Introduction

Heavy metals (HM) are ubiquitous pollutants owing to their toxicity, persistent and bioaccumulative characteristics in the environmental matrices including water (Zamora-Ledezma et al. 2021), sediment (Usman et al. 2021), soil (Ullah and Muhammad 2020), and living beings (Yuvaraj et al. 2021b), which has been primly focused in recent decades (Muhammad and Ahmad 2020). The soil ecosystem retains the toxic HM by influencing their mobility and detoxification serves as a sink (Bai et al. 2019; Mounissamy et al. 2021; Oruko et al. 2021). These processes resulted in a high level of HM and other contaminants deposition and accumulation, and its quality deterioration that reduced the buffering capacity of the soil (Sun et al. 2010; Wang et al. 2020). Soils HM sourced from various natural or geogenic (weathering and erosion of bedrocks and ore deposits) (Saddique et al. 2018) and anthropogenic (agrochemical, industries, mining, atmospheric deposition, and wastewater) activities (Xu et al. 2020; Yang et al. 2020).

Soil contamination resulted in alteration of soil structure and reduced its fertility, disturb the balance between residing biota, lead to contamination of groundwater, plants, and food chains, and had threatened the living organisms (Gałwa-Widera 2021; Kumar and Bhattacharya 2021). The HM are associated with toxicity in exposed living beings including DNA damage, induction of oxidative stress, immune system problem, and carcinogenesis (Köktürk et al. 2021; Omidifar et al. 2021).

Earthworms are one of essential species and the dominant members of soil macrofauna (Khan et al. 2013; Khan et al. 2012; Šrut et al. 2019). For risk assessment, the earthworms are considered to be a key bio-indicator, used in various environmental studies to evaluate the potential of HM bioaccumulation and food chain transfer (Liu et al. 2020; Verma et al. 2021). Earthworms live in soil and have close contact with it both internally and externally (Ruan et al. 2021). They uptake the soil and its HM ingredients such as lead (Pb), zinc (Zn), copper (Cu), chromium (Cr), and cadmium (Cd), and bioaccumulate in their bodies (Richardson et al. 2020; Yuvaraj et al. 2021a), with few species presentation high tolerance to soil contamination, showing variances related to ecological groups or physiological adaptation and tolerance to HM (Mahohi and Raiesi 2021; Wu et al. 2020). The variation of HM bioaccumulation appears to be dependent on their bioavailability (Yuvaraj et al. 2021a) as affected by soil factors affecting metal solubility (pH and redox) and complexation with organic matter (Bandara et al. 2021; Nejad et al. 2021). The activity of earthworms can improve soil’s nutrient cycling, decomposition of organic portion, structure, porosity, and fertility (Lakshmi et al. 2020; Sofo et al. 2020). Furthermore, due to their ability to transform and accumulating toxic HM in their bodies, earthworms are said to be the best bio-indicators in soil ecosystems (Ezemonye and Enete 2021; Šrut et al. 2019).

Meanwhile, many investigations have been conducted on bioaccumulation of HM in earthworm bodies (Boughattas et al. 2019; Šrut et al. 2019) on exposure to artificially contaminated soil or single land type. However, no study has been conducted regarding the introduction of earthworms to various kinds of long-term HM contaminated soils of various land types. More focus needs to be paid towards the assessment of HM and their uptake in earthworms through soil that is exposed to long-term contamination either through natural processes or anthropogenic sources. To fill the gap in existing research, this study was designed to investigate the uptake of toxic HM by earthworms, assess the bioaccumulation of metals within their bodies, and the adverse effects of HM on body weight/biomass after exposure to soils collected from different land types i.e. urban, rural, industrial, and mining sites of Khyber Pakhtunkhwa, Pakistan.

Material and methods

Soil sampling and analyses

Soil samples (n = 40) at the depth of 0–20 cm were collected from twenty sites of four land types in Khyber Pakhtunkhwa, Pakistan i.e. (1) urban area of Peshawar, (2) industrial estate at Hayatabad, (Peshawar district), (3) rural zone of Mardan (Mardan district), and (4) mining sites of Mohmand District (Mohmand District) as shown in Fig. 1. The collected soil samples were stored in polyethylene zip-locked bags. Each soil sample was divided into three parts: one for basic parameters measurement, the second for HM analyses, and third was stored at 4 °C to use for pot experiments.

Fig. 1
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Location map of the sampling sites in the study area

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Soil samples for basic parameters measurements were air-dried for 10 days and sieved through 2 mm mesh size. Dry soil samples were mixed with deionized water, measured for the pH and electric conductivity (EC) using a multi-water quality analyzer (CONSORT 6030, Belgium) (Das and Maiti 2008). Soil texture was determined using Fritsch Analysts 3 PRO Sieve Machine. Particle size infraction for sand, silt, and clay was measured (Barman and Choudhury 2020). Organic matter (OM) of the soil was determined using the loss on ignition (LOI) technique (Muhammad et al. 2011).

Acids digestion was carried out to find out the HM contents in oven-dried soil. For this purpose, 1.0 g of soil was taken in acid-cleaned digestion tubes and 10 ml of aqua regia (HNO3: HCl (3:1) was added to it and kept the tubes overnight. On the following day, the tubes were kept in digestion block and heating was turned on. The solution was heated at 80 °C for 1 h. The solution was allowed to cool down, and then, 5 ml of HClO4 was added to the tubes and heated at 160 °C until the solution became clear. The digested solutions were filtered and diluted to 50 ml with deionized water. The solutions were preserved at 4 °C until further analysis (Khan et al. 2016).

Earthworm sampling

Earthworms belonging to genus Pheritema were collected from parks in Peshawar, Pakistan, and then, they were added to metal-free clean soil for two weeks to eliminate HM from their body. The earthworms were identified with help of an expert in the Department of Zoology, University of Peshawar. After this, they were added to the soils collected from different land types, as given in the design of experiments.

Experimental design of pot experiments

Soil mass of 2 kg was taken into pots and moisture content was adjusted with deionized water. Five earthworms of similar size and weight were added to each pot and exposed to the collected soils for 14 days. Earthworms were kept in a dark chamber where a constant temperature of 12–18 °C was maintained. Soil moisture was checked and maintained up to 19% by weight technique. The weight of the earthworm was checked before the start of experiment, 7 days and finally at 14 days times. As per requirement, the collected earthworms were washed with distilled water and weighed. The mortality rate of soil earthworms and their biomasses were also determined after exposure to soils collected from four different land types.

Earthworm’s digestion for heavy metals

Acid digestion was also used to check the bioaccumulated HM concentration within the body of the earthworm. After 14 days of exposure, the earthworms were washed, weighed, and then, they were frozen for 48 h in Petri dishes. The freeze earthworm samples were digested using 3 ml of HNO3 and 1 ml of HCl for 45 min at 105 °C in digestion tubes. Then, 5 ml of deionized water was added and heated at 80 °C for a further 30 min (Omouri et al. 2018). The digested samples were then diluted with deionized water up to 50 ml and stored at 4 °C until the quantification of HM contents.

Analytical procedure accuracy and precision

Pots, glasswares, and plasticware were washed with 10% HNO3 solution and deionized water. Reagent blanks were also included as a control (Leveque et al. 2013), while reference soil material was used for precision and accuracy. The required elements Cr, Mn, Cd, Ni, and Pb were analyzed in all the digested samples of soil and earthworm using Atomic Absorption Spectrophotometer (AAS, Perkin Elmer 700) in the Centralized Resource Laboratory (CRL), University of Peshawar, Pakistan. Samples were measured in triplicates and mean values were used for results interpretation. Reproducibility of HM concentrations in samples was observed at 91 ± 6%. All reagents used in experimentation were of analytical grade.

Statistical analysis

For the calculation of range, mean, and standard deviation, Microsoft Excel 2017 was used. SigmaPlot 12.5 was used for the graphical presentation of data and one-way ANOVA was performed using SPSS 25 (SPSS Inc., Chicago, IL, USA).

Results and discussion

Soil basic properties

Soil basic parameters for various land types are summarized in Table 1. The highest mean pH of 8.2 was observed for the soil collected from the mining land type, while the lowest 7.9 for soil collected from the industrial land type. Maximum mean soil EC values of 15.9 mS/cm were found for industrial land type, while the minimum 5.7 mS/cm for mining land type. Soil OM showed variable levels in four selected land types. The highest OM content (0.32%) was observed for soil collected from rural land type and lowest (0.12%) for industrial land type (Table 1). The texture of the soil was classified as clay for urban, loamy clay for industrial, sandy clay for rural, and sandy for mining land types as per the United State Department of Agriculture (USDA) triangle method classification (Barman and Choudhury 2020). Soil basic properties determine the bioavailability of HM (Watson et al. 2021). Soil OM content and pH play an important role in EC and bioavailability of HM in the soil ecosystem. Soil pH level facilitates the availability of HM in the soil solution because of the metal’s capacity to precipitate (Silveira et al. 2003). The mobility of HM in the soil was affected by the dispersion sites of the adsorbent, its crystallinity, and its morphology as well (Visa and Chelaru 2014).

Table 1 Basic properties of soil collected from different land types (n = 40) in the study area
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Heavy metal concentrations in soil

Heavy metal concentrations of various land types in the study area are summarized in Table 2. Among HM, the highest mean concentration (692 mg/kg) was observed for Mn in soil collected from the mining land types and the lowest for Cd (1.5 mg/kg) of industrial land types. The rest of HM concentrations were observed within these two extremes (Table 2). The highest concentrations of Mn were attributed to its natural abundance compared to other studied HM in this land type. The results revealed that mean Pb concentrations of studied land types were observed within normal agriculture soil limits. The soil collected from mining areas was prone to be highly contaminated with Mn, Cr, and Ni. Higher contamination of HM in the mining areas was due to mining and exploration activities of chromites and natural denudation of ore deposit and bedrocks (Shah et al. 2014).

Table 2 Mean concentrations of metals (mg/kg) in soils collected from different land types in the study area
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The availability of HM in soil depends not only on the properties of soil but also on the concentration and solubility of metals. The solubility of pollutants can influence metal bioavailability in the environment (Garcia-Carmona et al. 2017). The highest quantity of Cr and Ni was observed in soil collected from the mining land types. The concentrations of HM in soils of mining land type only had surpassed the safe limits set by SEPA-China (1995) and US-EPA (2002), except for Pb. The presence of HM was also influenced by the physicochemical properties of soil (Zhao et al. 2021).

Heavy metal concentrations in earthworms

Heavy metals showed variable concentrations and accumulation patterns in earthworms exposed to soils of different land types, as summarized in Table 3. The Cr and Pb accumulation in earthworm collected from the soil of mining land type were the highest 10.7 and 71.9 mg/kg, respectively. The highest concentrations of Mn (37.6 mg/kg), Cd (7.00 mg/kg), and Ni (3.92 mg/kg) were observed in earthworm collected from rural, industrial, and urban land type’s soils, respectively. The results revealed that earthworms are more receptive towards Pb than any other studied HM. They can accumulate Pb for a longer period than any other HM due to its low detoxification that is why the earthworm showed a higher concentration of Pb within the body. The intakes of HM in earthworm take place through dermal contact with the soil or by ingesting a bulk of soil via the mouth (Richardson et al. 2020).

Table 3 Heavy metal concentrations/accumulation (mg/kg) in the earthworms collected from different land types in the study area
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The distribution of HM in the soil can be considered an important factor helping in their bioavailability to earthworms (Huang et al. 2020; Shi et al. 2020). Soil results showed the highest concentrations for Cr and Ni (as discussed earlier), while the accumulation in earthworms was the highest for Pb. The uptake of a metal by an organism depends upon the pathway that it adapts i.e. either through ingestion or through its tegument (Handy et al. 2021). Various studies observed that earthworm accumulates HM within its body for a longer period (Baccaro et al. 2021). The uptake of HM is also dependent on various soil factors and the earthworm (Wu et al. 2020). HM is bounded to the food so if the soil is less saturated with water, ultimately, more soil will be taken up causing more uptake of the respective metal. However, other metal like Cd is easily taken up in higher water by the earthworm (Leveque et al. 2013). HM are also competitive among themselves. Pb in comparison to Cd is more competitive that is why it is taken up easily by earthworm and can be retained for a longer period within the body compared to Cd within the body that is why Pb showed high concentration in comparison to other metal within the body of earthworm.

Changes in earthworm biomasses

The decreases in earthworm biomasses were observed in the soils of all land types (Fig. 2). The weight of earthworms was checked before the start of the experiment, at 7 days, and 14 days times. After 7 days of exposure, earthworms were washed, cleaned, and weighed again in order to observe the change in body weight of earthworm. The highest biomass reduction was observed for soil collected from industrial land types, followed by urban and is the lowest in rural land types at 7 days (Fig. 2). The maximum decrease in biomass of earthworm was observed for polluted soils collected from urban, mining, and rural land types, while the lowest reduction was observed for earthworm in soils of industrial land types at 14 days. These results showed that HM exerts stress after feeding by earthworms that had led to reduction of their biomasses. Various studies have observed biomass reduction of earthworms with exposure to HM and other environmental stresses (Khan et al. 2013, 2012; Yue et al. 2021). A study showed that increase in the concentration of contamination showed higher DNA damages in the earthworm species making it most suitable candidate for using as bio indicator (Khan et al. 2012).

Fig. 2
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Biomass changes in earthworms collected form soils of various land types in the study area

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Statistical analysis

One-way ANOVA results showed that HM concentrations in the soil of various land types contributed significantly (p < 0.05) different to the mean contamination except for Pb (Table 4). This significantly different HM contamination of soil was attributed to variation in the land types. In each land type, various factors contributing to HM were involved such as in rural areas natural denudation of bedrock and ore deposit the only source of contamination. The sources of contamination for mining-impacted soil could be mining activities in addition to natural denudation of bedrock and ore deposit. For industrial and urban land types, the sources of HM contamination were industrial and domestic wastewater, industrial and vehicular emission. The bioaccumulation of Mn and Ni concentrations in earthworms were significantly (p < 0.01) varied (Table 5). These higher the variations could be attributed to change of HM concentration in feeding soil of various land type. Table 6 summarized the Pearson correlation of physicochemical parameters concentrations in the soil of the study area. Results revealed a significant correlation between the metal pairs such as Mn-Ni, Cr-Ni, and Cr-Mn (Table 6). These significant correlations between the metal pairs could be attributed to common source of contamination.  

Table 4 One-way ANOVA for heavy metals in soil collected from different land types in the study area
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Table 5 One-way ANOVA for heavy metals in earthworm collected from soil of different land types in the study area
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Table 6 Pearson correlation of physicochemical parameter concentrations in soil of the study area
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Conclusions

This study concluded that HM concentrations of soil among the different land types and its bioaccumulation in earthworms were significantly (p < 0.05) varied. Among HM concentration, the highest accumulation was observed for Pb and the lowest for Ni in the studies earthworms. Variations of the HM concentrations in soil of land types and bio-uptake resulted in variable reductions in biomasses of earthworms. The highest biomass reduction was observed for industrial and mining land types. It means that the stresses caused by HM uptake depend not only on their accumulation rates but also on their long-term retention in the earthworm bodies. This study was limited to four heavy metals in soil of different land types and their accumulation in earthworms. It is recommended that in future studies, that physicochemical factors along with HM speciation need to be measured for the detailed influence on their accumulation in earthworm.