Spatial distribution, formation mechanism, and health risk assessment of high-fluoride groundwater at the largest antimony mine in Hunan Province, China Open Access

Fluorine (F) is an important trace element and the lightest member of the halogen group. Although fluoride plays an important role in the growth and strength of bones and teeth, it is linked with several serious health problems in high concentrations (Duggal & Sharma 2022). For example, dental and skeletal fluorosis can occur when F⁻ concentrations reach 1.50 and 4.00 mg/L, respectively (Lafayette et al. 2020). The World Health Organization (WHO) accordingly has issued that the concentration of fluoride in drinking water should not exceed 1.50 mg/L. In China, the maximum permitted F⁻ concentration in drinking water is 1.00 mg/L (Liu et al. 2021). Worldwide, however, high-fluoride groundwater continues to pose a threat to the health of millions. Consequently, assessing the spatial distribution of high-F⁻ groundwater is significant for the protection of regional groundwater resources and drinking water safety. Globally, groundwater fluorosis affects more than 260 million individuals (Ali et al. 2019). The high level of fluoride pollution in groundwater is observed in various countries mainly Pakistan, India, Sri Lanka, Iran, Bangladesh, and China (Noor et al. 2022). Especially in rural areas of developing countries, residents are more exposed to fluoride contamination-related diseases because of drinking fluoride excess concentration water (Dehghani et al. 2019). The F⁻ content in groundwater ranged from 10.00 to 20.00 mg/L in many regions such as Balod and Brahmaputra floodplain regions in India (Yadav et al. 2019).

Factors such as temperature, rainfall, evaporation, and human activities contributed to environmental pollution (Kousali et al. 2022). Generally, the F⁻ in groundwater originates from two sources, namely, anthropogenic input and geological genesis (Onipe et al. 2021). Geological genesis includes the dissolution and precipitation of F-bearing minerals, the evaporation effects, competitive adsorption, and cation exchange (Duggal & Sharma 2022). In addition, an alkaline environment with low Ca²⁺ content and high contents of Na⁺ and is helpful for increasing the F⁻ content in groundwater (Dong et al. 2022). Anthropogenic sources are derived mainly from the use of fertilizers and pesticides and excess fluoride emissions from coal burning and mining activities (Huang et al. 2022). In most arid mining situations, fluoride may come from the leaching of coal, coal combustion, and mining activities. However, in humid and semi-humid mining areas, the role of mining activities in fluoride enrichment in groundwater is not fully understood.

Drinking water with F⁻ concentrations in excess of 1.50 mg/L can have detrimental effects on the intelligence quotient (IQ) of children, with one of the early indicators of impaired IQ in children being dental fluorosis (Hu et al. 2021). In the present study, the health risk assessment model was recommended by the United States Environmental Protection Agency (U.S. EPA) to evaluate the health risk of fluorine in high-fluoride areas. The scope of health risk assessment mainly includes estimating the number of pollutants entering the human body and evaluating the relationship between dose and adverse health effects (U.S. EPA 2004).

The Xikuangshan antimony mine has been nicknamed ‘World's Antimony Capital’, owing to annual 40,000 t Sb production capacity. Previous reports had found that the content of Sb in shallow groundwater ranged from 0.01 to 23.00 mg/L, with a mean value of 3.87 mg/L exceeding the Chinese Environmental Quality standard for groundwater (Sb ≤0.005 mg/L) 774 times (Hao et al. 2020). However, fluorides in groundwater have received little attention. Although the distribution of high- F⁻ groundwater and the factors contributing have been extensively focused, the formation mechanism of high-F⁻ groundwater in humid and semi-humid areas affected by Sb-bearing mineral mining has received little attention.

Therefore, the objectives of this study are (1) to determine the spatial distribution and geochemical behavior of fluorides in shallow groundwater of the Xikuangshan antimony mine area, (2) to elucidate the formation mechanism of high-fluoride shallow groundwater affected by Sb-bearing mineral mining, and (3) to conduct a health risk of high-F⁻ shallow groundwater in the Xikuangshan antimony mine area. The findings of this study will make a valuable contribution toward devising strategies for ensuring the safety of drinking water and sound management of the geological environment.

The Xikuangshan antimony mine area (geographical coordinates: 27°46′–27°49′ N and 111°28′–111°32′ E) is located in a mountainous region of Hunan Province, with an area of 26 km². The mining area has a typical continental monsoon humid subtropical climate, with an average annual precipitation of approximately 1,381.60 mm, an average annual wind speed of 1.90 m/s, and an average temperature of 16.70 °C, respectively. This mining area is considered to have the largest deposits of antimony in the world, with estimated reserves of 2.11 × 10⁶t (Hao et al. 2020). However, geological environmental problems, such as ground subsidence, water and soil pollution, and ground landscape damage, had occurred (Wen et al. 2018).

The main aquifers within the Xikuangshan mine area are the Magunao (D₃x⁴) and Shetianqiao (D₃s²) aquifers. The aquifuge between D₃x⁴ and D₃s² has weak water hydraulic connectivity. The D₃x⁴ aquifer is the main source of drinking water in the study area, and the main water chemical types are HCO₃–Ca and HCO₃–Ca–Mg. The study area is located in a closed water-bearing structure trending east-west, with impervious boundaries to the east (the lamprophyre vein) and the west (the F₇₅ fault). The primary source of D₃x⁴ groundwater recharge is atmospheric precipitation. D₃x⁴ groundwater mainly flows in silicified tuff fissures, tuff solution spaces, and caves and eventually discharges through the springs. Owing to years of intensive mining activities, the upper part of the D₃x⁴ aquifer has been covered with large amounts of solid waste (including waste rock and slag). Additionally, little agricultural activities occurred in the study area.

Generally, in situ parameters including pH and total dissolved solids (TDS) were analyzed in the field via a pH meter (HANNA, Italy) and an electrochemical analyzer (HANNA, Italy). Then properly sealed and stored water samples were transported to the laboratory for further chemical analysis. The concentrations of K⁺, Na⁺, Ca²⁺, Mg²⁺, Cl⁻, ⁠, ⁠, and F⁻ were determined by ion chromatography (Dionex Integrion IC; Thermo Fisher Scientific, USA), and the concentrations of were determined using acid-base titration. The precisions of F⁻ and pH analyses and TDS analyses were 0.01 and 1 mg/L, respectively. Additionally, the analysis of other ions was 0.1 mg/L.

The storage and transportation of groundwater samples were carried out in accordance with the requirements of Regulation for Water Environmental Monitoring (GB/T 14848-2017) Set by the Ministry of Water Resources, P.R.C. To ensure precision and accuracy, all chemical analyses were performed in triplicate, and the data presented are the arithmetic mean values of the three replicate analyses, which should have a maximum relative standard deviation of less than 10%. Besides, the values obtained for all ions were subject to ion balance error calculation, with the permissible error being no more than 5%.

Origin 2021 (Mcmahon et al. 2020) software was used for data description and statistical analysis. Trilinear and Gibbs diagrams were used to elucidate the hydrogeochemical facies and processes. The effective spatial analysis tool ArcGIS 10.8 (Hu et al. 2022) was used for health risk assessment and comprehensive analyses of shallow groundwater utilization and management. Using ArcGIS 10.8, the data distribution of the samples in space was characterized. In addition, to assess the degree of equilibrium between the shallow groundwater and minerals, the geochemical model PHREEQC (Nawale et al. 2021) to calculate saturation index (SI) values was used, and the EPA health risk evaluation model was selected to evaluate health risks for different age groups. The geochemical data for shallow groundwater are shown in Table 1.

Table 1

Geochemistry data for high-F⁻ and low-F⁻ shallow groundwaters

Types .		Project .	K+ .	Na+ .	Ca2+ .	Mg2+ .	Cl− .	.	.	.	F− .	TDS .	pHa .
Low-F− shallow groundwater	F− < 1.00 mg/L	Max	7.7	47.8	258.0	27.9	11.6	587.0	260.0	20.8	0.88	1,054	7.74
		Min	0.2	0.0	14.0	1.6	0.3	47.1	9.2	0.0	0.08	135	4.97
		Mean	2.2	7.6	74.3	5.7	2.7	169.7	156.8	5.7	0.29	363	6.98
		SD	2.4	13.1	48.6	5.4	3.2	132.5	74.1	5.9	0.26	205	0.56
High-F− shallow groundwater	1.00 mg/L ≤ F− ≤4.00 mg/L	Max	7.7	151.0	175.0	54.6	34.7	642.0	248.0	45.4	3.60	1,082	7.78
		Min	1.6	25.6	52.3	6.0	2.6	262.0	138.0	0.9	1.09	527	7.05
		Mean	4.7	71.4	112.3	25.0	9.3	460.2	196.2	16.7	2.33	821	7.39
		SD	2.0	37.5	37.4	17.0	10.6	136.2	41.0	16.6	0.91	194	0.26
	F− > 4.00 mg/L	Max	5.0	395.0	210.0	53.4	21.7	912.0	600.0	22.9	15.00	1,311	9.32
		Min	4.9	98.0	20.4	1.6	9.5	795.0	187.0	12.9	5.05	1,229	7.14
		Mean	4.9	246.5	115.2	27.5	15.6	853.5	393.5	17.9	10.03	1,270	8.23
		SD	0.06	210.0	134.1	36.6	8.6	82.7	292.0	7.1	7.04	57	1.54

Types .		Project .	K+ .	Na+ .	Ca2+ .	Mg2+ .	Cl− .	.	.	.	F− .	TDS .	pHa .
Low-F− shallow groundwater	F− < 1.00 mg/L	Max	7.7	47.8	258.0	27.9	11.6	587.0	260.0	20.8	0.88	1,054	7.74
		Min	0.2	0.0	14.0	1.6	0.3	47.1	9.2	0.0	0.08	135	4.97
		Mean	2.2	7.6	74.3	5.7	2.7	169.7	156.8	5.7	0.29	363	6.98
		SD	2.4	13.1	48.6	5.4	3.2	132.5	74.1	5.9	0.26	205	0.56
High-F− shallow groundwater	1.00 mg/L ≤ F− ≤4.00 mg/L	Max	7.7	151.0	175.0	54.6	34.7	642.0	248.0	45.4	3.60	1,082	7.78
		Min	1.6	25.6	52.3	6.0	2.6	262.0	138.0	0.9	1.09	527	7.05
		Mean	4.7	71.4	112.3	25.0	9.3	460.2	196.2	16.7	2.33	821	7.39
		SD	2.0	37.5	37.4	17.0	10.6	136.2	41.0	16.6	0.91	194	0.26
	F− > 4.00 mg/L	Max	5.0	395.0	210.0	53.4	21.7	912.0	600.0	22.9	15.00	1,311	9.32
		Min	4.9	98.0	20.4	1.6	9.5	795.0	187.0	12.9	5.05	1,229	7.14
		Mean	4.9	246.5	115.2	27.5	15.6	853.5	393.5	17.9	10.03	1,270	8.23
		SD	0.06	210.0	134.1	36.6	8.6	82.7	292.0	7.1	7.04	57	1.54

 Values less than the limit of detection (LOD) were set to zero for statistical purposes.

^aDimensionless

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The F⁻ concentrations in collected shallow groundwater samples were found to be in the range of 0.08–15.00 mg/L (mean:1.21 mg/L), with 25.64% of the samples having F⁻ concentrations higher than the Chinese national standard for drinking water (GB 5749-2022: 1.00 mg/L). Considering the drinking water guidelines in China and the degree of harm caused by high- F⁻ concentration in shallow groundwater to the human body (Toolabi et al. 2021), shallow groundwater samples were divided into the following three groups: below 1.00 mg/L (low-F⁻ shallow groundwater), 1.00–4.00 mg/L, and above 4.00 mg/L. All shallow groundwater samples in which the recorded F⁻ concentrations exceeded 1.00 mg/L were designated high-F⁻ shallow groundwater.

The measurement of shallow groundwater pH revealed values ranging from 4.97 to 9.32 (mean: 7.13). Among the collection sites, the shallow groundwater at 33.33% was acidic and that at 66.67% was alkaline. The concentrations of TDS were found in the range of 135–1,331 mg/L (mean: 504 mg/L), with 10.25% of samples having TDS concentrations higher than reported in the Chinese national standard for drinking water (GB 5749-2022: 1,000 mg/L). The mean concentrations of anions in shallow groundwater could be ranked as (mean: 264.4 mg/L) > (177.0 mg/L) > (8.6 mg/L) > Cl⁻ (4.7 mg/L), whereas the ranking for cations was Ca²⁺ (mean: 84.2 mg/L) > Na⁺ (32.9 mg/L) > Mg²⁺ (10.8 mg/L) > K⁺ (2.8 mg/L).

As shown in Figure 5, all shallow groundwater samples are distributed around the fluorite equilibrium line (PK_CaF2 = 10.6), indicating that F⁻ in the shallow groundwater is derived primarily from the dissolution of fluorite minerals (CaF₂ → Ca²⁺ + 2F⁻). While all samples of high-F⁻ shallow groundwater cluster around the fluorite dissolution line (Figure 5), all samples of the low-F⁻ shallow groundwater are distributed around the fluorite:calcite (mass ratio) = 1:200 line. These results accordingly indicate that F⁻ concentrations are mainly determined by mineral precipitation or cation exchange. With an increase in F⁻ concentration, the sample points move leftward along the ion exchange or calcite precipitation line, indicating that cation exchange or calcite precipitation also makes an important contribution to the increase in F⁻ concentrations in shallow groundwater.

Adsorbed F⁻ is an important source of F⁻ in shallow groundwater (Onipe et al. 2021), and generally, sediments in groundwater are capable of adsorbing certain anions, including fluorine (Huang et al. 2022). The presence of has been shown to reduce available adsorption sites of sorbents, thereby leading to a release of adsorbed F⁻ into shallow groundwater (Li et al. 2018). As shown in Supplementary Fig. S1, F⁻ content in high-F⁻ shallow groundwater was weakly positively correlated with the /(⁠ + Cl⁻) ratio, which could be attributable to the fact that can displace F⁻ adsorbed on mineral surfaces, resulting in its release into shallow groundwater.

As shown in Figure 4, there was a strong correlation between pH and F⁻ in high-F⁻shallow groundwater (R² = 0.85), which could be ascribed to the fact that in an alkaline environment, the surface of minerals is neutral or negatively charged, thereby inhibiting the adsorption of F⁻ and leading to its release (Hao et al. 2021).

If the values of CAI 1 and CAI 2 are both greater than 0, this indicates that K⁺ and Na⁺ in shallow groundwater have been exchanged for Ca²⁺ and Mg²⁺ in sediments. In contrast, if the CAI 1 and CAI 2 values are less than 0, Ca²⁺ and Mg²⁺ in shallow groundwater have been exchanged with K⁺ and Na⁺. Finally, CAI 1 and CAI 2 values of 0.00 indicate that there has been no ion exchange in shallow groundwater (Kousali et al. 2022). As shown in Supplementary Fig. S2, the CAI 1 and CAI 2 values of the shallow groundwater samples analyzed in the present study were all less than 0, thus indicating that Ca²⁺ and Mg²⁺ in shallow groundwater had been exchanged for K⁺ and Na⁺ in sediments. Specifically, the determined CAI 1 values ranged from −17.34 to −0.31 (mean: −6.54), whereas CAI 2 values ranged from −0.24 to −0.03 (mean: −0.05) (Supplementary Fig. S2). Notably, the absolute CAI 1 and CAI 2 values obtained for the high-F⁻ shallow groundwater samples were found to be considerably higher than those of the low-F⁻ samples, thereby indicating that ion exchange reactions are an important factor contributing to an increase in shallow groundwater F⁻ concentrations.

It has been established that the solubility of fluorite (K_fluorite = 10^−10.6) is less than that of gypsum (K_gypsum = 3.1 × 10⁻⁵) (Sharma & Kumar 2020), and the produced via the oxidation of stibnite (Sb₂S₃) and pyrite (FeS₂) will reduce the activity of Ca²⁺ and promote the precipitation of CaF₂. With an increase in the concentration of in shallow groundwater, the forward progress of reaction (Equation (8)) is inhibited, the precipitation of fluorite is reduced, and the increase in F⁻ concentration is further promoted.

F⁻ concentration, uncertainty, and other water chemistry variables were incorporated in the Positive Matrix Factorization (PMF) 5.0 model with factor number, random start seed number, and run number set to 3, 20, and 100, respectively.

As shown in Supplementary Fig. S4, factor 1 shows strong loadings for (85.70%), pH (80.40%), and Ca²⁺ (74.20%). The positive loading values for these three variables tend to indicate that their presence in shallow groundwater is a consequence of the dissolution and precipitation of minerals. Factor 2 shows strong loadings for Na⁺ (90.20%) and F⁻ (75.30%), and weak loadings for Ca²⁺ (2.26%) and Mg²⁺ (1.57%), thereby indicating that cation exchange occurs between Na⁺ and Ca²⁺ or Mg²⁺ in shallow groundwater. Hence, factor 2 is associated with a cation exchange source. Factor 3 shows a predominant loading of (86.50%), the primary sources of which in shallow groundwater are industrial pollution, agricultural activities, urban solid waste, and mining activities (Enalou et al. 2018; Huang et al. 2022). Given the predominance of mining activities at the study sites in the antimony mining area, hence factor 3 was associated with mining activities. Emphatically, the loading value for (8.50%) is all lower than 0.50 in factors 1, 2, and 3, implying that the oxidation of stibnite and pyrite was not the key factor for elevating F⁻ concentration in shallow groundwater.

Based on these factor fingerprints, the overall percentage contribution of each source was calculated, as shown in Supplementary Fig. S5. The main source of F⁻ in the shallow groundwater is cation exchange, accounting for 73.40%, followed by the dissolution and precipitation of F-bearing minerals (15.10%) and human activities (11.50%). Overall, geological processes, including the weathering-related cation exchange and dissolution and precipitation of F-bearing minerals, are the predominant factors influencing the F⁻concentrations in shallow groundwater in the Xikuangshan antimony mining area.

For infants, children, teenagers, adult males, and adult females, HQ values ranged from 0.05 to 10.42 (mean: 0.84), 0.09 to 18.75 (1.51), 0.04 to 7.87 (0.63), 0.05 to 10.00 (0.81), and 0.04 to 8.33 (0.67), respectively (Figure 8). Among these groups, children were found to have the highest likelihood of exceeding the maximum HQ recommended by the health risk index threshold, with a proportion reaching 36.38%, followed by adult males (23.12%), teenagers (22.21%), infants (21.22%), and adult females (21.11%). For children, the non-carcinogenic health risks associated with the intake of fluoride in shallow groundwater are at an unacceptable level (HQ > 1) and highlight the fact that children in the antimony mine area have a greater exposure to fluoride-related health hazards than adults (Chicas et al. 2022; Duggal & Sharma 2022). Consequently, there is a clear need for appropriate monitoring and management of shallow groundwater resources in this study area to protect children from diseases associated with the drinking of high-F⁻ shallow groundwater (Nawale et al. 2021).

Furthermore, as shown in Figure 9, within the study area, there is a clear east-to-west increase in the non-carcinogenic health risks associated with the consumption of shallow groundwater. Notably, the population living in the north district of the Xikuangshan antimony mining area is generally exposed to the highest non-carcinogenic health risks, with HQ values exceeding 4.00. Alarmingly, the population living in the north of the Xikuangshan antimony mining area is identified as being at the highest risk of developing dental and skeletal fluorosis, which is consistent with our measurement of high-F⁻ shallow groundwater concentrations (15.00 mg/L) in the study area.

In this study, 39 shallow groundwater samples were collected from the largest antimony mining area in Hunan Province, China, to determine the spatial distribution, formation mechanism, and health risk of high-F-shallow groundwater. The results can be summarized as follows.

The F⁻ concentrations in shallow groundwater were found in the range of 0.08–15.00 mg/L (mean: 1.21 mg/L), with 25.64% of the samples having F⁻ concentrations higher than reported in the Chinese national standard for drinking water (GB 5749-2022: 1.00 mg/L) and 20.51% of the samples having F⁻ concentrations that exceed the WHO recommended limit of 1.50 mg/L. Furthermore, the concentrations of fluoride in shallow groundwater clearly increase from the east to the west, with the highest F⁻ concentrations being detected in the north district of the Xikuangshan. High F⁻ concentrations were typically detected in shallow groundwater with a Ca–Mg–Cl–SO₄-type water chemistry, and the transition from a Ca–Mg–HCO₃ type to a Ca–Mg–Cl–SO₄ type was found to be helpful for the enrichment of F⁻ in shallow groundwater.

Principal component analysis revealed that the main source of F⁻ in shallow groundwater samples is cation exchange, accounting for 73.40%, followed by dissolution and precipitation of F-bearing minerals (15.10%) and human influence (11.50%). Overall, geological processes, including weathering-related cation exchange and the dissolution and precipitation of F-bearing minerals, are the predominant factors influencing the F⁻ content in shallow groundwater in the Xikuangshan antimony mining area.

Among the population age categories assessed, children were found to have the highest proportion of individuals (36.38%) with an F⁻ intake exceeding the HQ safety limit (1.00), followed by adult males (23.12%), teenagers (22.21%), and infants (21.22%). Thus, children are considered to have the highest non-carcinogenic health risks associated with the intake of F⁻ in shallow groundwater. Most of the infants, children, teenagers, adult males, and adult females with an HQ value exceeding 4.00 occurred in the north district of the Xikuangshan antimony mining area.

Despite F⁻ enriched shallow groundwater in the study area, the majority of residents are not aware of the risks of fluorosis from drinking. Our results are of limited use in predicting seasonal distribution and supporting informed management of F⁻ pollution in shallow groundwater. A major health danger to the populace can arise if high-F⁻ shallow groundwater is drank for an extended length of time without previous treatment. It is therefore strongly recommended to carry out the pretreatment of drinking water to ensure the safety of drinking water in the study area.

This work was supported by the Natural Science Foundation of Hebei Province (D2021508004), the Open Fund of Key Laboratory of Mine Water Resource Utilization of Anhui Higher Education Institutes, Suzhou University (Grant No. KMWRU202101), and the Open Fund of State Key Laboratory of Coal Resources and Safe Mining (Grant No. SKLCRSM22KFA03).

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.