release site. Due to its valence and chemical properties similar to barium and calcium, Ra can be absorbed from the soil by plants and integrated into the food chain. In the human body, Ra isotopes are deposited mostly in the bone and teeth structure. In turn, inhaled radium can remain in the lungs for several months [29]. Therefore, its elevated radioactivity at depths where the plant roots exist poses a potential radiological threat to human and non-human biota.

The statistical data processing for the 0–20 cm soil layer is presented in Table 1. The skewness values for Ra-226 and Th-232 were close to normal distribution, ranging from 0.62 to 0.88, respectively. The data calculated for K-40 was slightly negative, suggesting a symmetrical distribution (the skewness of −0.16). The kurtosis results suggest the approximately symmetric (for Ra-226) or moderately skewed distribution (Th-232, K-40). The low COV value of 9.09% for K-40 indicates its natural origin, in contrast to Ra-226 and Th-232 (COV values are 31.9% and 50.8%, respectively).

Table 1. Statistical parameters for radioactivity measurements for the 0–20 cm layer.

Radioactivity Measurements for 0–20 cm
Radionuclide		Ra-226	Th-232	K-40
Concentration in Bq/kg	Average	42.9	184	537
	Median	39	172	545
	Minimum	23	55	445
	Maximum	75	450	616
Skewness		0.62	0.88	−0.16
Kurtosis		−0.11	0.89	−1.08
Coefficient of variation		31.9	50.8	9.09

Table 1. Statistical parameters for radioactivity measurements for the 0–20 cm layer.

Radioactivity Measurements for 0–20 cm
Radionuclide		Ra-226	Th-232	K-40
Concentration in Bq/kg	Average	42.9	184	537
	Median	39	172	545
	Minimum	23	55	445
	Maximum	75	450	616
Skewness		0.62	0.88	−0.16
Kurtosis		−0.11	0.89	−1.08
Coefficient of variation		31.9	50.8	9.09

Figure 3. Changes in gross alpha activity (a) and gross beta activity (b) in samples collected at different depths.

Figure 3. Changes in gross alpha activity (a) and gross beta activity (b) in samples collected at different depths.

Figure 4. Activity of Ra-226, Th-232, and K-40 in samples taken from depth: (a) 0–20 cm, (b) 40–60 cm.

Figure 4. Activity of Ra-226, Th-232, and K-40 in samples taken from depth: (a) 0–20 cm, (b) 40–60 cm.

Contrary to the K-40, Th-232 and Ra-226 activity increases significantly with the sampling depth. The average contribution at the depth of 20–40 cm was in the range 14–66% and 4.6–11% for Th-232 and Ra-226, respectively. The processing of monazite, the main source of thorium and rare earth elements, explains the considerable increase in Th-232 radioactivity in soil samples. It is important to note that the raw material contains trace quantities of radium isotopes (Ra-228 and Ra-226), constituting the decay products of Th-232 and U-238. The content of radium in monazite is determined by its grade and the relative proportions of uranium and thorium in the ore. Estimates suggest that the average amount of Ra isotopes in monazite is lower than 1 mg/metric ton of thorium for Ra-228, and 300 mg/metric ton of uranium. This discrepancy is associated with the half-life values of both radionuclides. The considerably longer half-life of Ra-226 (1600 years in contrast to 5.75 years for Ra-228) results in the accumulation of this isotope in the tailings, thereby posing a serious radiological hazard [30,31,32].

In the case of a 40–60 cm soil layer (Figure 4b,

Table 2. Statistical parameters for radioactivity measurements for the 40–60 cm layer.

Radioactivity Measurements for 40–60 cm
Radionuclide		Ra-226	Th-232	K-40
Concentration in Bq/kg	Average	129	738	549
	Median	102	598	549
	Minimum	21	47	465
	Maximum	404	2530	646
Skewness		1.79	1.80	0.09
Kurtosis		3.62	3.42	−0.84
Coefficient of variation		66.8	77.8	9.22

; data for other depths are presented in Figure S2, Table S4), the average radionuclide concentrations were 129, 738, and 549 Bq/kg. Sample A-10 exhibited the lowest concentration of Ra-226 (21 Bq/kg) and Th-232 (47 Bq/kg). The minimum K-40 concentration was detected in A-14 (465 Bq/kg). The maximum radioactivity of Ra-226 and Th-232 was noted for A-10 (404 and 2530 Bq/kg, respectively), whereas for K-40, the highest activity was measured for A-9 (646 Bq/kg). The skewness values for Ra-226 and Th-232 are almost the same (1.79 and 1.80, respectively). The lowest value for this parameter is reached for K-40 (0.09), which approaches a symmetrical distribution. At the same time, the kurtosis decreases in the Ra-226 > Th-232 > K-40 series, ranging from 3.62 for Ra-226 to −0.84 for K-40. The high COV values for radionuclides originating from the natural radioactive decay series (66.8% for Ra-226, 77.8% for Th-232) clearly indicate that the source of the radionuclides in the 40–60 cm layer is post-production waste (compared to K-40, COV = 9.22%). It can be concluded that the mean concentrations of Th-232 and Ra-226 are higher at depths of 40–60 cm and even 60–80 cm than at the surface. Atmospheric factors and processes, including erosion, runoff, and weathering, have been demonstrated to affect the concentration of radionuclides on the surface. Consequently, the concentration of analytes is lower in the deeper soil layers than in the shallower ones. It is possible that such a situation is a result of waste burial. Nevertheless, the migration cannot be excluded as one of many potential mechanisms.

Beyond a depth of 40–60 cm (Figure S2), an increase in the radioactivity of Th-232 and Ra-226 is still observed, although the magnitude is lower in contrast to the topsoil layer. Furthermore, a significant increase in the Ra-226 and Th-232 concentrations is evident in samples situated in close proximity to areas of elevated activity. In case of A-9 located at some distance from the A-1, the Th-232 radioactivity rises from 913 Bq/kg at depth 40–60 cm to 1760 Bq/kg at 60–80 cm. The analysis of statistical parameters can lead to interesting conclusions (Table S4). The skewness values for samples from a depth of 80–100 cm are significantly lower than for the topsoil layer, suggesting close proximity to symmetrical distribution (with kurtosis ranging from −0.67 for Ra-226 to −0.99 for Th-232). The COV values for Ra-226 (48.0%) and Th-232 (51.3%) decrease significantly compared to previously analysed depth. It can thus be concluded that horizontal migration of the stored radioactive substances has occurred. The supposition is corroborated by the analysis of the Th-232 and Ra-226 radioactivity levels in samples collected from the most distant locations from the hotspot (e.g., A-14, A-23). It should be remembered that thorium preferentially adheres very tightly to soil particles [32]. One of the monazite processing methods involves the acid cracking step to extract the thorium and lanthanide content. In order to do so, the starting mineral is heated with concentrated sulfuric acid to temperatures ranging from 120 °C to 150 °C for several hours. As a result, the final product is thorium–containing mud (in the phosphate cake form) and a solution of trivalent lanthanides. It is notable that other REE processing techniques also yield water-insoluble thorium compounds, such as ThO₂, or phosphates [33]. The major pathways of exposure to Th isotopes are ingestion and inhalation of thorium–contaminated dust. Despite the fact that Th-232 is characterised by low specific activity, its progeny is highly radioactive (4.07 × 10³ Bq/g for Th-232, 10.09 × 10¹² for the Ra-228 [34]. Although thorium is generally considered to be immobile, reports of its migration in soil can be found in the literature. The concentration of total dissolved thorium in soil and groundwater can increase through the formation of various aqueous complexes formed with inorganic anions such as dissolved carbonate, fluoride, phosphate, chloride, and nitrate [35]. The solution pH and the presence of humic substances are considered to be particularly important in the Th-232 sorption process [36,37]. In addition, Th-232 can migrate with groundwater flow, particularly in fractured rock systems [38].

The data and statistical information pertaining to gross alpha and gross beta radioactivity in water-soluble and ion-exchangeable forms are presented in

Table 3. The total radioactivity of the water-soluble and ion-exchangeable forms of the alpha-emitting radionuclides.

	The Experimental Value of Gross Alpha Water-Soluble Forms [Bq/kg]					The Experimental Value of Gross Alpha Ion-Exchangeable Forms [Bq/kg]
	Sampling Depth [cm]					Sampling Depth [cm]
Sample Code	0–20	20–40	40–60	60–80	80–100	0–20	20–40	40–60	60–80	80–100
A-1	38	131	187	167	124	37	133	164	155	148
A-2	15	57	113	121	171	12	63	112	145	162
A-3	25	29	38	64	80	21	35	43	62	113
A-4	29	59	67	56	32	28	71	71	69	40
A-5	21	41	57	121	124	16	44	57	109	164
A-6	23	34	40	62	71	26	34	48	71	65
A-7	15	63	139	104	124	16	68	123	122	132
A-8	25	64	79	105	83	25	77	95	100	85
A-9	12	13	87	153	127	9	14	77	143	128
A-10	34	44	12	11	12	33	50	9	9	10
A-11	16	104	193	200	122	12	88	216	181	128
A-12	32	48	28	42	49	38	51	28	46	53
A-13	26	50	46	82	130	20	43	43	85	141
A-14	14	20	51	91	127	14	23	46	98	121
A-15	24	32	50	66	70	19	38	42	57	62
A-16	29	42	47	92	151	23	48	47	93	124
A-17	30	59	64	120	136	29	47	52	99	143
A-18	22	31	48	79	93	18	35	49	76	106
A-19	17	43	53	66	81	16	35	64	74	74
A-20	19	42	66	78	77	14	37	60	68	78
A-21	22	28	56	75	107	18	28	58	94	118
A-22	20	39	41	53	60	20	35	43	55	50
A-23	17	49	84	100	68	15	44	83	86	87
A-24	17	27	48	51	58	13	31	41	53	68
A-25	17	24	21	20	19	16	20	21	17	17
A-26	17	25	31	55	91	18	28	28	58	107
A-27	14	18	31	26	17	8	17	30	19	19

,

Table 4. The total radioactivity of the water-soluble and ion-exchangeable forms of the beta-emitting radionuclides.

	The Experimental Value of Gross Beta Water-Soluble Forms [Bq/kg]					The Experimental Value of Gross Beta Ion-Exchangeable Forms [Bq/kg]
	Sampling Depth [cm]					Sampling Depth [cm]
Sample Code	0–20	20–40	40–60	60–80	80–100	0–20	20–40	40–60	60–80	80–100
A-1	16	57	87	70	47	16	56	74	60	54
A-2	10	29	42	50	57	9	28	41	45	51
A-3	9	15	15	27	43	10	15	17	34	45
A-4	11	24	32	22	18	12	24	28	23	16
A-5	13	18	25	52	57	11	14	23	53	51
A-6	14	14	18	30	26	13	13	20	34	32
A-7	10	21	49	44	43	10	26	43	42	52
A-8	15	27	29	36	37	14	25	37	33	38
A-9	7	8	35	55	53	7	7	35	47	60
A-10	11	19	6	6	9	10	20	7	7	8
A-11	8	38	77	69	55	8	32	83	73	56
A-12	19	23	13	20	25	19	22	14	16	24
A-13	12	19	19	35	41	11	15	15	32	47
A-14	7	10	17	27	50	9	10	20	27	42
A-15	10	16	22	19	21	10	17	21	24	24
A-16	12	17	19	42	55	12	18	20	34	53
A-17	14	22	31	49	63	16	20	27	41	64
A-18	10	19	15	25	45	11	18	16	30	39
A-19	10	18	26	34	22	8	17	25	32	29
A-20	8	15	29	29	26	8	16	27	32	30
A-21	7	16	18	37	51	8	17	17	37	49
A-22	11	14	16	22	26	11	13	18	20	28
A-23	8	21	33	29	31	8	19	40	32	27
A-24	8	14	24	19	19	7	13	23	24	23
A-25	9	9	12	9	11	9	9	12	9	9
A-26	9	12	16	23	44	8	10	15	23	44
A-27	6	11	11	12	12	8	9	14	12	19

, Tables S5 and S6 (in Supplementary Materials). As demonstrated, radionuclides emitting alpha and beta radiation are strongly bound to soil particles. The analysis of total radioactivity indicates that negligible quantities of the radioactive substances are capable of permeating into aqueous solutions. Among the analysed samples, there are points with the highest gross alpha or gross beta activity, coinciding with the occurrence of hot-spots present on the gamma dose rate map. In comparison with the total radioactivity data obtained for soil samples, the results of the water-soluble and ion-exchangeable forms are found to be several times lower. The minimum gross alpha value for the topsoil layer is 583 Bq/kg (A-9), while the maximum radioactivity of water-soluble substances is 38 Bq/kg (A-1 sample; the same value was recorded for the gross alpha for ion-exchangeable forms in A-12). A comparable phenomenon is observed in the context of radionuclides that emit beta-minus radiation. The lowest recorded level of total beta radioactivity was 641 Bq/kg (A-27), whereas the maximum level of radioactive species present in aqueous solutions was 19 Bq/kg (A-12, both water-soluble and ion-exchangeable forms).

A comparative analysis of the values for the aqueous forms clearly indicates discrepancies in the radioactivity of the gross α/β forms. For the topsoil layer, the maximum value for water-soluble alpha forms is reached for the A-1 sample (38 Bq/kg), while the highest radioactivity for the gross beta water-soluble alpha species is two times lower (19 Bq/kg, A-12). As the sampling depth increases, the radioactivity of water-soluble forms also increases. This phenomenon can be attributed to the presence of elevated concentrations of radioactive substances in the deeper layers of the soil, as evidenced by the results of individual radionuclide content measurements. The maximum values of forms migrating to aqueous solutions occur for the 60–80 cm layer. It is demonstrated that the values obtained for the water-soluble and ion-exchange forms are practically identical. It can be concluded that only ion-exchange forms of radionuclides have passed into the water phase. It has therefore been confirmed that Ra-226 and Th-232 exhibit a strong affinity for the soil particles. The skewness and kurtosis data (Tables S5 and S6) for alpha and beta emitters indicate a non-uniform distribution of results. High values of the coefficient of variation parameter are indicative of the anthropogenic nature of radionuclides contained within aqueous solutions.

The Pearson correlation coefficients (confidence level of 0.05) are shown in Table S7a–e. The data analysis reveals that the majority of relationships are sustained irrespective of sampling depth, except for K-40. A highly significant positive correlation close to 1 is observed between gross alpha and gross beta activity. The relationship may be attributed in part to the fact that the raw materials processed at the former Ulba site are accompanied by both alpha and beta radiation emitters. Furthermore, the process gives rise to the generation of waste that contains various radionuclides. The significant positive correlations (in the range of 0.8560–0.9886) between gross alpha and Ra-226 and gross alpha and Th-232 are unsurprising—both radionuclides are alpha emitters. It was observed that the lowest values were obtained for samples taken at a depth of between 40 and 60 cm (Table S7c) as a result of possible horizontal migration of radioactive substances. A number of noteworthy correlations were obtained for gross beta versus Ra-226 and Th-232. Despite the fact that both radionuclides are classified as alpha emitters, the Pearson correlation coefficient for Ra-226 ranges from 0.9089 to 0.9854, while for Th-232, the coefficient varies from 0.9050 to 0.9859. Moreover, the strongest correlation was identified in samples collected from depths ranging between 40 and 60 cm. It is important to acknowledge that alpha and beta radiation do not only originate from Ra-226 and Th-232, but also from their decay products. These, in turn, are trapped in deeper soil layers, where they are predominantly accessible to the plant root systems [29].

It was found that the lowest correlation coefficient was obtained for K-40. Surprisingly, the correlation between β-emitting K-40 and gross beta radiation is not observed at any sampling depth. The collected data indicate that K-40 was not introduced into the soil with the tailings. Additionally, its high mobility results in a significant tendency to migrate in the environment. However, a gradual increase in the Pearson correlation coefficients for K-40 is observed with increasing sampling depth (excepting the 20–40 cm layer) [39]. A literature review shows that most grass species typically have root systems between 30 and 90 cm deep, depending on the grass species, soil quality, climate, and environmental conditions [33,34,40]. Consequently, the intensive uptake of nutrients by roots abundant at a depth of 20–40 cm results in a disruption to the potassium levels within the soil, including K-40. The maximum values of the correlation coefficients for K-40 are observed for samples taken at a depth of 80–100 cm, where perennial plant roots are typically found, though their volumetric biomass is often lower compared to the shallower grass roots. This results in less noticeable changes in the K-40 content.

3.3. Radiological Hazard Analysis

The calculation of radiological risk factors for different sample collection depths is presented in

Table 5. Radiological hazard calculations made for superficial soil layer (0–20 cm).

Sample Code	Radiological Hazard Indices
	Raeq [Bq/kg]	ADR [nGy/h]	AEDE [mSv/y]	ELCR [10−3]	AGDE [μSv/y]	Iγ
1	766	332	0.41	1.63	2 306	5.41
2	269	120	0.15	0.59	845	1.95
3	376	166	0.20	0.81	1158	2.69
4	471	206	0.25	1.01	1430	3.34
5	366	162	0.20	0.79	1131	2.62
6	394	173	0.21	0.85	1203	2.80
7	290	128	0.16	0.63	898	2.08
8	470	205	0.25	1.00	1422	3.33
9	141	65	0.08	0.32	459	1.04
10	449	197	0.24	0.96	1369	3.19
11	215	97	0.12	0.47	678	1.56
12	616	268	0.33	1.31	1857	4.35
13	468	205	0.25	1.00	1426	3.33
14	201	91	0.11	0.44	638	1.46
15	376	166	0.20	0.81	1161	2.69
16	509	223	0.27	1.09	1550	3.62
17	482	211	0.26	1.03	1467	3.43
18	410	180	0.22	0.88	1256	2.92
19	255	114	0.14	0.56	803	1.85
20	232	104	0.13	0.51	732	1.68
21	299	133	0.16	0.65	931	2.15
22	398	175	0.21	0.86	1222	2.84
23	235	106	0.13	0.52	744	1.71
24	233	104	0.13	0.51	728	1.68
25	246	111	0.14	0.54	777	1.79
26	269	120	0.15	0.59	837	1.93
27	158	72	0.09	0.35	505	1.15
28	148	68	0.08	0.33	483	1.09
Average	348	154	0.19	0.75	1072	2.49
Std. Dev	145	62	0.08	0.31	436	1.01
Min.	141	65	0.08	0.32	459	1.04
Max.	766	332	0.41	1.63	2 306	5.41

(for the samples collected at 0–20 cm) and

Table 6. Radiological hazard calculations for superficial soil layer (40–60 cm).

Sample Code	Radiological Hazard Indices
	Raeq [Bq/kg]	ADR [nGy/h]	AEDE [mSv/y]	ELCR [10−3]	AGDE [μSv/y]	Iγ
1	4064	1737	2.13	8.50	11,994	28.4
2	2444	1046	1.28	5.12	7228	17.1
3	658	285	0.35	1.40	1978	4.64
4	1412	607	0.74	2.97	4201	9.90
5	1116	482	0.59	2.36	3336	7.84
6	780	339	0.42	1.66	2347	5.50
7	2273	975	1.20	4.77	6739	15.9
8	1759	756	0.93	3.70	5226	12.3
9	1526	657	0.81	3.22	4548	10.7
10	127	59	0.07	0.29	421	0.95
11	3532	1511	1.85	7.39	10,436	24.7
12	499	218	0.27	1.07	1513	3.53
13	723	313	0.38	1.53	2169	5.09
14	822	356	0.44	1.74	2461	5.77
15	902	390	0.48	1.91	2702	6.35
16	1097	473	0.58	2.32	3277	7.70
17	1319	568	0.70	2.78	3934	9.26
18	825	357	0.44	1.75	2471	5.80
19	1065	460	0.56	2.25	3183	7.47
20	1405	603	0.74	2.95	4174	9.85
21	1065	459	0.56	2.25	3181	7.49
22	771	335	0.41	1.64	2320	5.44
23	1469	631	0.77	3.09	4364	10.3
24	957	413	0.51	2.02	2864	6.74
25	312	138	0.17	0.67	962	2.23
26	506	221	0.27	1.08	1539	3.59
27	550	240	0.29	1.17	1665	3.89
28	373	164	0.20	0.80	1147	2.67
Average	1227	528	0.65	2.59	3657	8.61
Std. Dev	892	380	0.47	1.86	2624	6.22
Min.	127	59	0.07	0.29	421	0.95
Max.	4064	1737	2.13	8.50	11,994	28.4

(40–60 cm). The layer selection was made on the basis of radioactivity measurements and analysis of Pearson’s correlation coefficients. The values relating to doses defined as those received by the radionuclides distribution in the superficial soil layer (e.g., ADR or Iγ) refer to a situation in which this layer would be exposed as a result of human activity or weathering processes.

The radium equivalent activity (Raeq) is a key factor in the delineation of regions where elevated radiation doses pose a threat to public health. It is important to note that remaining in areas where the recommended threshold of 370 Bq/kg has been exceeded may lead to the development of cancer [17].

As demonstrated in Table 5, the range of Raeq values for the surface soil layer exhibited a variability from 141 Bq/kg (sample A-9) to 766 Bq/kg (highly contaminated site A-1), with an average value of 348 ± 145 Bq/kg. It was established that the recommended threshold was exceeded at 13 sample sites. Based on the data analysis, it can be concluded that the waste was stored in close proximity to point A-1 and within the lower left sector of the analysed area (Figure 2). In turn, at a depth of 40–60 cm, the range of Raeq was found to vary from 127 Bq/kg (sample A-10) to 4,064 Bq/kg (A-1) with an average value of 1227 ± 892 Bq/kg. Overall, at this depth, the acceptable threshold was exceeded in 26 out of 28 samples. This finding indicates that the exposure of deeper soil layers poses an immediate threat to human health and the environment. Since 370 Bq/kg corresponds to a dose of 1.5 mGy/y, in the vicinity of point A-1, the safe threshold will be exceeded more than 16 times. In the topsoil layer, the largest contribution to Raeq comes from K-40 (range of 54.0–86.5%), while the lowest is attributed to Ra-226 (3.79–7.65%). At a depth of 40–60 cm, the predominant contribution is thorium-232 (accounting for 8.15% to 72.8% of the total value). This phenomenon can be explained by considering the distribution of the waste on the former site, and secondly by its capacity to migrate. Th-232 in the form of ThO₂ or phosphates exhibits a strong adherence to soil particles and therefore is considered non-mobile. In contrast, Ra-226 and K-40 are capable of migrating freely in the environment. Comparison of the Raeq values with data for other locations (Table S6) indicates the significant environmental pollution characteristic of mining areas.

The global ADR range is 55–57 nGy/h, while the safe threshold recommended by UNSCEAR and ICRP is 1000–1500 nGy/h. In the surface layer of the tested soil, the ADR values were in the range of 65–332 nGy/h (average 154 ± 62 nGy/h). In the layer situated at a depth of 40–60 cm, this parameter attained values ranging from 59 to 1737 nGy/h, with an average value of 528 ± 380. Evidently, it has been demonstrated that the global average has been exceeded in both cases. In soil sampled from a depth of 40–60 cm, at three points (A-1, A-2, A-11), the ADR value was found to be greater than the recommended safe range.

The annual (outdoor) effective dose equivalent corresponds to the gamma radiation dose received as a result of existing in a particular location for one year. Globally, AEDE values are known to be less than 0.07 mSv/y, which is widely considered to be a safe limit. The analysis of the topsoil layer (Table 5) revealed that the values obtained at all measurement points were greater than the specified recommendation. The AEDE values vary from 0.08 mSv/y to 0.41 mSv/y (with a mean of 0.19 ± 0.08 mSv/y). A considerable increase in the AEDE values is observed in the 40–60 cm soil layer (0.07–2.12 mSv/y, with an average of 0.65 ± 0.47 mSv/y). With regard to the superficial layer, at 27 out of 28 sampling points, the radiation dose increases from 1.25 times at A-25 to more than 15 times (A-11). This phenomenon may be attributed to the vertical and horizontal migration of radioactive substances present within the tailings. Even at peripheral points in the study area, the AEDE exceeds the suggested standard of 0.07 mSv/y (e.g., for A-28 AEDE = 0.20 mSv/y). Hence, the calculation provides substantiation for the conclusion that exposure to deeper soil layers will be associated with an increased radiological risk.

The ELCR and AGDE radiation hazard indices can be considered in the context of the direct effects of ionising radiation on humans. The former is associated with an elevated probability of carcinogenesis, whereas AGDE serves as an indicator of the potential genetic material damage to human gonads. It is a well-established fact that an elevated value of both parameters is indicative of an increased probability of developing cancer or foetal malformations. In addition, the determination of ELCR and AGDE values is of particular significance in the context of the location of the former landfill site in Ust-Kamenogorsk city. It was observed that in the near-surface layer, the ELCR values in all of the samples taken were greater than the recommended 0.29 × 10⁻³ (ranging from 0.32 to 1.63 × 10⁻³, with an average of 0.75 ± 0.31 × 10⁻³), providing the higher cancer risk. A parallel conclusion can be deduced through the analysis of AGDE changes. Irrespective of the sample collection point (at the depth of 0–20 cm), the soil is characterised by elevated levels of the aforementioned hazard index (459–2306 μSv/y, with an average of 1072 ± 436 μSv/y). At A-9, the recommended value of 300 μSv/y is exceeded by 1.5 times, while at the highly radioactive A-1 site, the AGDE value is greater than the safe threshold by almost 8 times. In samples collected at a depth of 40–60 cm, a substantial increase in ELCR and AGDE results is observed compared to the surface soil layer. The ELCR dose ranged from 0.29 to 8.50 × 10⁻³, with an average value of 2.59 ± 1.86 × 10⁻³. Concurrently, the lowest AGDE values were recorded for sample A-10 (421 μSv/y) and the highest for A-1 (11,994 μSv/y), with an average of 3657 ± 2624 μSv/y. The results obtained from the study indicated significant soil contamination, which consequently gave rise to radiological hazards. The consequences of elevated radiation levels are manifold. These include the potential for a multitude of genetic alterations, which can result in the development of cancerous cells. Furthermore, such levels can have a detrimental effect on the process of gestation, as well as giving rise to a plethora of pathologies that are associated with foetal development.

Iγ refers to the presence of radioactive contaminants in the superficial soil layer. Irrespective of the sampling depth, the value obtained for samples exceeds 1; the sole exception is the A-10 at a depth of 40–60 cm. As demonstrated in Table 5, the gamma index ranged from 1.04 to 5.41 with an average of 2.49 ± 1.01. Furthermore, at a depth of 40–60 cm (Table 6), the Iγ results indicated substantial environmental contamination with NORM radionuclides (0.95–28.36 with an average of 8.61 ± 6.22). It can be concluded that these values confirm a high degree of contamination, predominantly attributable to thorium compounds strongly adhering to soil grains. Consequently, the exposure of deeper soil layers endangers the propagation of radioactive contamination through weathering and erosion, posing a direct radiological threat to surrounding human habitats. It follows that any reclamation work must be carried out in strict accordance with safety procedures.

As can be seen from a comparison of the radiological indices obtained for different locations in the world (see Table 7), the value of a specific dose depends on the human activity profile. It has been observed that an increase in parameter values is significant when the extracted materials are accompanied by uranium or thorium compounds. A pertinent example of the issue in question is the REE mine in Lao Cai, Vietnam, where the calculated hazard factors exceed the recommended values [41]. Conversely, the diamond mining process is associated with a lower contamination by NORM-classified substances [42]. An additional contribution to the emission occurs in the processing of the extracted ores and the subsequent disposal of post-mining waste. A comprehensive analysis of the radiation hazard indicators in the area of former tin tailings and monazite processing site in Kazakhstan, it was concluded that the levels of ionising radiation far exceed the recommended global values. A comparison with other regions of the world with known radioactive contamination led to the conclusion that the study area has a comparable, and in some places even higher, level of radiation risk.

The next stage in the quantitative interpretation of the borehole gamma survey data, together with the laboratory testing of samples, was the geo-mapping of the surveyed area according to specified radioactivity criteria. In order to carry out this stage, the results of the mathematical processing of the data (including radionuclide activity, ADR, gross alpha, and beta activity) were transferred and processed in the MapInfoProfessional Geographic Information System (GIS), resulting in the necessary cartographic materials (see Figure S3). In accordance with the legislation of the Republic of Kazakhstan, the contaminated and reclaimed areas formed as a result of mining and processing enterprises should be classified as “radioactive waste” if the specific activity of radioactive elements contained in them exceeds the minimum significant specific activity value established for each radionuclide (Annex 26 of the State Standard on Radioactive Waste). Another criterion for the classification is the level of specific activity of alpha-emitting radionuclides ≥ 10,000 Bq/kg in accordance with Article 369.4 of the Environmental Code of the Republic of Kazakhstan [36]. In addition, the classification and categorization of waste were based on existing IAEA standards [39,43]. On the basis of the above-mentioned documents, the soils in the area contaminated with wastes from the mechanical concentration of REE ores up to the level of the total specific activity of alpha emitters exceeding 10,000 Bq/kg shall be classified as solid low-level radioactive waste.

Given the source of contamination in the area as well as proximity to buildings and urban infrastructure, its reclamation must be carried out in accordance with both local and international norms (e.g., Environmental Code of the Republic of Kazakhstan No. 400-VI ZRK dated 2 January 2021 [44]). Therefore, the decontamination of the territory should involve the removal of contaminated soil and burying it either in a tailings storage facility, in special burial grounds, or together with waste dumps in reclaimed denudation sites. The removed soil must be replaced with potentially fertile soil that is free of radioactive and chemical contamination. Mining and processing activities produce a lot of waste and by-products that contain natural radionuclides. In the absence of such control, this can lead to the introduction of anthropogenic changes to the radiation background in human habitats and the environment. It is imperative to emphasise the necessity of safeguarding the well-being of personnel engaged in the remediation of contaminated soil, with a view to averting potential harm to human health arising from physical injuries and excessive effective doses of radiation. The subsequent phase in the long-term strategy should entail the continuation of the radioactivity monitoring in the reclaimed area.

Table 7. Comparison of the radiation hazard indices with other places in the world.

Location	Raeq (Bq/kg)	ADR (nGy/h)	AEDE (mSv/y−1)	AGDE (μSv/y)	ELCR × 10−3	Region, Country	Ref.
Granite area	121–624	57.6–280	70.7–344	ns	ns	Xiazhuang, China	[45]
Gold mine	76–104 (control soil area) 1668–2812 (tailing no 3)	37.5–49.1 (control) 203.9–1297.3 (tailing no 3)	0.1–1.6	ns	ns	Gauteng Province, South Africa	[46]
Uranium-rich area	ns	52.7–1722	ns	0.06–2.14	0.21–7.49	Siwalik region, India	[47]
Lomonosovov diamond deposit	17–198	9–94	0.01–0.12	65–675	0.15–1.62	Arkhangelsk, Russia	[42]
Abandoned U mine	160–1746 (for soil, for rhizosphere up to 9494)	78–856 (for soil, for rhizosphere up to 4 811)	0.095–1.05 (27.0 for rhizosphere sample)	ns	0.38–4.19 (10.6 for rhizosphere sample)	Salamanca, Spain	[48]
Rare earth element mine	ns	1597–25,057	1.96–30.7	ns	0.008–0.124	Lao Cai, Vietnam	[41]
Tin tailing	0.6–20.9 (soil) 3.2–17.0 (tailing)	1.7–8.9 × 103 (soil) 1.2–7.2 × 103 (tailing)	ns	ns	1.2–42.0 (soil) 5.9–34.0 (tailing)	Perak, Malaysia	[49]
Pb–Zn deposit	78.9–225	36.8–108	0.045–0.133	256–770	0.158–0.463	Yunnan, China	[50]
Former REE processing site	141–766 (at 0–20 cm) 127–4064 (at 40–60 cm)	65–332 (at 0–20 cm) 59–1737 (at 40–60 cm)	0.08–0.41 (at 0–20 cm) 0.07–2.13 (at 40–60 cm)	459–2306 (at 0–20 cm) 421–11,994 (at 40–60 cm)	0.32–1.63 (at 0–20 cm) 0.29–7.39 (at 40–60 cm)	Kazakhstan	This study
	<370	59.0	0.07	300	0.290	World recommended value	[17]

ns—not specified.

Table 7. Comparison of the radiation hazard indices with other places in the world.

Location	Raeq (Bq/kg)	ADR (nGy/h)	AEDE (mSv/y−1)	AGDE (μSv/y)	ELCR × 10−3	Region, Country	Ref.
Granite area	121–624	57.6–280	70.7–344	ns	ns	Xiazhuang, China	[45]
Gold mine	76–104 (control soil area) 1668–2812 (tailing no 3)	37.5–49.1 (control) 203.9–1297.3 (tailing no 3)	0.1–1.6	ns	ns	Gauteng Province, South Africa	[46]
Uranium-rich area	ns	52.7–1722	ns	0.06–2.14	0.21–7.49	Siwalik region, India	[47]
Lomonosovov diamond deposit	17–198	9–94	0.01–0.12	65–675	0.15–1.62	Arkhangelsk, Russia	[42]
Abandoned U mine	160–1746 (for soil, for rhizosphere up to 9494)	78–856 (for soil, for rhizosphere up to 4 811)	0.095–1.05 (27.0 for rhizosphere sample)	ns	0.38–4.19 (10.6 for rhizosphere sample)	Salamanca, Spain	[48]
Rare earth element mine	ns	1597–25,057	1.96–30.7	ns	0.008–0.124	Lao Cai, Vietnam	[41]
Tin tailing	0.6–20.9 (soil) 3.2–17.0 (tailing)	1.7–8.9 × 103 (soil) 1.2–7.2 × 103 (tailing)	ns	ns	1.2–42.0 (soil) 5.9–34.0 (tailing)	Perak, Malaysia	[49]
Pb–Zn deposit	78.9–225	36.8–108	0.045–0.133	256–770	0.158–0.463	Yunnan, China	[50]
Former REE processing site	141–766 (at 0–20 cm) 127–4064 (at 40–60 cm)	65–332 (at 0–20 cm) 59–1737 (at 40–60 cm)	0.08–0.41 (at 0–20 cm) 0.07–2.13 (at 40–60 cm)	459–2306 (at 0–20 cm) 421–11,994 (at 40–60 cm)	0.32–1.63 (at 0–20 cm) 0.29–7.39 (at 40–60 cm)	Kazakhstan	This study
	<370	59.0	0.07	300	0.290	World recommended value	[17]

ns—not specified.

4. Conclusions

The gamma radiation measurement revealed dose levels that were significantly elevated (up to 1.00 µSv/h), which exceeded the natural background (0.16–0.18 µSv/h). This finding indicates a radiological anomaly is associated with the former processing of thorium ores and RRE. Vertical profiling revealed an increase in dose rates with depth, with a maximum recorded at 80–100 cm. Furthermore, the identification of ‘hot spots’ (e.g., A-1, A-11, A-21) indicated an uneven distribution of radioactive waste.

The measurement of gross alpha and beta activity has been shown to be an effective method of confirming the presence of contamination. The obtained results demonstrate a clear correlation between the increase in alpha and beta emitter activity and depth, with the greatest increases observed in the 40–100 cm range. For the topsoil layer, the total radionuclide activity was in the range of 583–5275 Bq/kg and 641–1749 Bq/kg for alpha and beta, respectively. However, at the depth of 80–100 cm, the measured values were increased to 725–22,482 Bq/kg (for gross alpha) and 729–6845 Bq/kg for the β-emitters. It has been shown that the obtained values exceed the background levels and correspond to contaminated industrial sites. It is imperative to acknowledge that the surface layer (0–20 cm) poses a significant risk, as the radionuclides have the capacity to penetrate plants and animals, subsequently migrating with surface water.

A multivariate correlation analysis reveals that significant long-term anthropogenic contamination has been identified, thus justifying the necessity of ongoing monitoring, risk assessment, and consideration of remediation measures. A strong positive correlation was found between alpha and beta radioactivity, particularly at a depth of 40–60 cm. The maximum radionuclide concentrations are observed at a depth of 40 cm, suggesting the vertical migration from the surface.

The identified exceedances of the permissible radiological risk levels clearly indicate the need for remedial action. It is recommended that the area be the subject of continuous radiological monitoring. Subsequently, there is a necessity for the contaminated soil to be removed in an appropriate and secure manner. It is indisputable that concerted efforts should be implemented to curtail the propagation of contamination and to ensure the preservation of the environment and the well-being of the local population.

Regarding hypotheses, it can be concluded that the deeper soil layers contain significantly higher levels of radioactivity than the surface layers. The reasons may be vertical migration and the storage of post-production waste. Analysis of soil samples in contact with aqueous solutions revealed that radionuclides are tightly bound to soil particles; only the ion-exchangeable form passes into the aqueous phase. Therefore, it can be stated that the radioactive isotopes cannot penetrate water when in contact with soil. The third hypothesis was confirmed, resulting in a higher risk to both the local population and the environment.

Further research directions will include the analysis of radionuclide content in groundwater, underground water, and plants found in the area. In order to determine migration mechanisms, transfer coefficients will be calculated. The authors also intend to analyse radionuclides in atmospheric aerosol samples. A significant constraint pertains to the temporal framework of measurements, which consequently restricts the number of samples and repetitions (in order to ensure reliable results). The resolution of this issue will be achieved through intensive collaboration between two or more research centres equipped with modern research facilities. This approach will facilitate a comprehensive analysis of the area in question, allowing for the drawing of conclusions that could be applied to other regions contaminated by human activity. The measures delineated above are in perfect alignment with the overarching sustainable development programme of developing countries, including Kazakhstan. The broad perspective and final conclusions resulting from the analyses will enable the development of a strategy for dealing with cases of uncontrolled release of radioactive substances into the environment, resulting in improved quality of life and health protection for the local population.