Pollution and health risk assessment of soil, water and vegetables from an ion-adsorption rare earth mining area in Ganzhou, China

doi:10.21203/rs.3.rs-4883640/v1

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Pollution and health risk assessment of soil, water and vegetables from an ion-adsorption rare earth mining area in Ganzhou, China

https://doi.org/10.21203/rs.3.rs-4883640/v1

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Rare earth elements (REEs) are increasingly recognized as significant environmental pollutants due to their environmental persistence, bioaccumulation, and chronic toxicity. This study assessed REEs pollution in soil, water, and vegetables in an ion-adsorption rare earth mining area in Ganzhou, and evaluated the associated health risks to the local population. Results indicated that the REEs content in soil ranged from 168.58 to 1915.68 mg/kg, with an average of 546.71 mg/kg, substantially surpassing the background level for Jiangxi Province (243.4 mg/kg) and the national average (197.3 mg/kg). Vegetables displayed an average REE content of 23.17 mg/kg in fresh weight, far exceeding the hygiene standard of 0.7 mg/kg. Water samples contained REEs at a concentration of 4.09 µg/L. The estimated daily intake (EDI) of REEs from vegetables exceeded the threshold for subclinical damage, posing potential health risks, particularly for children and adolescents. Further analysis of the adjusted average daily intake (ADI) suggested that while most vegetable consumption remains within safe threshold, the intake of REEs from high-risk vegetables such as pakchoi and radish should be limited.

Rare earth elements

Emerging pollutants

Health risk assessment

Rare earth elements (REEs) consist of 17 elements, including 15 lanthanides—lanthanum (La), cerium (Ce), praseodymium (Pr), neodymium (Nd), promethium (Pm), samarium (Sm), europium (Eu), gadolinium (Gd), terbium (Tb), dysprosium (Dy), holmium (Ho), erbium (Er), thulium (Tm), ytterbium (Yb), and lutetium (Lu)—plus scandium (Sc) and yttrium (Y). Based on their atomic numbers and masses, REEs are typically classified two groups known as light rare earth elements (LREEs: La, Ce, Pr, Nd, Pm, Sm, and Eu) and heavy rare earth elements (HREEs: Gd, Tb, Dy, Ho, Er, Tm, Yb, Lu, and Y) (Ramos et al., 2016; Arienzo et al., 2022). REEs are crucial in diverse fields such as military technology, artificial intelligence, metallurgy, and medicine due to their unique physicochemical properties (Chakhmouradian & Wall, 2012). China holds approximately 23% of the global REEs reserves, with significant contributions from Ganzhou, Jiangxi Province, where one-thirds of the country's ion-absorbed REE resources are found (Li et al., 2010). However, traditional mining methods such as pool, heap, and in-situ leaching have caused severe environmental issues, including soil acidification, structural damage, and potential risks to water and vegetation (Li & Wu, 2017; Li et al., 2020a; Nie et al., 2020).

Pollutants from REEs mining activities readily contaminate local soil and water systems, creating ecological hazards. An ecological study in Dingnan County, Ganzhou, has demonstrated that tailing soils contained REE concentrations ranging between 187 and 1148 mg/kg, with an average of 392 mg/kg. Surface water near these mining sites revealed average REE concentrations of 7730 µg/L, while both sediments and farmland soils far exceeded natural background levels, indicating notable environmental contamination (Liu et al., 2019). Another study in the same county identified tailings with low organic content and high REE concentrations, reaching up to 808.5 mg/kg (Zhao et al., 2019). Along the Dongjiang River near Xunwu County, REE concentrations ranged from 0.18 to 87.7 mg/L, surpassing background levels by up to 14,126 times (Wei et al., 2001). In Longnan County, Ganzhou, the average soil REE concentration was 976.94 mg/kg, significantly higher than both Jiangxi and national background levels (Jin et al., 2014). Moreover, this contamination extends beyond environmental media to agricultural products. For instance, REEs measured in ten different crops from mining sites in Longnan County ranged from 1.04 to 78.57 mg/kg, exceeding China's vegetable hygiene standard limit of 0.7 mg/kg (Jin et al., 2014). A study examining various food types from six Chinese provinces found that some vegetables exceeded the safety limits for total REEs (Jiang et al., 2012). Such widespread contamination underscores the urgent need for comprehensive studies on REEs transfer and ecological risks, especially regarding their impact on human health (Li et al., 2013; Gwenzi et al., 2018; Arbalestrie et al., 2022).

REEs can enter the human body through various exposure routes and accumulate in different tissues or organs, posing a threat to human health. Prolonged exposure to REE dust in mining and processing industries has been linked to pneumoconiosis development (Hirano & Suzuki, 1996; Censi et al., 2011). As non-essential elements, excessive REE intake can lead to physiological and immune dysfunction (Yu et al., 2007). Additionally, elevated REE levels have been also associated with severe health issues like DNA damage and carcinogenesis (Bai et al., 2019; Gaman et al., 2021). Notably concerning is the potential neurotoxic effect on children, where exposure could impair cognitive functions, such as intelligence quotient (IQ) and memory (Li et al., 2020b; Wei et al., 2020). Animal studies further confirm that REE exposure can negatively impact respiratory, cardiovascular, neurological, and reproductive systems (Nakamura et al., 1997; Toya et al., 2010; Lin et al., 2017; Gao et al., 2022; Lee & Park, 2022; Xiong et al., 2023).

The present research focuses on the environmental and health impacts of REEs in Northern Ganxian District, Ganzhou, Jiangxi Province, and an area known for its ion-adsorption rare earth deposits. The study aims to comprehensively assess REEs levels in soil, water and vegetables to understand how REEs transfer from the environment to agricultural products and ultimately to human consumers. It also examines variations in REE absorption by different vegetable types, essential for identifying crops that pose higher consumption risks. Moreover, the study evaluates the potential health risks from ingesting REEs through these environmental mediums. The findings are expected to inform the development of targeted environmental protection policies and health measures to mitigate the impact of REE contamination on local population health.

Study area and sample collection

This study was conducted in an ion-adsorption rare earth mining area within Jiangkou Town, northern Ganxian District of Ganzhou City (115°7'33"~115°8'4"E longitude and 25°52'53"~25°53'22"N latitude). In the study area, the soil, primarily composed of red soil, is characterized by a deep, loose structure with poor corrosion resistance and limited water and fertilizer retention capabilities. These conditions are conducive to the presence of rare earth.

A total of 21 soil samples (0–20 cm depth), 19 vegetable samples and 15 water samples were collected during October 2023 and April 2024 (Fig. 1). Soil samples were air-dried, cleared of gravel and organic debris, and then passed through a 10-mesh nylon sieve. The samples were then ground in an agate mortar, sieved through a 100-mesh nylon sieve, and analyzed. The soil pH, determined using a 1:2.5 soil-to-water ratio, ranged from 3.94 to 6.14, averaging 4.52, indicating mild acidity. Vegetable samples were washed three times with deionized water, oven-dried at 75°C until constant weight, ground, sieved through a 100-mesh nylon sieve, and stored in polyester bottles at -20°C for later analysis. Water samples were filtered through a 0.45 µm nylon membrane and stored at 4°C until further analysis.

Sample analysis

Samples were digested using a microwave digestion system (MARS 6, CEM, USA) with 0.500 g of each sample treated with 8 mL of a 3:1 nitric and hydrochloric acid mixture. Vegetable samples received an additional 2 mL of hydrogen peroxide. Following the acid-driven treatment, the solution was diluted to 10 mL in a volumetric flask using a 2% nitric acid solution, and then filtered through a 0.45 µm nylon membrane before analysis.

The content of 16 REEs in the samples was determined by inductively coupled plasma mass spectrometry (ICP-MS, NexION 350X, PerkinElmer, USA). Concentrations were expressed in milligrams per kilogram dry weight for soil and wet weight for vegetables. Calibration used standard solutions ranging from 0.1 to 20 µg/L (GNM-M160181-2013, National Nonferrous Metals and Electronic Materials Analysis and Testing Center, China). In, Re, and Rh (GNM-M030025-2013) were used as internal standards at a concentration of 25 µg/L to ensure analytical quality. The detection limits ranged from 0.0063 to 0.0611 µg/L, calculated as three times the standard deviation from eleven blank solutions (Table S1). Each sample underwent duplicate analysis. Accuracy and precision were assessed using national standard samples (soil GBW07565 and spinach GBW10015a, national certified reference materials of China), with recovery rates from 89.45–116.77% (Table S1).

Geostatistical and REEs characteristic parameters analysis

Geochemical parameters provide insights into the enrichment, fractionation, and sources of REEs in soils. The North America Shale Composite (NASC) normalization method is commonly applied to minimize the zigzag patterns in REE distributions, a phenomenon attributed to the Oddo-Harkins rule. This study specifically examines Eu and Ce anomalies (denoted as δEu and δCe) to assess REE geochemical behavior in soil, calculated using the equations (1) and (2):

δEu = Eu_N/(Sm_N×Gd_N)^1/2 (1)

δCe = Ce_N/(La_N×Pr_N)^1/2 (2)

where the subscript N indicates the NASC-normalized values of the measured concentrations of REEs. Anomalies are classified as positive when values exceed 1.05 and negative when below 0.95 (Algeo & Maynard, 2004).

Additionally, to assess REE differentiation, ratios such as LREEs/HREEs, the NASC-normalized La/Sm ((La/Sm)_N) and Gd/Yb ((Gd/Yb)_N) were calculated. These ratios are defined as:

(La/Sm)_N$\:\:=\frac{{\text{L}\text{a}}_{\text{s}\text{a}\text{m}\text{p}\text{l}\text{e}}/{\text{L}\text{a}}_{\text{N}}}{{\text{S}\text{m}}_{\text{s}\text{a}\text{m}\text{p}\text{l}\text{e}}/{\text{S}\text{m}}_{\text{N}}}$

(Gd/Yb)_N$\:\:=\frac{{\text{G}\text{d}}_{\text{s}\text{a}\text{m}\text{p}\text{l}\text{e}}/{\text{G}\text{d}}_{\text{N}}}{{\text{Y}\text{b}}_{\text{s}\text{a}\text{m}\text{p}\text{l}\text{e}}/{\text{Y}\text{b}}_{\text{N}}}$

where "sample" represents the concentration of La, Sm, Gd and Yb in the studied soils, and the subscript N represents the NASC-normalized values.

Pollution assessment of REEs in soil

The pollution status of the studied soil was assessed by the geo-accumulation index (I_geo), using the following Eq. (3):

$\:{I}_{\text{g}\text{e}\text{o}}$=$\:{\text{log}}_{2}{[\text{C}}_{\text{i}}/(1.5{\text{B}}_{\text{i}}$)] (3)

where C_i represents the concentration of a specific REEi in the soil; B_i denotes the background value for that REEi in the soil of Jiangxi Province, China. The classification criteria of I_geo is shown in Table S2.

Potential ecological risks of REEs in soil

Ecological risks from REEs were assessed using method developed by Hakanson (Hakanson, 1980), which considers both individual and cumulative element risks. The equations involved are (4), (5) and (6):

$$\:{C}_{f}^{i}={C}_{i}/{C}_{b}^{i}$$

$$\:{E}_{r}^{i}={C}_{f}^{i}\bullet\:{T}_{r}^{i}$$

RI = Σ$\:{\:E}_{r}^{i}$ (6)

where $\:{C}_{f}^{i}$ and $\:{E}_{r}^{i}$ denote the contamination factor and ecological risk factor of element i in soil; C_i is the concentration of element i in soil; $\:{C}_{b}^{i}$ and $\:{T}_{r}^{i}$ refers to the pre-industrial background value and toxic response factor of element i in soil, respectively. Risk classification is detailed in Table S2.

Health Risk Assessment for REEs

The health risks assessment of REEs in vegetables for both children and adults near mining areas were evaluated using the estimated daily intake (EDI) model proposed by the Environmental Protection Agency of the United States (US EPA, 2000). This model calculates exposure doses based on the following Eq. (7):

EDI= $\:\frac{{\text{C}}_{\text{R}\text{E}\text{E}}\text{*}\text{C}\text{R}}{\text{B}\text{W}}$ (7)

EDI (µg/kg/day) indicates the estimated daily intake of REE, where C_REE represents the sum of REEs in vegetable based on wet weight (mg/kg ww); CR is the vegetable consumption rate by different age groups (%) (from the Fifth China Total Diet Study, FCTDS) (Wu et al., 2018), BW is the average body weight of different age groups (%) (from FCTDS). The vegetable consumption rate for children is typically estimated to be 57% of that of adults (US FDA, 2003).

To assess the long-term health risks of REEs from consuming vegetables and water, the average daily intake (ADI) for children and adults was calculated as following Eq. (8):

ADI = $\:\frac{\:\:\text{C}\times\:\text{G}\text{W}\times\:\text{E}\text{F}\times\:\text{E}\text{D}}{\text{B}\text{W}\times\:\text{A}\text{T}}$ (8)

where ADI is the average daily intake of REEs (mg/kg/day); C is the content of REEs in the sample (mg/kg); GW is the daily intake (kg/day), the average daily intake of total vegetables is in 0.22 kg/day for children and 0.85 kg/day for adults (from FCTDS), respectively, and the average daily intake of water is 1.1 L/day for children and 1.6 L/day adults (from the Chinese Dietary Guidelines)(Chinese Nutrition Society. (2022)); EF is exposure frequency, which is 365 days/year for children and adults; ED refers to exposure period, which is 6 years for children and 25 years for adults; BW is the bodyweight, which is 15 kg and 60 kg for children and adults, respectively; AT refers to the average time, which is 2190 days for children and 9125 days for adults.

Statistical analysis

Statistical analyses were conducted to compare average results from various soil and vegetable samples by using a one-way analysis of variance (ANOVA) followed by Scheffe's test for multiple comparisons. Spearman correlation and principal component analysis were performed to explore the relationships between soil and vegetable REEs. A significance level of (P < 0.05) was applied for multiple comparisons and analyses. Data analysis was carried out using SPSS software package version 14.0 (SPSS Inc., USA) and Excel 2010.

REEs content in study soils

The total concentration of REEs (ΣREEs) in study soils ranged from 168.58 to 1915.68 mg/kg, with an average of 546.71 mg/kg (Table S3), exceeding the natural background levels in Chinese soil (197.3 mg/kg) and Jiangxi soil (243.4 mg/kg) (Wang et al., 2022). The hierarchy of average concentrations for individual REEs is as follows: Ce > Y > Nd > La > Sc > Gd > Pr > Sm > Dy > Er > Yb > Tb > Ho > Eu > Lu > Tm. Notably, Ce, Y, Nd, and La collectively contributed over 80% of the total REE concentration, with Ce being the most prevalent, ranging from 55.18 mg/kg to 458.31 mg/kg. The concentrations of HREEs varied from 44.80 to 771.74 mg/kg with an average of 175.64 mg/kg, whereas LREEs ranged from 119.50 to 1068.59 mg/kg with an average of 348.61 mg/kg. Moreover, the Y was the most abundant HREE, comprising 23.63% of total REEs, while the HREEs accounted for 32.13%, indicating typical of a Y-enriched ion-adsorption rare earth deposit.

Distribution characteristics of REEs in study soils

The distribution of REEs in the study soils was characterized by the ratios of LREE/HREE, NASC-normalized La/Sm and Gd/Yb, along with the anomalies of Ce and Eu (Table 1). The ratios of LREE/HREE, ranging from 1.02 to 7.10 with an average of 2.94, suggest a higher content of LREEs relative to HREEs. However, these ratios are significantly lower than those in northern China, indicating a comparative enrichment of HREEs in the study area (Liang et al., 2014). The slope of LREE and HREE concentration curves can be reflected by normalized (La/Sm)_N and (Gd/Yb)_N ratios, respectively (Schilling & Winchester, 1966). In this study, the (Gd/Yb)_N ratio of 3.15 points to a high degree of internal fractionation of HREEs compared to LREEs, which had a (La/Sm)_N ratio of 1.16. This aligns with the broader distribution pattern of "light north, heavy south" across China.

Table 1

Geochemical characteristic parameters of soil rare earth elements in the study area
Statistics	ΣLREEs/ΣHREEs	(La/Sm)_N	(Gd/Yb)_N	δCe	δEu
Minimum	1.02	0.74	1.43	0.56	0.13
Maximum	7.10	1.69	5.55	1.14	0.35
Mean	1.98	1.16	3.15	0.84	0.25
SD	1.88	0.29	1.14	0.16	0.06
Median	2.62	1.17	3.22	0.81	0.26
Note: SD: Standard deviation.

Ce and Eu can exist in multiple valence states (Ce⁴⁺ and Eu²⁺), differentiating their fractionation behavior from other REEs. The anomalies of Ce and Eu are indicative of specific geochemical processes and sources of REEs (Yu et al., 2022). In our study, all soil samples consistently showed negative Eu and Ce anomalies, with mean values of 0.25 and 0.84, respectively. The δCe/δEu ratio in studied soil averaged 3.36, suggesting a predominance of reducing conditions, as this ratio exceeds 1. Under such conditions, Eu³⁺ can be reduced to Eu²⁺ and potentially lost, especially under strong reducing conditions. Meanwhile, positive Ce anomalies often correlate with high sediment alkalinity (Kim et al., 2012). As soil acidity increases, the transformation of Ce³⁺ to Ce⁴⁺ occurs, often alongside the precipitation of iron or manganese oxides, contributing to the observed negative Ce anomalies. Consequently, the studied soil exhibited marked negative δEu and slightly negative δCe anomalies, attributable to mild acidic and strongly reducing conditions.

Pollution and potential ecological risks in study soils

The pollution status of REEs was assessed using I_geo, which provided insights into the ecological risks of soil contamination. The hierarchy of average I_geo values for REEs ranged from Gd (1.51) to Tm (-0.29), with Gd reaching moderate pollution levels (Fig. 1a). The results indicated that Nd (0.83) and Y (0.79), both HREEs, displayed similar distribution patterns, ranging from no pollution to high pollution across all sample sites. Seven other elements had average I_geo values above zero (Tb (0.68), Pr (0.56), La (0.21), Ce (0.13), Dy (0.11), Sc (0.05)), indicating varying levels of pollution from none to moderate. Approximately one-third of the sampling sites were considered unpolluted, with only two out of 21 sites (S20 and S7) showing moderate pollution with I_geo values exceeding 1. The remaining sites indicated between unpollution to moderate pollution categories. The LREEs primarily revealed the lower contamination categories, in contrast to the broader dispersion observed for HREEs, which contributed more significantly to soil pollution.

The ecological risk of REEs is shown in Fig. 2b. Based on the results, all REEs had average $\:{E}_{r}^{i}$ values lower than 40.0, indicating a low ecological risk. The cumulative average RI values for 21 soil sites ranged from 82.32 to 573.29, ranging between 150 and 300, indicating that the overall mining area is at a moderate risk. According to criteria, two of 21 sites (S20 and S7) had RI values (573.29 and 418.41, respectively) above 300, indicating a considerable ecological risk of REEs in these sites. At the same time, 57% of the sites were medium risk, while the rest were low risk. Eu, Ho, Lu, Dy and Gd were identified as primary contributors to ecological risk, with numerous sites exhibiting medium risks due to high concentrations and toxicity coefficients of these elements (Nkrumah et al., 2021), posing potential threats to the ecosystem. To sum up, the I_geo and RI assessments suggest that the soil in the study area generally faces low to moderate ecological risks, with sites S7 and S20 requiring further attention and potential remediation.

Transport and impact of REE in soil-plant system

REEs may disrupt plant ion transport, influencing its growth by modifying nutrient uptake and physiological functions. Study indicates that REEs absorbed from the soil tend to accumulate in the edible parts of crops, potentially increasing human exposure through diet (Li et al., 2013). Thus, the present study detected REE concentrations in 19 vegetable samples, finding a range from 0.52 to 78.93 mg/kg and an average of 23.17 mg/kg (Table S4). These levels substantially surpass the Chinese National Standard of 0.7 mg/kg fresh weight (GB2762-2005) and those found in commercially available vegetables in China, which range from 0.28 to 1.40 mg/kg (Jiang et al., 2012). In the examined mining area, REE concentrations in vegetables were also markedly higher than those reported from other mining areas in Fujian (3.58 mg/kg) and Jiangxi provinces (6.55 mg/kg) (Zhang et al., 2000b; Li et al., 2013), akin to findings from a HREEs mine in Longnan county ranging from 1.55 to 78.57 mg/kg (Jin et al., 2014). The uptake and translocation of REEs across plant species are diverse, typically, LREEs are more easily enriched in plants compared to HREEs (McLennan and Taylor, 2012). Consistently, we found that over 50% of the total REEs in vegetables were attributed to La and Ce in this study (Table S4).

Besides these differences in enrichment, this study observed varying REE absorptive capacities across different vegetable species. We categorized the vegetable samples into three groups: root vegetables, leafy vegetables, and solanaceous vegetables. Solanaceous vegetables exhibited the lowest REE concentrations, ranging from 0.52 to 3.09 mg/kg with an average of 1.81 mg/kg. In contrast, root vegetables had higher concentrations, ranging from 6.11 to 78.93 mg/kg with an average of 25.69 mg/kg. Leafy vegetables presented concentrations between 0.61 and 47.40 mg/kg, averaging 25.67 mg/kg (Table S5). Consistent with the results reported by Zhang et al, our findings show comparable REE content in both leafy and root vegetables, suggesting that roots and lush foliage facilitate greater REE uptake from soils (Zhang et al., 2000a).

REE accumulation in vegetables near mining sites correlates with soil bioavailability. We conducted a systematic analysis using Spearman correlation matrices to investigate the migration characteristics of REEs in different vegetables. Our findings indicated significant positive correlations between the soil concentrations of Pr, Nd, Sm, Dy, Ho, Er and Tb and their presence (P < 0.05) in leafy vegetables (Fig. 3a). Root vegetables showed significant correlations with all soil REEs except Ce and Tm (P < 0.05, Fig. 3b) and solanaceae vegetables correlated significantly with all soil REEs except Tm, Lu, and Sc (P < 0.05, Fig. 3c). These findings suggest selective absorption and accumulation of REEs by different vegetable species.

Principal component analysis (PCA) of soil and vegetable samples revealed two distinct clusters along the first two principal components (PC1 and PC2), explaining 86.6% and 8.9% of the variance, respectively (Fig. 3). This analysis confirmed consistent REE distribution patterns across different sites, emphasizing their ecological interconnectedness. Notably, the PCA indicated distinct groupings of LREE and HREE in PC2, with Sc and Y being the most significant variable. Excluding Sc and Y, the remaining REEs significantly influenced PC1, demonstrating unique REE patterns for each vegetable variety despite identical soil and climate conditions.

Human health Risk Assessment

REEs are increasingly recognized as emerging pollutants due to their environmental persistence, bioaccumulation, and chronic toxicity (Queiros et al., 2023; Wang et al., 2023). Concerningly, REEs may adversely affect childhood neurodevelopment, potentially leading to reduced IQ and memory deficits (Wei et al., 2020). Previous study suggested that a daily REE intake of up to 70 µg/kg is generally safe for human health, but levels between 100 and 110 µg/kg/day could cause subclinical damage (Zhu et al., 1997). Consequently, it is essential to closely monitor the REE intake of residents near mining areas, particularly children aged 2 to 12 years, who are particularly vulnerable to pollutants through ingestion pathway.

Our study estimated that daily intake of REEs through vegetables significantly exceeded the safe threshold of 70 µg/kg/day for all age and gender groups. Notably, children aged 2 to 12 years were at the highest risk, with intake levels reaching 2.29 times (ages 2 to 7) and 1.73 times (ages 8 to 12) the subclinical threshold (Table 2). In contrast, individuals over 13 years exhibited intake levels near the risk threshold, with minimal variation across genders. Regarding water sources, REE concentrations were lower than in other mining regions such as Dingnan and Longnan counties, ranging from 2.00 to 6.87 µg/L with an average of 4.09 µg/L (Jin et al., 2014; Liu et al., 2019). Although the average daily intake (ADI) of REEs from water was significantly below the recommended limits, suggesting minimal risk (Table 3), the ADI from vegetables in children and adults consistently reached up to three times the subclinical threshold, indicating significant health risks. When adjusting for consumption frequency of different vegetable species, the daily cumulative REEs intake from selected vegetables was 49.46 µg/kg/day (Table 4), nearing the tolerable ADI of 51.5 µg/kg·BW set by the China Scientific Committee (Yang et al., 2022). Pakchoi and radish, which contribute 76.99% to the dietary REE risk, are of primary concern for residents in mining areas. It is important to note that the daily dietary intake also includes grains, meats, fruits, and other foods. For instance, Jin et al. reported that the daily cumulative intake of REEs through various crops and well water in mining areas reached up to 295.33 µg/kg (Jin et al., 2014). Given the high REE levels observed in studied soil and vegetables, further comprehensive risk assessment of REEs in daily diet is warranted.

Table 2

EDI of total REEs via vegetable consumption in mining areas by different gender and age groups
Gender/age group	BW^a (kg)	CR^a (g)	EDI
2 ~ 7 years	17.9	194.8	252.2
8 ~ 12 years	33.1	272.4	190.7
Male, 13 ~ 19 years	56.4	396.7	163.0
Female, 13 ~ 19 years	50.0	317.9	147.3
Male, 20 ~ 50 years	63.0	436.4	160.5
Female, 20 ~ 50 years	56.0	412.1	170.5
Male, 51 ~ 65 years	65.0	477.9	170.4
Female, 51 ~ 65 years	58.0	447.0	178.6
Male, > 65 years	59.5	413.3	160.9
Female, > 65 years	52.0	364.1	162.2
Note: EDI: estimated daily intake, µg/kg/day. ^a Data were from the fifth China total diet study.

Table 3

Average daily intake dose of REEs of local children and adults in mining area via vegetables and water
Group	Sample	Average content (mg/kg)	Days of consumption (days/year)	Daily consumption (kg)	Average bodyweight (kg)	Average daily intake (µg/kg/day)
Children Adult	Vegetables	23.17	365	0.22 0.85	15 60	339.83 338.24
Children Adult	Water	4.09^a	365	1.10^b 1.60^b	15 60	0.30 0.10
Note: ^a The unit of REEs in water is µg/L; ^b The unit of REEs in water daily consumption is L.

Table 4

Adjusted ADI of REEs based on consumption frequency of different vegetable species
Vegetable species	Average content (mg/kg)	Days of consumption^a (days/year)	Daily consumption (kg)	Average daily intake (µg/kg/day)	Proportion (%)
Pakchoi	33.78	180	0.10	27.76	56.13
Radish	11.31	100	0.20	10.33	20.88
Spinach	17.56	60	0.05	2.41	4.86
Bamboo shoot	8.16	30	0.10	1.12	2.26
Celery	50.47	60	0.05	6.91	13.98
Water fennel	9.34	30	0.05	0.64	1.29
Red pepper	1.81	180	0.02	0.30	0.60
Total				49.46	100.00
Note: ^a Data were from a study conducted by Jin et al (Jin et al., 2014).

The concentration of REEs in soil from the studied area was significantly higher than the background levels in China and Jiangxi province. The distribution patterns in the mining area showed a trend of HREE enrichment. Vegetables from these sites, particularly leafy and root vegetables, exhibited higher REE concentrations than those permitted by China's National Standards. Health risk assessments reveal that both the EDI and ADI for REEs via vegetables exceed the critical values for subclinical damage, posing a potential health threat to local residents. Given the strong enrichment capacity of leafy and root vegetables, particularly pakchoi and radish, it is advisable for locals to limit their consumption to mitigate exposure risks. Additionally, the potential impacts of prolonged low-level REE via dietary intake exposure on residents, especially children and adolescents, require further investigation.

Author Contribution

Conceptualization: Liang Xiong; Formal analysis: Liye Zhu and Jinyu Huang; Investigation: Liye Zhu, Jinyu Huang, Hui Huang, Sihui Wang, Ziyue Sun, Yi Fang and Ming Hao; Methodology: Jinyu Huang, Qi Wang, and Liang Xiong; Data curation: Qi Wang, Chunmei Wu and Ming Hao; Writing - original draft: Liye Zhu; Writing - Review & Editing: Liang Xiong; Funding Acquisition: GongHua Hu and Liang Xiong. All authors reviewed the manuscript.

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Pollution and health risk assessment of soil, water and vegetables from an ion-adsorption rare earth mining area in Ganzhou, China

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