Modelling the Uptake of ²²⁶Ra and ²³⁸U Stable Elements in Plants during Summer in the Vicinity of Tailings from an Abandoned Copper Mine

Abstract

Transfer factors (TFs) are widely used tools for assessing the uptake of radionuclides by plants. The literature contains numerous studies on TFs in tropical and temperate climates; however, the existing data on TFs in arid and semi-arid climates are very scarce. Furthermore, the current trend in nuclear energy expansion in countries with this type of climate necessitates knowledge of the mechanisms of radionuclide incorporation by plants as well as the TF values. For this reason, this work investigates the TFs of ²³⁸U and ²²⁶Ra in plants in a study area during the summer period under conditions equivalent to a semi-arid climate. The selected plants were Scolymus hispanicus L., Eryngium campestre L., Chenopodium vulvaria L., and Chenopodium album L., which were collected in the vicinity of a waste dump from an abandoned copper mine. The selected study area has radionuclide levels above the global average, in addition to heavy metals, as it is a waste dump from an abandoned copper mine. The range of transfer factors for ²³⁸U varied between 1.5 × 10⁻⁴ kg⁻¹ kg⁻¹ and 7.8 × 10⁻³ kg⁻¹ kg⁻¹, while for ²²⁶Ra, the range was between 1.8 × 10⁻⁴ kg⁻¹ kg⁻¹ and 4.0 × 10⁻² kg⁻¹ kg⁻¹. The correlations found with PCA were (i) ²³⁸U with Fe and Al, and (ii) ²²⁶Ra with S, Ti, Ca, and Sr. A transfer model of ²³⁸U and ²²⁶Ra was created using multiple linear regression analysis. The model showed how ²³⁸U was related to the presence of Al, while ²²⁶Ra was related to Al, Fe, and Ti. The results obtained have allowed us to propose a model for the incorporation of ²³⁸U and ²²⁶Ra, taking into account the chemical composition of the soil. The results obtained indicate that both Scolymus hispanicus L. and Eryngium campestre L. could be utilized in phytoremediation for soils contaminated by natural radionuclides in semi-arid climates. The TFs, as well as the proposed model, allow us to expand the knowledge of the absorption of natural radionuclides by plants in regions with arid and semi-arid climates, which is necessary for the radiological risk assessment of future nuclear fuel cycle facilities.

1. Introduction

The uptake of radionuclides such as ²³⁸U and ²²⁶Ra by plants is a well-studied aspect of the literature [1,2,3,4,5]. The parameter used to evaluate the availability and uptake of radionuclides by plants is the transfer factor [6]. This parameter is defined as the dry mass ratio of the activity concentration of a given radionuclide in a plant to its content in the soil. The values reported in the literature for ²³⁸U and ²²⁶Ra range widely from 10⁻⁴ to 1 kg⁻¹ kg⁻¹ [6,7,8]. The high variability in transfer factors is mainly because they are a very conservative parameter, as they do not take into account the interactions of radionuclides with the physicochemical properties of soil [9,10]. In relation to this, there are few studies that relate chemical composition regarding the concentration of stable elements with unstable or radioactive ones such as ²²⁶Ra and ²³⁸U [11]. Rea et al. [12] studied the uptake of radionuclides from the uranium and thorium radioactive series considering the presence of stable elements; however, the study did not relate the uptake of stable elements with the natural radionuclides. Similarly, Vera-Tomé et al. [13] found no correlation between the stable elements and the natural radionuclides. Therefore, more studies are needed on this possible relationship in the uptake of stable and radioactive elements by plants.

On the other hand, the study of the uptake of both naturally occurring and artificial radionuclides by plants in semi-arid climates is increasing in importance due to the absence of transfer factors for these types of climates [14]. The soils typical of these climates possess characteristics that distinguish them those of from tropical climates and affect plant absorption. The primary characteristic is a lack of moisture, mainly due to the texture and absence of organic matter, which results in high-level water stress for the plants [14,15]. Additionally, temperatures cause precipitation to evaporate quickly, leading to radionuclides being retained in the surface layer of the soils, making them less accessible to the roots [1]. These characteristics imply that fewer radionuclides are incorporated than those in the soils of tropical climates, which have a greater presence of organic matter and moisture. The need for this knowledge lies in the possible future construction of nuclear power plants in countries with this type of climate in the near future. The literature includes many studies of this type of climate for different radionuclides, both naturally occurring and anthropogenic, and for different types of soils, including in areas impacted and not impacted by uranium mining [1,16,17,18,19]. However, these studies only report transfer factors, without finding a relationships between the studied radionuclides and the stable elements. The transfer factors are used to assess the potential incorporation of radionuclides through the food chain. The initial step involves the ingestion of plants by livestock, which may subsequently be consumed by humans, potentially increasing the risk of cancer [20].

The objective of this study was to investigate the potential relationship between the uptake of two major natural radionuclides, ²³⁸U and ²²⁶Ra, by plants and stable elements in a semi-arid climate. The necessary conditions for this study were provided by the waste heap and the surroundings of an abandoned Cu mine during the summer period, where the absence of rainfall and high temperatures resembled a semi-arid climate zone. Our working hypothesis was that there is a relationship between the uptake of ²³⁸U and ²²⁶Ra and stable elements under semi-arid climate conditions. To test our hypothesis, the partial objectives were (i) the radiological and chemical characterization of the soils in the study area, (ii) the determination of the transfer factors of ²³⁸U and ²²⁶Ra, along with those of the stable elements in four types of plants (Scolymus hispanicus L., Eryngium campestre L., Chenopodium vulvaria L., and Chenopodium album L.), and (iii) the statistically relation of the transfer factors through principal component analysis, obtaining absorption models through multiple linear regression analysis.

2. Materials and Methods

2.1. Study Area and Plants Analyzed

The waste heap of the “Antigua Pilar” mine is located in the municipality of Colmenarejo in the western part of the Community of Madrid (Spain) (Figure 1). The area is situated at an elevation between 780 and 815 m, with an average temperature of 13.4 °C and an annual precipitation average of 637 mm. The soils in the Colmenarejo area are classified as cambisols formed from granites, gneisses, mica schists, and quartz [21]. Summer temperatures can reach up to 35 °C, which combined with the absence of precipitation, gives the climate a semi-arid character.

The study area was the abandoned waste heap of the “Antigua Pilar” copper mine in the municipality of Colmenarejo in the Community of Madrid (Spain). The mine’s waste heap mainly contains the following minerals: chalcopyrite (CuFeS₂), arsenopyrite (FeAsS), malachite (Cu₂CO₃(OH)₂), azurite (Cu₃(CO₃)₂(OH)₂), pyrite (FeS₂), siderite (FeCO₃), fluorite (CaF₂), olivenite (Cu₂AsO₄(OH)), and chalcanthite (CuSO₄·5H₂O) [C1196]. In addition to these minerals, the mine’s waste heap contains torbernite (Cu(UO₂)₂(PO₄)₂·8–12H₂O) and zeunerite (Cu(UO₂)₂(AsO₄)2·12H₂O), whose presence is caused by geological formation through hydrothermal processes [22]. Therefore, the samples from the waste heap and its surroundings have a ratio of the activity concentration of uranium and thorium radioactive series that is significantly greater than 1 [23]. Figure 2a–c depicts the mine tailings pile and a detailed view of the minerals present. The four plant species used in this study were selected based on their relevance to the following aspects: (i) Scolymus hispanicus L. (9 plants) for its use in phytoremediation by absorbing numerous heavy metals [24]; (ii) Eryngium campestre L. (5 plants) for the uptake of radionuclides such as ¹³⁷Cs by mycorrhizal fungi associated with its roots [25,26]; (iii) and Chenopodium vulvaria L. (7 plants) and (iv) Chenopodium album L. (4 plants) due to their use as fodder in livestock feed as well as their medicinal applications as anti-inflammatory agents [27]. The 25 sampled plants were collected from the sampling points shown in Figure 1.

2.2. Sampling and Preparation of Soil and Plant Samples

The plants were selected based on their location within the study area. The plants were identified by collecting the entire aerial part of each plant, which was placed in a paper bag to avoid alterations during transport to the laboratory. The roots were not used to prevent errors in the interpretation of the results due to the presence of adhered soil, which is very complex to remove [13]. Subsequently, a surface soil sample near the plant was taken using a metal grid measuring 20 cm × 20 cm and 5 cm deep (Figure 2d) until a total of 1.0 kg of soil was collected. The samples were placed in plastic bags. The plant sample and its corresponding soil were located using a GPS. The samples were taken to the laboratory, where they were air-dried for one day.

Each soil sample was then placed in a Selecta brand oven, model 2000209 (Spain), and dried at 105 °C for 24 h until a constant weight was achieved. Next, the samples were ground in an FRITSCH Pulverisette 5 planetary ball mill (Figure 2e). Subsequently, the samples were sieved using a 120 mesh sieve, achieving a particle size of 250 µm.

The plant samples were dried in an HOBERSAL oven model 2003741 (Spain) at 65 °C for 3 days, and subsequently ashed in an HOBERSAL muffle furnace model HD150PA at 450 °C for 24 h (Figure 2f).

The soil and plant samples were mineralized in a microwave oven (Milestone, ultraWAVE, Italy) (Figure 2e). Three aliquots of 0.2 g of sample were introduced into 3 Teflon vials to which 1 mL of HF, 1 mL of HNO₃, and 4 mL of HCl were added. The three aliquots obtained after mineralizing each sample were added to a Teflon beaker and evaporated to dryness. Subsequently, the HF was removed by dissolving the residue with 3 mL of concentrated HCl and evaporating the solution to dryness again. This step was repeated twice. The tracers used to determine the chemical yield of the radiochemical methods were (i) ²²⁹Th solution with an activity concentration of 20.96 ± 0.16 Bq g⁻¹ (supplied by the National Laboratory of Metrology of Ionising Radiations (LMRI) of the Centre for Energy, Environmental and Technological Research (CIEMAT), Madrid, Spain) and (ii) 232U solution of 13.54 ± 0.16 Bq g⁻¹ supplied by the LRMI. The Ba²⁺ carrier had a concentration of 10 mg mL⁻¹ and was prepared from BaCl₂·2H₂O (Merck, Germany) with analytical grade purity. The tracers were diluted with HNO₃ 1:100 (v/v) to obtain the required activity concentration for the radiochemical methods employed.

2.3. Determination of the Chemical Composition of Soil and Plant Samples

The chemical composition of the soil and plants was determined using a Malvern-PAN-analytical AXIOS spectrophotometer. The samples were prepared using a planetary ball mill, and the samples prepared in Section 2.2 were again sieved to a particle size of 74 µm with a 200 mesh sieve. The determination of the concentration of stable cations by WDXRF was performed using a semi-quantitative method developed by Malvern-Panalytical. This method uses fundamental parameters for quantification, with a minimum of one standard per element that ensures an accuracy of at least 1%. The use of this method allows for the same aliquot to be used for subsequent radiochemical analyses, also avoiding its handling. For this purpose, the samples were placed in a special container that permits the analysis of the sample in a He atmosphere. The WDXRF technique is very stable and does not require recalibration for each sample measured since the standard deviation is very low. On the other hand, although WDXRF is capable of quantifying the uranium present in the samples, these results have not been used as they are much less sensitive and reproducible than those determined using radiochemical methods.

2.4. Radiochemical Methods for the Separation of ²³⁸U and ²²⁶Ra

2.4.1. ²³⁸U Separation Method

The radiochemical separation method for ²³⁸U consisted of a liquid–liquid extraction with TBP dissolved in xylene in a 1:4 (v/v) ratio. Prior to extraction, 0.25 M EDTA solution was added to mask Th and the other cations that were co-extracted in TBP. Finally, U was re-extracted using 1 M (NH₄)₂CO₃ solution. The separated ²³⁸U was electrodeposited onto a stainless steel planchet using the Hallstadius method [28] (Figure 2f). The chemical yield of the method was determined from the ²³²U tracer described in Section 2.2. The different steps of the method used, as well as the expressions employed for the determination of the activity concentration, uncertainty, and limit of detection (LoD) of ²³⁸U, are detailed in [29]. The samples were measured for 500,000 s, achieving an LoD of 5·10⁻³ Bq kg⁻¹ for the plant samples and 1.2 Bq kg⁻¹ for the soil samples.

2.4.2. ²²⁶Ra Separation Method

The method used for the radiochemical separation of ²²⁶Ra was based on the co-precipitation of Ra in a BaSO₄ precipitate through the formation of mixed crystals. The method began with the formation of a [Pb-Ba-Ra]SO₄ precipitate, which was dissolved with EDTA in a strongly ammoniacal medium. The [Ba-Ra]SO₄ was re-precipitated with CH₃COOH and transformed into [Ba-Ra]CO₃ using saturated Na₂CO₃ solution at 90 °C. Subsequently, [BaRa]Cr₂O₇ was precipitated to eliminate the co-precipitated alkaline earth elements (Ca, Mg, and Sr). [BaRa]Cr₂O₇ was transformed into [BaRa]Cl₂ with a HCl + CH₃CH₂OCH₂CH₃ mixture in a 1:6 (v/v) ratio. [BaRa]Cl₂ was dissolved in H₂O, and [BaRa]SO₄ was precipitated with 1 M H₂SO₄. BaSO₄ was filtered, deposited on a cellulose nitrate filter with a pore size of 0.45 µm, and dried to constant weight under an infrared lamp. The chemical yield of the method was determined gravimetrically from the weight of the final BaSO₄ obtained. A more extensive description of the method, as well as the expressions used to determine the activity concentration, uncertainty, and LoD, is detailed in [30]. The LoD obtained for the plant samples was 0.006 Bq kg⁻¹, and it was 0.7 Bq kg⁻¹ for the soil samples.

2.4.3. Equipment for the Measurement of ²³⁸U and ²²⁶Ra

The stainless steel planchets with electrodeposited ²³⁸U were measured using an AlphaAnalyst model A450-18AM (Mirion Canberra, Bretonneux, France) system equipped with 12 vacuum chambers with 12 PIPS (passivated implanted planar silicon) detectors from Canberra Industries (Figure 2f). Each detector had an active area of 450 mm² and a resolution of 18 keV full width at half maximum (FWHM) at 5.486 MeV for ²⁴¹Am. The spectra were acquired and analyzed using Genie 2000 software.

The [BaRa]SO₄ precipitates were measured using 6 ZnS(Ag) detectors. These detectors detect the light emitted when an alpha particle interacts with the ZnS(Ag) deposited on one side of a transparent plastic. This plastic was placed over the filter containing the [BaRa]SO₄ precipitate. The detection of the scintillations was carried out with a photomultiplier tube (Canberra model 2007P) whose signal was amplified with a Canberra model 2007A amplifier, which, in turn, was powered by a Canberra model NHQ 202M high voltage source.

2.5. Determination of Transfer Factors

The transfer factors (TFs, kg kg⁻¹) were determined as the ratio between the activity concentration or chemical concentration in the dry mass of a radionuclide or stable element in the plant and in the soil using the following expression [6]:

$$TF={\displaystyle \frac{C_{plant(d.m.)}}{C_{soil(d.m.)}}}$$

(1)

where $C_{plant(d.m.)}$ is the activity concentration of a radionuclide (²³⁸U and ²²⁶Ra) or a stable element in the plant (dry mass), and $C_{soil(d.m.)}$ is the activity concentration of a radionuclide (²³⁸U and ²²⁶Ra) or a stable element in the soil (dry mass). The uncertainty associated with the transfer factor (u(TF), Bq kg⁻¹) is determined by the following expression:

$$u\left(TF\right)=TF·\sqrt{{\left({\displaystyle \frac{u\left(C_{plant(d.m.)}\right)}{C_{plant(d.m.)}}}\right)}^2+{\left({\displaystyle \frac{u\left(C_{soil(d.m.)}\right)}{C_{soil(d.m.)}}}\right)}^2}$$

(2)

2.6. Statistical Analysis of Results

Various statistical analyses conducted in this study were carried out using the RStudio statistical software version 2024.04.2 Build 764, utilizing the following libraries: stats, car, writexl, psych, FactomineR, and Factoextra. The statistical analyses performed are summarized below.

Firstly, principal component analysis (hereafter, PCA) was conducted, through which the samples were grouped based on variables with an influence above the average value of the entire set of variables. Additionally, it was verified that the selected variables had a Kaiser–Meyer–Olkin parameter value above 0.5 [31].

Subsequently, multiple linear regression analysis (MLRA) was performed to verify that the variables obtained in PCA had statistical significance and to obtain a model relating the incorporation of ²³⁸U and ²²⁶Ra with the incorporation of stable elements. The variables used in the models did not exhibit collinearity among them, a factor that was verified using the Variance Inflation Factor (VIF). Finally, the weight of the model parameters allowed us to draw conclusions from this study, selecting those with the highest values of the standardized coefficients as the most influential. The standardized coefficients were determined by transforming the model coefficients using their standard deviation [23].

3. Results and Discussion

3.1. Chemical Composition of Soil Samples

Figure 3 shows box-and-whisker plots, illustrating the concentrations of different stable elements in the soil samples. The data indicate that SiO₂ is the principal compound, followed by these other major elements: Al, Ca, Fe, K, P, and Mg. Cu, Mn, S, and Ti were the predominant elements with higher percentages. The concentrations of the major elements, namely O, Si, Al, Ca, Fe, K, P, and Mg, were equivalent to those typically found in soil samples [32]. Cu had an average concentration of 1600 mg kg⁻¹, which is significantly higher than the European average value of 35.3 mg kg⁻¹ in unimpacted soils [33] and 496.3 mg kg⁻¹ in soils contaminated by pesticides [34]. However, the levels found were consistent with those in other areas affected by Cu mining, where a value of 1600 mg kg⁻¹ was also obtained [35]. Sulphur also showed values well above the average in soils of 24.6 mg kg⁻¹. On the other hand, the As value was consistent with that obtained in other studies on soils affected by U mining, whose average value of 233 mg kg⁻¹ is equivalent to the one obtained in this work of 303 mg kg⁻¹ [36]. The elements identified were consistent with the mineral species characteristic of a copper mine tailings site. On the one hand, Cu is associated with chalcopyrite (CuFeS₂), bornite (Cu_1.8Fe_1.2S₄), and chalcocite (Cu₂S) [37]. Mn is a part of pyrolusite (MnO₂); Fe is found in hematite (Fe₂O₃), pyrite (FeS₂), and magnetite (Fe₃O₄); and Ti is found in rutile or ilmenite (FeTiO₃) [38]. The trace elements found were also consistent with those typically encountered in such tailings [39]: (i) As as arsenopyrite (FeAsS), (ii) Zn as sphalerite (ZnS), (iii) Ba as barite (BaSO₄) and witherite (BaCO₃), and Cr due to the presence of chromite (FeCr₂O₄).

3.2. Activity Concentration Values of ²³⁸U and ²²⁶Ra

Table 1. The activity concentrations of ²³⁸U and ²²⁶Ra for the 25 plants and soils together with the soil–plant transfer factors.

		Soil (Bq kg−1)		Plant (Bq kg−1)		Transfer Factors (kg kg−1)
Species	Sample	238U	226Ra	238U	226Ra	238U	226Ra
Scolymus Hispanicus L.	1	79.3 ± 3.7	113 ± 13	0.232 ± 0.067	4.03 ± 0.50	0.00293 ± 0.00086	0.0357 ± 0.0060
	2	107.6 ± 4.6	161 ± 20	0.232 ± 0.058	1.40 ± 0.21	0.00216 ± 0.00055	0.0087 ± 0.0017
	3	247.5 ± 9.1	222 ± 23	0.73 ± 0.11	3.93 ± 0.45	0.00295 ± 0.00046	0.0177 ± 0.0027
	4	68.6 ± 2.9	117 ± 13	0.402 ± 0.073	0.73 ± 0.14	0.0059 ± 0.0011	0.0062 ± 0.0014
	5	70.6 ± 2.8	125 ± 15	0.196 ± 0.059	1.34 ± 0.19	0.00278 ± 0.00084	0.0107 ± 0.0020
	6	432 ± 15	735 ± 88	0.332 ± 0.076	3.07 ± 0.38	0.00077 ± 0.00018	0.00418 ± 0.00072
	7	70.2 ± 3.1	122 ± 13	0.101 ± 0.053	1.55 ± 0.23	0.00144 ± 0.00076	0.0127 ± 0.0023
	8	776 ± 27	904 ± 97	0.467 ± 0.094	3.55 ± 0.52	0.00060 ± 0.00012	0.00393 ± 0.00071
	9	432 ± 15	537 ± 65	0.153 ± 0.095	1.57 ± 0.22	0.00035 ± 0.00022	0.00292 ± 0.00054
Eryngium Campestre L.	10	65.0 ± 2.5	93 ± 11	0.328 ± 0.062	0.92 ± 0.18	0.00505 ± 0.00097	0.0099 ± 0.0023
	11	58.9 ± 2.3	64.6 ± 7.3	0.071 ± 0.027	1.17 ± 0.17	0.00121 ± 0.00046	0.0181 ± 0.0033
	12	54.5 ± 2.3	65.1 ± 8.3	0.091 ± 0.030	1.12 ± 0.16	0.00167 ± 0.00055	0.0172 ± 0.0033
	13	55.9 ± 2.3	68.1 ± 8.8	0.143 ± 0.044	0.80 ± 0.12	0.00256 ± 0.00079	0.0117 ± 0.0023
	14	47.0 ± 2.0	55.6 ± 6.9	0.064 ± 0.033	2.20 ± 0.28	0.00136 ± 0.00070	0.0396 ± 0.0070
Chenopodium Vulvaria L.	15	44.6 ± 2.3	59.3 ± 6.9	0.35 ± 0.11	0.093 ± 0.032	0.0078 ± 0.0025	0.00157 ± 0.00057
	16	61.5 ± 2.7	98 ± 11	0.0189 ± 0.0075	2.89 ± 0.72	0.00031 ± 0.00012	0.0295 ± 0.0081
	17	71.4 ± 3.1	87 ± 10	0.059 ± 0.057	0.016 ± 0.012	0.00083 ± 0.00080	0.00018 ± 0.00014
	18	66.7 ± 3.2	105 ± 12	0.181 ± 0.069	2.97 ± 0.58	0.0027 ± 0.0010	0.0283 ± 0.0064
	19	60.1 ± 3.1	73.1 ± 7.4	0.112 ± 0.058	0.63 ± 0.26	0.00186 ± 0.00097	0.0086 ± 0.0037
	20	207.2 ± 7.9	167 ± 23	0.104 ± 0.073	1.07 ± 0.78	0.00050 ± 0.00035	0.0064 ± 0.0048
	21	202.3 ± 7.7	315 ± 37	0.068 ± 0.052	0.37 ± 0.18	0.00034 ± 0.00026	0.00117 ± 0.00059
Chenopodium Album L.	22	90.3 ± 3.8	100 ± 10	0.070 ± 0.062	0.54 ± 0.35	0.00078 ± 0.00069	0.0054 ± 0.0035
	23	71.4 ± 3.1	87 ± 10	0.0110 ± 0.0057	1.03 ± 0.79	0.000154 ± 0.000080	0.0118 ± 0.0092
	24	62.8 ± 2.9	88 ± 10	0.0095 ± 0.0017	0.22 ± 0.11	0.000151 ± 0.000028	0.0025 ± 0.0013
	25	94.7 ± 3.9	101 ± 11	0.069 ± 0.043	0.56 ± 0.27	0.00073 ± 0.00046	0.0055 ± 0.0027

The uncertainties are quoted for a coverage factor k = 2.

shows the activity concentration results of ²³⁸U and ²²⁶Ra in the 25 sampled soils, whose values were higher than the average value for Spanish soils of 30 Bq kg⁻¹ [40]. The elevated presence of ²³⁸U and ²²⁶Ra is due to the hydrothermal formation of the study area, which caused a significant increase in uranium radioactive series [41]. This presence is primarily due to the presence of torbernite (Cu(UO₂)₂(PO₄)₂·8–12H₂O) in the area, as described in Section 2.1. Figure 4 shows the activity concentration of ²²⁶Ra as a function of the activity concentration of ²³⁸U. The ratio of ²²⁶Ra/²³⁸U was obtained from the slope of the linear equation, with a value of 1.237 ± 0.065. The obtained ratio showed that U(IV) is oxidized by external agents into UO₂²⁺, which is more soluble than ²²⁶Ra and is thus released into the environment through rain and air [42].

3.3. Relationship between the Chemical Composition of Soils and the Activity Concentration of ²³⁸U and ²²⁶Ra

Figure 5 shows a biplot graph with principal component analysis (PCA) that allows for visualization of the correlation between the activity concentration of ²³⁸U and ²²⁶Ra and the chemical composition and scoring of the different soils based on these variables. The two factors obtained represent 90.4% of the variance, indicating statistical dependence between the different variables. The Kaiser–Meyer–Olkin (KMO) index value was 0.79, which indicates a satisfactory correlation [43]. Some variables were eliminated as they did not exceed the average contribution of the set of variables or 1/(number of variables). The correlation between the different variables is obtained from the cosine of the angle they form with each other. The results showed a correlation between the activity concentrations of ²²⁶Ra and ²³⁸U with Sn, Fe, and Cu. This results indicate that the radionuclides were associated with areas with higher mineral content from the waste heap. Soils 8, 6, 3, 9, 2, and 21 are those that would have the highest content of natural radionuclides, with soil 8 standing out. These soils are located in the waste heap and the mine loading area, which, as verified in a previous study, had the highest radioactive content [23]. The remaining stable elements did not show any correlation with ²³⁸U or ²²⁶Ra. However, coherent correlations are observed, such as those of Si with O (due to SiO₂) and Na and K, which have the same chemical behavior. Additionally, a correlation was observed between Ca, Sr, and S, possibly due to the presence of sulfates, an aspect that would require further analysis to confirm. However, ²²⁶Ra is not correlated with Ca, Sr, and S, indicating that its presence is more commonly associated with U, its progenitor. Therefore, the correlation between U and Fe is consistent with the findings of other authors such as Ha et al. [44], who found that Fe-rich minerals can reduce heavy metals like U from UO₂²⁺, the most soluble species, to U(IV), which is less soluble. Furthermore, the subsequent formation of Fe oxyhydroxides such as ferrihydrite would favor the release of U under high-temperature conditions, which is consistent with the meteorological conditions of the study area [45,46]. The presence of carbonate would also favor the release of U from the oxyhydroxides [47], which, as verified in a previous study, is a significant fraction in the tailings [48]. Therefore, the formation of oxyhydroxides would explain the ²³⁸U/²²⁶Ra ratio shown in Figure 3.

Another observed aspect was the qualitative relationship between the presence of As and ²³⁸U, as they are usually associated in this type of deposit [49]. However, the data did not show any correlation. Therefore, the presence of As would only indicate that the samples were taken from the tailings and the mineral loading area (soils 1, 2, 3, 5, 7, 8, 9, 17, 22, 23, and 25).

3.4. Transfer Factors of ²³⁸U and ²²⁶Ra between Soil and Studied Plants

Figure 6 shows the transfer factors (TF) of ²²⁶Ra and ²³⁸U for the four species studied. On the other hand, Figure 6 presents a biplot graph in which the TFs for ²²⁶Ra, ²³⁸U, and the stable elements (K, Mg, Fe, Al, Na, S, Ti, Ca, and Sr) are represented as variables. The range of transfer factors obtained for ²³⁸U varied between 1.5·10⁻⁴ kg⁻¹ kg⁻¹ and 7.8·10⁻³ kg⁻¹ kg⁻¹, while for ²²⁶Ra, the range was between 1.8·10⁻⁴ kg⁻¹ kg⁻¹ and 4.0·10⁻² kg⁻¹ kg⁻¹. The ranges obtained for the transfer factors of ²³⁸U and ²²⁶Ra were similar to those observed in semi-arid climates [12,18,50,51]. Furthermore, these transfer factors were also comparable to those found in mine tailings present in semi-arid climates [16]. Moreover, the transfer factors followed a logical pattern, with those obtained for ²²⁶Ra being higher than those for ²³⁸U [52]. The elements represented in the graph were those that exceeded the average contribution, as shown in the biplot graph of Figure 4. The KMO index was 0.54, which is low, but represents 57.3% of the variance and is higher than the value of 0.5 necessary to obtain statistically representative results [31]. The TFs obtained for ²²⁶Ra and ²³⁸U are consistent with those obtained by other authors in areas impacted and not impacted by U mining, which ranged between 2·10⁻⁴ and 0.1 kg kg⁻¹, whose characteristics are equivalent to those of the samples analyzed in this study [7,8,31,53,54].

The biplot graph clearly shows four data sets directly related to the species studied (Figure 7). The species Chenopodium vulvaria L. and Chenopodium album L. are the ones least related to the radioactive content of the soils, as the angle formed by the variables ²³⁸U and ²²⁶Ra is 90°, indicating no correlation. These plants are associated with the incorporation of K and Mg. K is often associated with a greater presence of mycorrhizal fungi, as confirmed by other authors such as El-Mesbahi et al. [55]. The presence of mycorrhizal fungi is, in turn, associated with less U absorption, as demonstrated in the previous studies [56]. Similarly, the incorporation of Mg is related to the reduced incorporation of ¹³⁷Cs and ²²⁶Ra due to the competition of Mg for accessible positions in the soil [57]. The reduced incorporation of K and Mg by Scolymus hispanicus L. and Eryngium campestre L. imply a greater incorporation of ²²⁶Ra mainly. On the other hand, the presence of Ti is related to the absorption of ²²⁶Ra, as it favors its mobility; thus, increasing the incorporation of Ti would increase that of ²²⁶Ra [58]. The results also show how the incorporation of ²²⁶Ra is related to the incorporation of Ca and Sr due to their similar behaviors. However, as previously mentioned, Mg, although it is an alkaline earth metal, competes with these cations. Therefore, the incorporation of ²²⁶Ra is coherent from the perspective of Mg, Ca, and Sr. Regarding U, the results showed that its content is greater in the case of Scolymus hispanicus L. and this is associated with the incorporation of other heavy metals such as Fe and Al. The relationship between U, Fe, and Al was observed in the previous studies of granites [59], and it is consistent with the type of soil in the study area (see Section 2.1). Finally, the incorporation of ²³⁸U and ²²⁶Ra did not show any correlation with each other, as the angle of these two variables is practically 90°. This effect is consistent with the TFs obtained for these two elements (Figure 4). The results obtained show that both Scolymus Hispanicus L. and Eryngium Campestre L. are the plants that absorb the most ²³⁸U and ²²⁶Ra, along with other heavy metals such as Fe and Al. Scolymus Hispanicus L. has the ability to alter the soil as its roots generate organic acids and sugars that lead to an increase in pH, the formation of chelates with heavy metals, and an increase in the presence of microorganisms [24]. Eryngium Campestre L. exhibits a high level of mycorrhization in its roots, which also promotes the greater mobility of ²³⁸U and ²²⁶Ra, as well as heavy metals, enhancing their incorporation into the plant [25]. Although future research is necessary to validate these hypotheses, the correlation with sulfur present in the composition of minerals found in a copper tailings site (see Section 3.1.) indicate the possible decomposition of the sulfides present. Therefore, these two plants could be candidates for use in phytoremediation.

3.5. General Model of Transfer of ²³⁸U and ²²⁶Ra Based on the Transfer of Stable Elements

The transfer models of ²³⁸U and ²²⁶Ra obtained through multiple linear regression analysis have the following equations:

$${TF}_{{238}_U}=\left(1.84\pm 0.49\right)·{TF}_{Al}+\left(0.00024\pm 0.0001\right)·{TF}_S-\left(0.00066\pm 0.00069\right)$$

(3)

$$\begin{gathered} {TF}_{{226}_{Ra}}=\left(10.0\pm 4.6\right)·{TF}_{Al}+\left(0.0036\pm 0.0014\right)·{TF}_{Ca}-\left(4.2\pm 2.0\right)·{TF}_{Fe}+\left(0.016\pm 0.010\right)·{TF}_{Mg}-\cdots \\ \cdots -\left(0.62\pm 0.30\right)·{TF}_{Na}+\left(1.72\pm 0.84\right)·{TF}_{Ti}+\left(0.0032\pm 0.0074\right) \end{gathered}$$

(4)

The residuals obtained through the proposed models for ²³⁸U and ²²⁶Ra were evaluated using RSE and RMSE statistics, yielding 1.5·10⁻⁴ and 1.5·10⁻⁴ for ²³⁸U and 8.8·10⁻³ and 7.4·10⁻³ for ²²⁶Ra. Therefore, the values estimated by the models yield consistent results. The models were simplified by eliminating the parameters with a p-value greater than 0.05 in the analysis of variance. The weight of the most influential factors in the absorption of ²³⁸U and ²²⁶Ra was determined by examining the standard coefficients of the fit. The parameter that had the greatest weight in the fit for ²³⁸U was Al, while for ²²⁶Ra, it was Al >> Fe > Ti. The interaction observed between ²³⁸U and Al may be equivalent to that found by Bachmaf and Merkel [60], who observed an interaction between Al and UO₂²⁺ in clay samples due to available sites on the surface where Al was present. This same effect was found by Yamaguchi et al. in clays [61]. Although the interactions between ²³⁸UO₂²⁺ and ²²⁶Ra²⁺ with Al³⁺ were found in the clay samples, they are extrapolable to soils due to their similar behaviors. Therefore, the absorption of ²³⁸U and ²²⁶Ra is related to that of Al and is directly proportional since the sign found in the model is positive. A relationship between ²²⁶Ra and Fe and Ti has been found in previous studies on granite samples [61]. The interaction of Ti with ²²⁶Ra, according to our results, is positive, which is equivalent to the results obtained by Rubinos and Barral [58]. On the other hand, the negative relationship found between ²²⁶Ra and Fe is due to the presence of a high concentration of Fe, as is the case with the soils in the study area, which can influence the mobility of ²²⁶Ra as it forms complexes with Fe₂O₃ that prevent its absorption by plants [62]. The proposed models emphasize the significance of mycorrhization in the uptake of radionuclides and heavy metals by plants [63,64,65]. The microorganisms present in the roots play a crucial role in the mobility of ²³⁸U and ²²⁶Ra, as well as Fe and Al, due to pH variation and chelate formation [24]. The microorganisms may induce the chemical dissociation of the compounds formed by ²³⁸U with Al and Fe and by ²²⁶Ra with S and Ti. These mechanisms need to be investigated and confirmed in future studies on the transfer factors of radionuclides under arid and semi-arid conditions.

4. Conclusions

The results obtained have allowed us to verify our working hypothesis that the incorporation of ²³⁸U and ²²⁶Ra in a climate similar to a semi-arid one is related to that of stable elements.

The study area was characterized by the high presence of radionuclides from uranium and actinium series, along with Cu and Fe values well above the global values for soils. Despite the complexity of this type of soil and the similarity of the climate in the study area to a semi-arid climate, no different TFs were obtained compared to those of areas with different climates, such as tropical ones. The range of transfer factors for ²³⁸U varied between 1.5·10⁻⁴ kg⁻¹ kg⁻¹ and 7.8·10⁻³ kg⁻¹ kg⁻¹, while for ²²⁶Ra, the range was between 1.8·10⁻⁴ kg⁻¹ kg⁻¹ and 4.0·10⁻² kg⁻¹ kg⁻¹. Likewise, in addition to the values obtained for the TFs, the behavior of the natural radionuclides in relation to the stable elements, mainly K, Mg, Ca, Fe, and Ti, is similar to those observed in the literature. The results showed how K and Mg lead to the reduced incorporation of ²³⁸U and ²²⁶Ra due to the possible presence of mycorrhizal fungi, resulting from the greater incorporation of K and also competition between Mg and the remaining alkaline earth metals (Ra, Ca, and Sr). The presence of arbuscular mycorrhizal fungi has not been verified, and this conclusion is based on the results of previous studies; however, a count of arbuscular mycorrhizal fungi vesicles is necessary to ensure this conclusion. Furthermore, the mobility of ²²⁶Ra is related to the presence of Ti, an aspect observed by the ²²⁶Ra/Ti correlation in the biplot graph representing the TFs. Finally, the incorporation of U showed the expected behavior and correlated with the presence of Fe and Al typical of granitic soils like that of the study area.

The correlations found were consistent with the parameters with the most weight in the absorption models of ²³⁸U and ²²⁶Ra. The interaction between Al and ²³⁸U and ²²⁶Ra was similar to that observed in clays, showing a positive relationship between them. On the other hand, Ti facilitates the movement of ²²⁶Ra, which matched the positive relationship in the model. Finally, Fe retains Ra by forming a complex with Fe₂O₃ when there is a high concentration of Fe, an aspect that is consistent with the levels found in the study area. Our results indicate that both Scolymus hispanicus L. and Eryngium campestre L. could be utilized in the phytoremediation of soils contaminated by natural radionuclides in semi-arid climates. Therefore, our results have confirmed that there is a significant relationship between the incorporation of stable and radioactive elements, which helps in understanding the absorption mechanisms of plants under conditions similar to a semi-arid climate.