Introduction

There has historically been large-scale use of organochlorine pesticides (OCPs) in agriculture due to their various benefits, including cost-effectiveness, efficiency, and their broad spectrum of toxicity to insects (Shen et al., 2021; Tang et al., 2023). However, their harm to both soil ecology and human health became increasingly clear due to their high levels of toxicity, persistence, and tendency to bioaccumulate (Chen et al., 2021a; Huang et al., 2021). Consequently, the Stockholm Convention listed certain OCPs, such as hexachlorocyclohexanes (HCHs) and dichlorodiphenyltrichloroethanes (DDTs), as class I carcinogenic organic compounds, and China prohibited their use in 1983 (Li et al., 2020; Sharkey et al., 2020; Yu et al., 2014).

However, recent studies have detected the presence of residues of these restricted pesticides, such as HCHs and DDTs, in various environments, including cities and agricultural areas (Khuman et al., 2022; Tesi et al., 2022; Yu et al., 2019). These findings suggest the persistence of OCPs in the environment, even after their use has been halted for decades (Khuman et al., 2022). Several recent studies have focused on the presence and distribution of OCPs in farmland (Kafaei et al., 2020), watersheds, river and lake sediments (Gandla et al., 2023), and dust and soil in urban areas (Aslam et al., 2021; Skrbic & Marinkovic, 2019). Although most of these studies focused on both traditional and emerging OCPs, the focus in environmental monitoring has remained on HCHs, which often account for 10–50% of OCPs and represent the highest risks to ecosystem and human health (Wang et al., 2023). There is still environmental contamination by OCPs, despite the gradual phasing out of their use (Vudamala et al., 2023; Zhang et al., 2023). Moreover, the distribution of OCPs in the environment is often dependent on the pollution source, which can be broadly categorized as mainly non-point sources generated by agriculture (Chen et al., 2021b; Cui et al., 2017). Furthermore, OCP contamination has been detected in areas very far from the original source of contamination (Tao et al., 2008). While their persistence in the environment may be largely due to their constant migration and transformation, it is possible that there still are neglected sources of OCP contamination in the environment (Siddique et al., 2023a, 2023b).

Historical sites of pesticide production are important sources of OCP contamination (Ma et al., 2020). Analysis of the pesticide residues in these sites is helpful for determining the characteristics of HCH persistence in the environment. Since their closure, many of these historical sites of pesticide production in China were gradually transformed into commercial and residential areas (Wang et al., 2023; Zhang et al., 2022). However, residues of OCPs often persist at these sites. Localized residual soil OCPs can be transferred through volatilization, diffusion, and mass flow processes, leading to pollution of the atmosphere and surface/groundwater, as well as bioaccumulation up the food chain (Derouiche et al., 2023; Siddique et al., 2023a). Therefore, pesticide residues at these historical sites of pesticide production pose risks to both current and future residents. There is a need to improve the understanding of soil pesticide residues in urban land planning and management.

The rapid urbanization of China has emphasized the need to develop land that was historically used for other purposes (Ai et al., 2022; Chen et al., 2022). However, pesticide contamination of land in China places limits on urban development, and not considering residual soil pesticide pollution can pose risks to ecological and human health (Yu et al., 2019, 2020). While pesticide residues tend to concentrate on the surface of the soil, they can also migrate to groundwater or deep soil layers (Vryzas et al., 2018).

Therefore, the aim of the present study was to characterize the occurrence and distribution of HCHs and underlying processes at sites of historical pesticide production in southern China. The objectives of the present study were to (1) characterize the soil HCH residues at sites of historical pesticide production and responses to the functional areas; (2) analyze the distribution, composition, and migration of HCHs to deep soil, surface water and groundwater; and (3) evaluate the presence of HCH residues in the soil of lands now used for urban development.

Materials and method

Study area

The present study examined a site of historical HCH production in Zhanjiang city in southern China. This site is located in the monsoon marine climate with a mean annual temperature of 22.7–23.5 ℃ and precipitation of 1395.5–1723.1 mm. The pesticide production plant covered an area of ± 40,000 m2 and produced HCHs in powder form from 1966 to 1979. Raw materials used at the plant were mainly benzene and chlorine gas, and the production process mainly included pesticide synthesis, distillation, drying, and packaging. The factory was spatially segregated into areas for pesticide synthesis, packaging, storage of finished-product, sewage-treatment, and offices (Fig. 1 and Fig. S1). The site is characterized by rice-type red soil divided according to formation into quaternary filling and alluvial soil layers (Fig. 2): (1) filling soil (0–1 m); (2) miscellaneous soil (0–2 m); (3) alluvial soil (0–12.0 m); (4) sand silt (0–13.0 m); and (5) clay (0–6.0 m). The site contains shallow groundwater at a depth of 5.35–8.36 m, which is concentrated in the silt layer and flows from northeast to southwest (Figs. 1 and 2). There are plans to develop the site into urban land in the future.

Fig. 1
figure 1

Diagram of the site. Sewage-treatment area (ST); synthesis area (SA); drying packaging area (DPA); finished-product storage area (FPSA); office area (OA)

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Fig. 2
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Structural characteristics of the underground soil layer

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Sampling, pretreatment, and analysis

The present study adopted grid distribution and encryption distribution methods and established 79 soil-monitoring sites according to the functional layout of the production area. Of the 475 soil profile samples collected, 132, 242, 29, 43, and 29 were from the areas involved in synthesis, dry-packaging, sewage-treatment, storage of the finished-product, and the office area, respectively, and were collected from different depths. Moreover, the study characterized the hydrogeological features of the site and collected 10 groundwater and 4 surface water samples to explore patterns of HCH migration. Soil and water samples were collected in October 2020. Each 1,000-g soil sample was sealed in a brown glass bottle, frozen, and maintained in the dark during transport to the laboratory, after which it was stored at − 18 ℃ pending analysis. During analysis, soil samples were freeze-dried for 24 h and then ground and filtered through 1-mm filters for the analysis of α-HCH, β-HCH, and γ-HCH. The analysis process followed the protocol of Soil and Sediment-Determination of Organochlorine Pesticides-Gas Chromatography (HJ 921-2017). Briefly, HCHs in the soil were prepared by Soxhlet extraction. Soil samples (0.50 g) were extracted using 100 mL of dichloromethane for 24 h, after which the extract solutions were concentrated to < 2 mL via rotary evaporation at 40 °C, and transferred onto a solid-phase extraction column that had been conditioned by successive passages of 10 mL of dichloromethane. After the loading of the concentrated samples, the HCHs were eluted from the column using 10 mL of dichloromethane. The eluents were concentrated to 1 mL via rotary evaporation at 40 °C and stored at 4 ℃, pending analysis. The 1-L water samples were enriched using an SPE cartridge with the speed of 1 drop/s after which the material was eluted with 10 ml DCM. The eluate fractions were collected and concentrated to 1 mL, pending analysis.

Samples were analyzed by gas chromatography (GC, 7980, Agilent J & W Scientific, Folsom, CA, USA) coupled with ECD detector. One microliter of sample was injected directly onto the column (30 m × 0.32 mm × 0.25 μm, Agilent J & W Scientific) using the splitless mode. (The temperature of the injection port was 220 °C.) The carrier gas was high-purity nitrogen at a flow rate of 2 mL per min. The process was conducted in an oven as follows: the GC oven temperature was initially set to 100℃, increased to 220 °C at a rate of 15 °C/min, maintained for 5 min, and then increased to 260 °C at the same rate, holding for 20 min. The temperature of the detector was 280 °C.

Quality assessment/control (QA/QC)

Prior to sample processing, all glassware was calcined at 450 ℃ for 4 h, and 10% of samples (n = 50) were tested in parallel. The limits of detection (LODs) of α-HCH, β-HCH, and γ-HCH in the soil were 0.07 mg kg−1, 0.06 mg kg−1, and 0.06 mg kg−1, respectively, while the LODs in water were all 0.01 μg L−1. Their recoveries were 59.3–113.5%, 65.2–98.3%, and 78.0–135.7%, respectively. Table 1 summarizes the screening and intervention values of α-HCH, β-HCH, and γ-HCH (GB 36600-2018) in the soil.

Table 1 The detection frequencies (DFs) and concentration of soil HCHs at the site (mg kg−1)
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Data processing

SPSS statistics 26 and Microsoft Office 2016 were used for data processing, while Origin 2023 and ArcGIS 10.6 were used to plot graphs. Data below the LOD were replaced by half the LOD. Significance was assumed at p < 0.05.

Results and discussion

Occurrence of soil residual HCHs

Table 1 summarizes the detection frequencies (DFs) and concentration of soil HCHs at the site. The DFs of α-HCH, β-HCH, and γ-HCH were 18.1%, 21.3%, and 16.2%, respectively, with concentrations of ND–70.80 mg kg−1 (mean of 0.45 mg kg−1), ND–98.50 mg kg−1 (0.44 mg kg−1), and ND–17.60 mg kg−1 (0.17 mg kg−1), respectively. The mean concentrations of α-HCH and β-HCH exceeded China’s Class 1 screening values (0.09 mg kg−1 for α-HCH and 0.32 mg kg−1 for β-HCH) in the soil (GB36600—2018, Soil environmental quality—risk control standard for soil contamination of development land, Ministry of Ecology and Environment of the People's Republic of China). The ranking according to the degree that they exceeded the screening value was α-HCH (17.26%) > β-HCH (5.68%) > γ-HCH (2.74%). Notably, there were many samples with extreme concentrations throughout the site; for instance, the maximum concentrations of α-HCH, β-HCH, and γ-HCH exceeded the screening values by factors of 787, 308, and 28.4, respectively, suggesting strong contamination. Additionally, the soil ∑3HCHs had a DF and concentration range of 24.6% and ND–186.90 mg kg−1 (mean of 1.05 mg kg−1), respectively, reflecting the persistence of HCHs in the site. Table S1 shows a comparison of the pesticide residue concentrations in elsewhere; the maximum value of HCHs reached 186.90 mg kg−1 in the site, which can be regarded as a moderate residue level. The coefficient of variation (CV) can reflect the spatial heterogeneity of HCHs in the soil, thus indicating possible anthropogenic effects (Khademi et al., 2019), classified as high CV ≥ 1, moderate CV < 1. The CV of HCHs reaching 6.24–11.5 (Table 1) indicated high levels of the effects of anthropogenic factors on their distribution.

Although there is currently reduced attention paid to HCHs in the urban environment, there is, nevertheless, evidence of HCH residues, indicating their continued relevance especially at sites contaminated with HCHs (Ma et al., 2020). The site investigated in the present study is surrounded by urban land, including schools, hospitals, and residences, and there are plans to develop the site itself into residential land. However, the presence of high residual soil HCHs at the site present limits to its development (Balazs et al., 2020). There is, therefore, an urgent need to reduce residual soil HCH prior to land-use planning to facilitate safe development in the transformation of functional areas during urban development (Wang et al., 2016; Yun et al., 2022). Moreover, the high levels of HCHs in the soil at the contamination source usually spread into the surrounding areas through diffusion. Thus, the investigation of the transfer of contamination from soil to water is of significance.

Characteristics of residual HCHs in soils in relation to areas with different historical functions

The concentration of residual HCHs in soils in relation to historical function

Figure 3 and Fig. S2 show the spatial distribution and relative proportions of α-HCH, β-HCH, and γ-HCH among the different functional areas. The results indicated that the concentrations of α-HCH exceeded those of β-HCH, and γ-HCH in most of the functional areas. The highest β-HCH residues were detected in the dry-packaging area, while only β-HCH was detected in the office area. The concentration of α-HCH in the finished-product area (1.50 mg kg−1) even exceeded the risk intervention value (0.9 mg kg−1). The maximum concentrations of α-HCH were in the dry-packaging and finished-product areas where they exceeded the screening value by factors of 787 and 504, respectively. The DF of α-HCH reached 37.9% in the sewage-treatment area, exceeding the values of the finished-product area (20.9%), the dry-packaging area (20.7%), the synthesis area (12.1%), and the office area (0%). β-HCH was detected in all functional areas, with its highest mean concentrations in the finished-product area (1.28 mg kg−1) and dry-packaging area (0.58 mg kg−1), and exceeded the screening value (0.32 mg kg−1), but not the controlled value (3.2 mg kg−1). The highest β-HCH concentrations were observed in the dry-packaging and finished-product areas, exceeding the screening value by factors of 307.8 and 145.6, respectively, indicating a high residue risk. In addition, the ranking for β-HCH concentrations exceeding the screening value was the sewage-treatment area (13.8%) > finished-product area (11.6%) > dry-packaging area (6.6%) > synthesis area (1.5%) > office area (0%). However, the ranking of the functional areas according to DFs of β-HCH was the finished-product area (30.2%) > sewage-treatment area (27.6%) > dry-packaging area (24.4%) > synthesis area (12.9%) > office area (6.9%). While γ-HCH was not detected in the office area, its highest mean concentrations in the finished-product area (0.33 mg kg−1) and the dry-packaging area (0.23 mg kg−1) were both within safety limits. Moreover, the ranking of the functional areas according to the DF of γ-HCH was the sewage-treatment area (27.6%) > finished-product area (25.6%) > dry-packaging area (19.0%) > synthesis area (7.6%) > office area (0%). The highest mean concentrations of ∑3HCHs were found in the finished-product area (3.11 mg kg−1) and dry-packaging area (1.33 mg kg−1), both of which exceeded the screening value (1.02 mg kg−1), while the sewage-treatment area (0.50 mg kg−1), synthesis area (0.18 mg kg−1), and office area (0.10 mg kg−1), were at safety levels. The ranking of the functional areas in terms of DF of HCHs was the sewage-treatment area (37.9%) > finished-product area (32.6%) > dry-packaging area (27.3%) > synthesis area (16.7%) > office area (6.9%).

Fig. 3
figure 3

Spatial distribution of HCHs in the site. Sewage-treatment area (ST); synthesis area (SA); drying packaging area (DPA); finished-product storage area (FPSA); office area (OA)

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The above results suggest that the soil HCH residues at the study site depended significantly on the source (Liu et al., 2022), seen especially in the high concentrations observed in the finished-product and dry-packaging areas. The maximum soil residual HCH was in a soil sample from a depth of 0–1 m in the dry-packaging area, consistent with the characteristics of most contaminated soils (Liu et al., 2022; Ma et al., 2020). The high concentration of HCHs in this area could possibly be attributed to the spilling and spreading of powdered HCH on the soil surface during the drying and packaging processes, followed by further migration and diffusion of HCH into the soil (Balazs et al., 2021; Ma et al., 2020). However, the higher DFs in the sewage-treatment areas reflected a difference compared with the high-concentration areas. Drainage in the area may show increases in the DFs and decreased concentrations. HCHs can be easily transferred to water systems, resulting in an increase in their DF values. However, the HCHs usually show > 70% removal in wastewater treatment systems after denitrification, resulting in lower concentrations in the sewage-treatment area (Cong et al., 2009).

Detection of HCHs at different depths

The maximum depths of α-HCH, β-HCH, and γ-HCH detection were all observed in the dry-packaging area, reaching 24 m, indicating significant migration and high penetration of the HCH pollution. Moreover, the maximum detection depth of both α-HCH and γ-HCH was found to be 19 m in the synthesis area, much deeper than the depth of 12.5 m observed for β-HCH. In the sewage-treatment area, the values were similar, between 14.5 and 16.5 m. However, the β-HCH concentrations were higher than those of α-HCH in the finished-product storage area but were not higher than those in the aquifer. Only β-HCH was detected in the office area, and its maximum detection depth was only 1 m. The ranking of the maximum detection depths of ∑3HCHs was the dry-packaging area (24 m) > synthesis area (19 m) > sewage-treatment area (16.5 m) > finished-product area (10.5 m) > office area (1 m). It was apparent that the residual HCHs gradually migrated and were retained in the deeper soil following the decommissioning of HCH production four decades ago, implying widespread and long-term contamination. The 24 m depth of detection in the drying-packaging area may have been related to the higher residue level, resulting in penetration. However, the finished-product area also had high HCH concentrations, which did not fully correspond with the detection depth, suggesting that other factors in the soil may have influenced migration, preventing deeper migration into the soil (Sun et al., 2020). Interestingly, while only β-HCH was detected in the office area (at 1 m depth), its concentration did not exceed the screening value. The different functional areas of the factory were very close to each other, and the results indicate that appropriate protection measures can effectively prevent HCH contamination. In the factory, the surface runoff was affected by the terrain, and it was difficult to migrate to the office area. There may have been only limited migration of HCH through dust and other means. After decades, it was apparent that the levels of HCH residues had decreased gradually to below the LOD in most of the sites (Table 2).

Table 2 The HCHs concentration of soil in different functional zones (mg kg−1)
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HCHs in surface water and groundwater

Table 3 shows the concentrations of HCHs in surface water and groundwater. HCHs were only detected in one of the four surface water samples. This sample was collected from the dry synthesis area (with higher soil residue), showing α-HCH (2.45 μg L−1) and β-HCH (0.73 μg L−1). There was no detection of HCH in the other three samples from the sedimentation and clear tanks, which indicated that HCH contamination may have been present for a long time in the surface water. It is more useful to explore the groundwater characteristics of HCHs. As shown in Table 3, in the groundwater, the DFs of α-HCH, β-HCH, and γ-HCH were 70%, 100%, and 40%, respectively, whereas their concentrations were ND–256 μg L−1 (mean of 29.9 μg L−1), ND–35.8 μg L−1 (5.57 μg L−1), and ND–26.7 μg L−1 (3.18 μg L−1), respectively. The concentrations of ∑3HCHs were 0.285-319 μg L−1 (mean of 38.6 μg L−1). This result confirms the migration of historical HCHs into groundwater and indicates that this process has more serious consequences than those in surface water. HCHs in groundwater may further influence the accumulation of HCHs in deep soil, causing more widespread effects. Furthermore, the high mobility of groundwater could enhance the spread of HCHs, causing technical obstacles and increasing the workload involved in the remediation of HCH-contaminated areas. This is not surprising if we take into account the fact that the levels of OCPs in the water column are subject to very rapid and significant changes over time and that water is not effective for conserving HCHs, given their poor water solubility, but functions rather as a means of transportation (Pardo et al., 2021).

Table 3 HCHs in the surface water (SW) and groundwater (GW) (μg L−1)
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Feature ratio characteristics

The study analyzed soil HCH residues and their migration and transformation at the study site to assess the risks posed by their presence. In general, industrial HCHs are composed of 60–70% α-HCH, 5–12% β-HCH, and 10–15% γ-HCH (Walker et al., 1999) with a α-HCH/γ-HCH ratio of approximately 4–7 and a β-HCH/ (α + γ)-HCH ratio of approximately 0.06–0.17. In the present study, the proportions of HCHs varied widely in the soil environment. In terms of the functional areas, the α-HCH/γ-HCH ratio (Fig. S3) ranged from 1.98 (sewage-treatment area) to 4.54 (synthesis area), while the β-HCH/(a + γ-HCH) ratio ranged from 0.29 (sewage-treatment area) to 0.77 (dry- packaging area), all of which are higher than 0.17, indicating a dominant historical influence (Li et al., 2015; Meng et al., 2013). In the synthesis area, the high α-HCH/γ-HCH ratio may be related to a continued input from the potential source.

The α-HCH/γ-HCH ratio in the deep soil was 1.4–3.7 (Fig. 4), less than the 4–7 range associated with industrial HCH production, indicating a historical source (Li et al., 2015; Meng et al., 2013). In the 0–1 and 1–2 m soil, the α-HCH/γ-HCH ratios were both 3.7, which is close to the characteristics of industrial HCHs, suggesting a historical source of HCHs in the topsoil. In addition, in the 3–7 and 9–24 m soil, the ratios were less than 1.6, reflecting a difference from an industrial source and suggesting the presence of different osmotic effects in the soil. Notably, the ranking of the different HCH monomers according to their water solubility was γ-HCH (7.30 mg L−1) > α-HCH (2.00 mg L−1) > β-HCH (0.24 mg L−1). The 5–9 m soil covers the aquifer, and in the lower layer of the aquifer (7–9 m), the ratio was found to be significantly increased, possibly due to difference in the water solubility of HCHs, causing changes in the migration patterns of HCHs. Moreover, characteristic ratio of β-HCH/(α + γ)-HCH (0.21–1.04) in all layers was much higher than 0.17, indicating that residual HCH was the dominant source of HCH at the site. The situation was significantly occurred in the 0–1 and 1–2 m soil layer (1.04 and 0.86). Willett et al. (1998) reported that the presence of HCH residues may be related to their historical sources as well as differences in volatility, water solubility, and biodegradability, affecting their breakdown and migration in the soil environment (Wang et al., 2006). Since the different isoforms of HCH have different stabilities and toxicities, they differ in terms of their ability to undergo microbial degradation in the soil environment (Garg et al., 2016). The four decades since the discontinuation of HCH production at the site would result in different degrees of attenuation and degradation, together with changes in the HCH compositions (Gao et al., 2008; Wang et al., 2015). The ranking of the different HCH monomers in terms of de-chlorination associated with degradation was γ-HCH > α-HCH > β-HCH. γ-HCH is the most susceptible to degradation by soil microorganisms and can be transformed to α-HCH by photochemical reactions, while α-HCH can be metabolically converted to β-HCH. The proportion of α-HCH (specific) in the site soil was significantly reduced compared with those of industrial HCHs (60–70%), which occurred concurrently with an increase in β-HCH. This result indicated that α-HCH was converted to β -HCH, with the latter characterized by lower biodegradability, water solubility, and volatility in the complex soil environment (Mrema et al., 2013; Yadav et al., 2015).

Fig. 4
figure 4

Ratio characteristics of HCHs in different depth; a scatter diagram characteristics; b trend characteristics of mean value in the depth

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Migration and transformation of soil HCHs

Concentrations of soil HCHs in relation to depth

HCHs can not only degrade but also migrate to deeper soil layers. Figure 5 shows the vertical distribution of HCHs at the different soil depths at the historical HCH production site. The concentration of total soil HCHs as well as those of the three isomers rapid decreased with 0–5 m, with the lowest observation at 3–5 m, indicating significant downward penetration from the surface soil layers. In contrast, the concentrations of HCHs gradually increased with increasing soil depth between 5 and 14 m and decreased further at 14–24 m, although these changes were insignificant (p > 0.05). Surface backfill of construction waste occurred at the site, which could have contributed to the downward migration of soil HCHs through hardened surface cracks (Wycisk et al., 2013). Moreover, it is possible that reconstruction and demolition of equipment may have promoted the accumulation of soil HCHs, mainly in the 0–5 m soil layer. The depth of the aquifer at the study site is 5.35–8.36 m and is characterized by silty clay and a weak permeable floor layer. Therefore, HCHs would have accumulated in the bottom layer of the aquifer, resulting in a minor increase in the concentrations of HCHs in the soil layer below 7–9 m.

Fig. 5
figure 5

The concentration and composition of α-HCHs, γ-HCH, and β-HCH in different depths in the site

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The results indicated an initial decrease in the total concentration of soil HCHs with increasing soil depth, followed by a gradual increase. The mean concentrations of α-HCH, β-HCH, γ-HCH, and ∑3HCHs in the 0–1 m soil layer reached 1.46, 1.92, 0.40, and 3.78 mg kg−1, respectively, significantly exceeding that below 3 m (< 0.4 mg kg−1). The significant downward migration of HCHs observed at the site may have been possibly due to the wet weather in southern China. The soil structure is an important factor that controls the combination of soil organic matter and water infiltration with HCHs, allowing co-migration into the deeper soil (Lupi et al., 2019; Zhang et al., 2009). The typical geological profile characteristics of the site from top to bottom are miscellaneous fill or plain fill (0–2 m); sandy silt, silt, or sand layer (2–12 m), and clay layer (> 12 m). The water permeability of soil tends to gradually decrease with increasing soil depth. Moreover, shallow groundwater at the site is mainly present in the silt soil layer, with groundwater at a depth of 5.35–8.36 m. The migration of HCHs to the groundwater resulted in lateral migration with groundwater and consequent increased concentrations of HCHs in the gas zone and aquifer of the local block.

Compositions of soil HCHs in relation to depth

Figure 5 shows the vertical distribution of soil HCHs at different depths. The ranking of the HCH monomers according to concentration at a soil depth of 0–24 m was α-HCH (42.5%; mean of 34.9–58.3%) > β-HCH (41.7%; mean 17.6–50.9%) > γ-HCH: (15.8%; mean 10.6–33.7%). The proportions of the different HCH isoforms changed greatly compared to their original proportions, during which the percentage of β-HCH significantly increased from 5–12% to a 17.6–50.9% level. This result indicated that residual soil HCHs underwent significant environmental degradation and transformation (Ding et al., 2021; Liu et al., 2022). The concentrations of α-HCH, β-HCH, and γ-HCH decreased rapidly from the depth of 0–5 m, while increasing gradually with increasing soil depth up until 14 m, following which they gradually decreased at soil depths of 14–24 m. The HCHs showed a similar trend in terms of concentration while differing in their proportions. The higher permeability of γ-HCH and α-HCH compared to that of β-HCH may have led to the accumulation of β-HCH in the 0–7 m soil layer (> 29%). In contrast, the concentrations of α-HCH and γ-HCH in deep soil significantly exceeded that of β-HCH due to their stronger permeability and higher water solubility (Heeb et al., 2012). The differences between the upper and the deeper soil layers were also related to the permeability and degradability of HCHs (Ding et al., 2021). These differences in water permeability in addition to the ready conversion of γ-HCH to α-HCH under the action of microorganisms contributed to the accumulation of α-HCH in the 5–14 m soil layer (composed of silty clay and the aquifer) and its gradual decline at depths exceeding 14 m (Kranzioch-Seipel et al., 2016). The permeability of γ-HCH increased with increasing soil depth, and as microorganisms are not active under anoxic conditions (Langenhoff et al., 2013), the degradation of γ-HCH was significantly reduced as the soil depth increased. Thus, the proportion of γ-HCH increased gradually increased in the deep soil, accounting for 30–34% of HCHs in the 9–24 m soil layer compared with that of β-HCH.

The relative proportions of the different HCH monomers have been found to change according to the soil depth (Concha-Grana et al., 2006; Yang et al., 2005). The proportion of stable β-HCH generally decreased with increasing soil depth, whereas that of active γ-HCH gradually increased, while α-HCH showed relatively little change (Ma et al., 2020). This result implied that the different HCH monomers showed different capacities for migration due to the influences of soil characteristics in different soil layers (Liu et al., 2021). While all three HCHs monomers showed downward migration, they migrated at different rates (Hites & Venier, 2023). The faster downward migration of γ-HCH reflected its relatively rapid increase, whereas the gradual decline in β-HCH in deep soil could be attributed to its stability and resistance to migration with dissolution and penetration by water (Ashesh et al., 2022; Srivastava et al., 2022). Overall, the permeabilities of γ -HCH and α-HCH exceeded that of β-HCH, resulting in increasing and decreasing proportions of β-HCH with increasing soil depth, respectively.

Environmental impact and management

Re-use of historical sites is a common practice in urban land management. Contaminants at these sites can continue to pose risks to human and ecological health due to the lack of prolonged management and the historical production processes conducted at a site. Therefore, there is an urgent need for an assessment of risks posed by soil contaminants at sites before re-development during urbanization. Despite the passage of four decades since the cessation of pesticide production, there remains a need to investigate and analyze historical activities at a site. The waste generated by pesticide industries remains an important source of urban OCP pollution. The presence of soil OCP residues can limit the land that can be used for urbanization. The results of the present study indicated the presence of significant HCH soil residues at the site that were closely related to historical activities. Although the HCH concentrations decreased rapidly with increasing soil depth (0-5 m), they were found to gradually increase in the deeper soil layers, with high detection observed at a depth of 24 m, indicating serious pollution of the deep soils. Moreover, the groundwater showed higher levels of HCH pollution than surface water, augmenting the pollution risk of the deeper soil layers. Therefore, the possible risks posed by residue HCHs to the soil ecology and human health should be considered during urbanization. Since pesticide residues have wide-ranging effects on groundwater, sites with HCH contamination should undergo comprehensive evaluation and remediation.

Conclusions

The results of the present study indicate that the soil at the study site remains HCH contamination. Additionally, variations in HCH concentrations across different functional areas of the site suggest that the contamination originated from historical sources. The ranking of the different soil HCH species according to mean concentration was α-HCH (0.45 mg kg−1) > β-HCH (0.44 mg kg−1) > γ-HCH (0.17 mg kg−1), with α-HCH and β-HCH exceeding the Chinese Class 1 screening levels, indicating the presence of serious residual pollution. The residual soil HCH concentrations of the site differed according to the functional areas of the site, with the highest residues observed in the finished-product and dry-packaging areas. We confirmed the presence of obvious migration and transformation trends of the HCHs in the soil at the site, even four decades after the cessation of HCH production. The γ-HCH and α-HCH forms were transformed into the more stable β-HCH and were found at higher levels in the surface soil rather than the deep soil. HCHs in the surface soil readily migrated into the aquifer, even as far as a depth of 24 m, indicative of widespread and serious pollution. However, in different layers of soil, the HCH isoforms were found to be distributed according to their specific composition, mainly due to their different water solubilities and stabilities. The study revealed that HCH production sites are potential sources of soil HCH pollution even after four decades and that HCH can migrate to deeper soils and groundwaters, with reduced degradation with increasing depth, thus causing a more widespread harm. Thus, despite the present reduced attention on HCHs in urban environments, these chemicals still pose a significant risk when historical production sites are transformed and redeveloped during urbanization.