Influence of cadmium-contaminated soil on earthworm communities in a subtropical area of China

Applied Soil Ecology xxx (xxxx) xxx–xxx

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Influence of cadmium-contaminated soil on earthworm communities in a subtropical area of China ⁎

Kun Wanga, Yuhui Qiaoa, , Huiqi Zhanga, Shizhong Yuea, Huafen Lia, Xionghui Jib, David Crowleyc a

China Agricultural University, Beijing, China Hunan Academy of Agricultural Sciences, China c University of California Riverside, Riverside, CA, USA b

A R T I C LE I N FO

A B S T R A C T

Keywords: Heavy metal Ecotoxicology Earthworm communities Biological indices

Research was conducted to examine the influence of historical cadmium (Cd) contamination and soil properties on the biomass, diversity and structure of earthworm communities in a subtropical area (Hunan province) of South China. Fourteen earthworm species were identified across the twelve field sampling sites. Metaphire californica was the most widespread and dominant species. Results showed that both earthworm density and the Simpson diversity index decreased inversely with increasing soil Cd concentrations. The proportion of adult earthworms was greater in soils with high levels of Cd contamination. The abundance of earthworms was also correlated with soil organic carbon and total nitrogen contents. Cd concentrations in M. californica were well predicted by both the total and available soil Cd concentrations (R2 = 0.83, R2 = 0.90, p < 0.01, respectively), and suggested that this species may have particular applications for risk assessment and use in bioremediation.

1. Introduction Over several decades, the mining, smelting, and other industrial activities of metal ores has resulted in soil pollution, which is now a serious problem in many areas across the world. A joint report on the current status of soil contamination in China issued by the Ministry of Environmental Protection (MEP) and the Ministry of Land and Resources (MLR) of China indicates that cadmium (Cd) ranks first among the metals and metalloids of concern based on the percentage of soil samples (7.0%) that exceed the limits established by the MEP (MEP and MLR, 2014). Among the potentially toxic elements that are introduced into the environment, Cd is generally more mobile than other metals and can cause acute or chronic toxicity to living organisms (Sparks, 2003; Steinnes and Friedland, 2006; Kirkham, 2006; Smith, 2009; Alguacil et al., 2011; Margesin et al., 2011; Aghababaei et al., 2014). Recently, the concentrations of metals in soils in some areas of southern China such as Hunan Province still appear to be increasing. Minimizing the transfer of contaminants from soil to the food chain has therefore become a top priority, and numerous studies have been carried out in this area to examine the potential for Cd exposure to the food chain and methods for remediation of contaminated soils to better assure food safety (Zhao et al., 2015, Williams et al., 2009; Schreck et al., 2012; Austruy et al., 2013; Liu et al., 2013). However, there have been

⁎

relatively few studies focusing on how the earthworm communities vary with different levels of metal contamination in the field. Such information is needed as baseline information to evaluate of impact of metals on the earthworms, which may be used as bioindicators for risk assessement of soil contaminants. Earthworms are key members of the soil fauna, and have important roles in many soil functions (Nahmani et al., 2007), being involved in nutrient cycling, soil organic matter decomposition, and modification of soil structure, all of which affect plant productivity (Edwards and Bohlen, 1996; Butenschoen et al., 2009; Koutika et al., 2001). By incorporating organic matter into the soil and soil mixing, earthworms also influence the distribution and bioavailability of contaminants in soils (van Gestel et al., 2009). They are increasingly recognized as indicators of soil health and serve as ecotoxicological sentinel species that are constantly exposed to soil contaminants (Lanno et al., 2004; Suthar et al., 2008; Pérès et al., 2011; Lévêque et al., 2013). At present, a large set of ecotoxicological data on contaminated soils has been collected for diagnosis and evaluation of ecological risks. Most studies on metal toxicity to earthworms have been carried out using methods that examine the effects of short-term exposure to high doses of metals in freshly-spiked soil (OECD, 1984; Depledge and Fossi, 1994; Martikainen, 1996; Lowe and Butt, 2007; Ramadass et al., 2015; Asensio et al., 2007; Lijun et al., 2005; Hobbelen et al., 2006; Hankard

Corresponding author. E-mail addresses: [email protected], [email protected] (Y. Qiao).

https://doi.org/10.1016/j.apsoil.2018.02.026 Received 2 August 2017; Received in revised form 26 February 2018; Accepted 26 February 2018 0929-1393/ © 2018 Elsevier B.V. All rights reserved.

Please cite this article as: Wang, K., Applied Soil Ecology (2018), https://doi.org/10.1016/j.apsoil.2018.02.026

2

1.92 0.06 14.1

± ± ± ± ± ± ± ± ± ± ± ± 0.2 0.2 0.3 0.6 1.4 1.3 2.2 2.1 1.8 1.2 2.2 9.7 0.2 0.2 0.1 0.2 0.7 1.0 1.0 0.4 0.8 1.0 1.5 2.6

0.3

4.8 0.18 24.1

± ± ± ± ± ± ± ± ± ± ± ± 0.81 1.31 1.75 2.39 3.45 4.03 4.30 4.53 4.65 5.11 7.49 17.8 7.3 16.0 10.4 7.1 32.9 10.7 22.4 2.0 4.3 26.6 10.9 42.4

200

2.67 1.24 3.93 18.9 12.2 24.9 17.7 14.2 22.4 6.02 4.88 7.24 a

Threshold

Mean Maximum Minimum

Abbreviations: CEC, cation exchange capacity; SOC, soil organic carbon; TN, total nitrogen; DTPA-Cd, DTPA-extractable Cd. a Threshold based on Chinese Environmental Quality Standard for Soils (GB 15618-1995) (State Environmental Protection Administration of China, 1995).

50 250

119 45.7 528

54.1 12.9 229

176 95.2 505

± ± ± ± ± ± ± ± ± ± ± ± 123 130 131 140 164 127 227 129 131 276 160 374 24.9 ± 1.4 28.8 ± 3.2 55.0 ± 9.0 19.5 ± 1.8 207 ± 14 20.1 ± 1.4 24.4 ± 1.1 23.9 ± 1.8 39.8 ± 1.7 55.8 ± 3.2 217 ± 22.7 26.6 ± 1.6 66.8 ± 2.1 78.6 ± 6.9 67.2 ± 2.2 64.8 ± 5.0 102 ± 10.5 58.4 ± 4.0 78.7 ± 5.3 61.3 ± 4.4 127 ± 18.3 158 ± 8.9 429 ± 32.2 139 ± 16.4 0.2 0.3 0.1 0.4 0.3 0.3 0.3 0.3 0.1 0.2 0.3 0.2 ± ± ± ± ± ± ± ± ± ± ± ± 3.0 3.1 3.2 2.9 2.9 2.1 2.2 2.4 2.7 3.0 3.0 2.1 1.0 1.0 2.1 1.5 1.1 0.9 1.4 1.4 0.7 0.7 0.4 2.1 ± ± ± ± ± ± ± ± ± ± ± ± 21.2 20.8 18.2 18.5 18.5 18.3 17.0 15.8 19.5 21.4 20.8 18.6 0.9 0.4 0.3 0.7 0.8 1.0 0.3 0.4 0.6 0.5 0.3 0.4 ± ± ± ± ± ± ± ± ± ± ± ± 28°25′44″ 28°27′01″ 28 °28′15″ 28°21′59″ 26°32′46″ 28°24′34″ 26°59′51″ 28°21′48″ 27°06′05″ 27°21′43″ 26°32′20″ 27°00′13″

113°08′13″ 112°56′60″ 112°54′13″ 113°05′39″ 112°28′03″ 112°53′16″ 112°26′38″ 113°10′32″ 112°29′28″ 113°08′05″ 112°26′39″ 112°28′27″

66 69 63 68 64 59 70 65 61 70 66 70

6.0 5.3 5.3 6.2 6.1 5.8 6.7 5.6 6.3 6.0 6.5 6.7

± ± ± ± ± ± ± ± ± ± ± ±

0.3 0.1 0.1 0.2 0.3 0.3 0.3 0.1 0.2 0.1 0.2 0.1

18.1 16.1 16.0 18.9 18.8 17.4 15.4 17.1 19.4 18.3 20.1 15.4

SOC (g kg−1) CEC (cmol kg−1) pH Elevation (m)

S1 S2 S3 S4 S5 S6 S7 S8 S9 S10 S11 S12

This study was conducted in the vicinity of the cities of Changsha (E113°04″30′, N28°02″30′), Zhuzhou (E113°04″42′, N27°54″26′) and Hengyang (E112°47″29′, N27°00″20′) in Hunan province, south China. The specific coordinates of the sample locations and site elevations are shown in Table 1, the elevations of the sites ranged from 59 to 70 m. Before the survey, site selection criteria were defined as: located near an abandoned industry/factory at distances of 0.5–2 km, similar land use (agricultural land) with loam-based typical red soil. The general area has a subtropical monsoon climate, with mean annual temperatures ranging from 16 °C to 19 °C and average annual precipitation ranging between 1200 and 1700 mm. The pollution sources in the study

Longitude

2.1. Location and site descriptions

Latitude

2. Materials and methods

Site

Table 1 Mean soil properties and metal concentrations of the selected sampling sites in Hunan province, south China (n = 5, ± SE).

TN (g kg−1)

Pb (mg kg−1)

Cu (mg kg−1)

Zn (mg kg−1)

Cd (mg kg−1)

DTPA-Cd (mg kg−1)

et al., 2004). However, such tests do not accurately reflect the actual type of exposure or the potential toxicity of metals that occurs in the field, where soils are instead slowly contaminated over time through processes involving wet and dry deposition and metal availabilities are affected by long-term partitioning into different mineral phases and compartments (Bolan et al., 2014; Li et al., 2014; Nannoni and Protano, 2016). Information gathered from laboratory studies thus might not comprehensively provide useful information for ecological risk assessment of contaminated soils (Nannoni et al., 2014). Laboratory studies also ignore the adaptability of organisms to long-term environmental contamination, and may overestimate the risks of metals to earthworms where the indigenous communities are comprised of members having different levels of metal tolerance (ISO 11268-2, 1998; OECD 222, 2004; McBride et al., 2009; Smolders et al., 2009; Spurgeon and Hopkin, 1995). To address these concerns, some authors have performed field studies or have exposed earthworms to field-contaminated soils in the laboratory (Hobbelen et al., 2006; Nahmani et al., 2009; Pérès et al., 2011; Lévêque et al., 2013; Nannoni et al., 2014; Beaumelle et al., 2016). In addition, it is recognized that variations in soil chemical factors, and the adaptability and tolerance of different earthworm species also affect earthworm community composition and structure (Pérès et al., 2011; Vijver et al., 2003; Harmsen, 2007; Peijnenburg et al., 2007). In field-contaminated areas, sensitive earthworm species may decrease in abundance or even disappear when they are exposed to heavily polluted soil, and earthworm community structure tends to become simplified or unstable. Despite these apparent differences in metal tolerance, many studies have still used sensitive species for the assessment of environmental risks, ignoring the advantages of tolerant species that might dominate contaminated areas. The latter species also may have the ability to bioaccumulate metals to higher levels, in this way potentially contributing to their food-chain transfer (Vijver et al., 2003; Hobbelen et al., 2006; Harmsen, 2007; Peijnenburg et al., 2007). Therefore, this study was aimed at understanding the variation of earthworm community composition and metal bioaccumulation in dominant species in field-contaminated soils, and addresses differences in earthworm species that are relevant to the use of earthworms for ecological risk assessment and soil remediation. The research described here was carried out in Hunan province, a subtropical area of South China, which has severe metal contamination due to the historical mining and metal smelting activities. We aimed to investigate (i) the earthworm community structure, species composition in different levels of historical contaminated soil; (ii) the impact of soil properties on earthworm community; (iii) the uptake and accumulation of Cd in the dominant earthworm species. The effects of soil contaminants on earthworms in field experiments are influenced by multiple interacting factors. By investigating the earthworm community from various perspectives (species composition, density, diversity and biomass), we hope to achieve a better understanding in the risk of metals on earthworm communities.

0.0 0.1 0.1 0.1 0.2 0.6 0.3 0.2 0.3 0.3 0.5 1.8

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91%, 92% and 90% of the reported certified concentrations for Cd, Cu, Zn and Pb, respectively. Tissue Cd concentrations were only analyzed in the adults of the dominant earthworm species Metaphire californica (individuals having clitella). The earthworms were digested with HNO3-H2O2 in a microwave oven, after which the acid digests were analyzed for Cd concentrations by inductively coupled plasma-mass spectrometry (ICP-MS, 7700ce, Agilent Technologies, Santa Clara, CA, USA). Five replicate soil samples and five earthworms were analyzed per site. GBW08571 (Marine muscle tissue, State Bureau of Technical Supervision, People’s Republic of China) was employed as certified reference material for earthworm analyses. Measured concentrations were within 98.8% of the reported certified concentration for Cd.

areas were mainly related to mining that was carried out from the 1980s to 1990s, and atmospheric deposition that introduced metals by wet and dry deposition from metal smelters. This is a unique scenario to evaluate the impact of in-situ soil contamination and any possible local deposition would be discernible at the study scale. A total of 60 samples were collected from 12 sites (n = 5) that showed variations in soil metal concentrations. The soil samples were processed in five replicates covering a radius of 10 m for each category. Earthworms and soils were sampled at each site in early November 2015. 2.2. Soil properties Soil samples were collected from the upper soil horizon (0–20 cm) in randomly selected plots at each site, using a plastic shovel. Samples were placed into polyethylene bags, labeled and transported to the laboratory for analyses of the soil properties and metal concentrations (Nannoni et al., 2014; Lévêque et al., 2015; Shen et al., 2017; Liu et al., 2017). The soil samples were air dried, ground and sieved to determine soil pH, organic carbon (SOC), total nitrogen (TN), and cation exchange capacity (CEC) (Table 1). Soil pH was measured in a mixture [1:2.5 (w/ v) soil/water] with a pH meter; SOC, TN, and CEC were determined by Walkley and Black (1934), Kjeldahl (Bremner and Mulvaney, 1982), and ammonium acetate exchange (Hendershot and Duquette, 1986) methods, respectively.

2.5. Biological indices For each site, Shannon (1948) and Simpson (1949) indices were calculated based on the density and diversity of earthworm species:

Shannon index: H = − ∑ Pi lnPi Simpson index: D = 1− ∑ (Pi )2 where Pi = percentage of the individuals represented by species i on the total number of individuals. High values of the Shannon index and low values of the Simpson index indicate a high diversity of earthworms. The representation of individuals by taxon (evenness) was evaluated using the Pielou (1969) index:

2.3. Earthworm sampling and preparation Earthworms were sampled and hand separated from five replicate 50 cm × 50 cm × 25 cm plots at each site (Lévêque et al., 2015). Living earthworms were placed on moist filter paper without food for 24 h to egest soil from the gut, and stored in 70% ethanol for further identification and analysis (Nahmani et al., 2007; Lévêque et al., 2015; Carnovale et al., 2015). Ethanol was replaced after 48 h to preserve the specimens in 95% ethanol and the earthworms were examined with a dissecting microscope, and identified to the species level based on anatomical and morphological characteristics with published keys (Xu and Xiao, 2010). Earthworms were washed, patted dry and lyophilized to weigh their dry biomass. Earthworm number and biomass for each site was expressed as the mean density (ind m−2) or biomass (g m−2) (n = 5). The number of adult earthworms (presence of a clitellum) was recorded along with the number of juvenile (non-clitellate) earthworms. The proportion of adults or juveniles is the ratio between adults or juveniles and total number of individuals in each site. The relationship between proportion of adults (%) and soil Cd concentrations was analyzed by linear regression. Normality of data was checked with the Shapiro-Wilk test (p > 0.05) using R3.3.3 with mvnormtest package. The unidentified specimens (e.g. absence of a clitellum or totting flesh), or very small earthworms (less than 0.1 g) were discarded from the analysis.

Pielou index: P = H /ln(total number of taxa) High values of the Pielou index indicate high evenness in the population sizes of different taxa, whereas low values for the Pielou index indicate dominance of the community by one or more species. 2.6. Statistical analysis Statistical analyses were performed using SPSS statistical software package (Version 17.0), and data were plotted with Sigmaplot 10.0. Analyses of variance (ANOVA) were performed and statistical differences between sites were determined using Tukey’s honest significant difference (HSD) test. Co-inertia analysis (CoIA) was also measured to test for significant co-variation among the earthworm community descriptors, soil properties and soil Cd concentrations (Dray et al., 2003). CoIA is a symmetrical approach allowing the use of various methods to model the structure in each data matrix. The analysis computed the covariance matrix crossing the variables. The sum of squared covariance is the total co-inertia. It also computed the eigenvalues and eigenvectors, which described the partitioning of the total co-inertia. Projecting the points and variables of the two original data tables on the co-inertia axes allowed us to compare the projections of the two data tables in the common co-inertia space (Borcard et al., 2011). The RV coefficient, as the multivariate generalization of the Pearson correlation coefficient, is the ratio of the total co-inertia to the square root of the product of the squared total inertias of the separate analyses (Robert and Escoufier, 1976). The CoIA analysis was performed using the R statistical software (version 3.3.1) with the ade4 package (Dolédec and Chessel, 1994; R Development Core Team, 2005; Dray and Dufour, 2007; R Core Team, 2014).

2.4. Metal concentrations in soils and earthworms Cd, Pb, Cu, and Zn concentrations were determined using soil extracts that were prepared by digesting soil samples with aqua regia [HNO3:HCl (v/v) = 1:3] (Bade et al., 2012) in a microwave (MARS-5, CEM Microwave Technology Ltd, Matthews, NC, USA). To assess the available Cd concentrations, the soils from the same samples were extracted with a solution of 5 mmol L−1 DTPA (diethylenetriaminepentaacetic acid) and 10 mmol L−1 CaCl2 (Lindsay and Norvell, 1978; Dai et al., 2004). All digests and extracts were measured by inductively coupled plasma-mass spectrometry (ICP-MS, 7700ce, Agilent Technologies, Santa Clara, CA, USA). GBW07456 (Polluted soil, Institute of geophysical and geochemical exploration by Chinese Academy of Geological Sciences, Inspection and Quarantine of the People’s Republic of China) was employed as certified reference material for soil analyses. Measured concentrations were within 93%,

3. Results 3.1. Soil properties and metal concentrations The soil properties (pH, CEC, TN, and SOC) and metal concentrations at the study sites are shown in Table 1. Descriptive statistics 3

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Pielou index values, and species richness, are shown in Table 3. The Shannon and Pielou values were not significantly different between sites (p > 0.05), and did not correlate with soil Cd concentrations (p > 0.05). In contrast, a significant correlation was observed between the Simpson index and total (R2 = 0.47, F = 8.85, p < 0.05), and available Cd concentrations (R2 = 0.47, F = 8.99, p < 0.05). Correlation strength was site dependent and was highest at S2. The species richness was 4–6 species in soils with Cd concentrations ranging from 0.81 to 2.39 mg kg−1, while there were 3–6 species at Cd concentrations from 3.45 to 5.11 mg kg−1, and 2–3 species at higher Cd concentrations (7.49 and 17.8 mg kg−1 in S11 and S12, respectively) (Table 3). Species richness was not correlated with the Cd gradient in soil (p > 0.05).

showed that Cd concentrations exceeded the threshold legal values at all of the sampling sites [as stipulated in the Chinese Environmental Quality Standard for Soils, prescribed by the Ministry of Environmental Protection of China (GB 15618-1995); Table 1]. In the standard of the soil environmental quality, grade II can be used for protecting human health and food security and for plant growth with the thresholds of Cd 0.3 mg kg−1, Cu 50 mg kg−1, Zn 200 mg kg−1, and Pb 250 mg kg−1, respectively (Bai et al., 2011). Soil total Cd concentrations varied by a factor of 22 with concentrations ranging from 0.81 to 17.8 mg kg−1 dry soil. DTPA-extractable soil Cd concentrations ranged from 0.18 to 9.73 mg kg−1. The total Cd concentrations were 1.83–6.75 times higher than the DTPA-extractable Cd concentrations; total and DTPA-extractable Cd concentrations were significantly and positively correlated (R2 = 0.95, F = 203.8, p < 0.01). Elevated Zn concentrations were observed at S7, S10 and S12, elevated Cu concentrations at S3, S5, S10 and S11 (Table 1). The soil Pb concentrations exceeded the legal threshold only at S11 (Table 1), and ranged from 58.4 to 429 mg kg−1 (7.4-fold variation). Soil organic carbon content ranged from 15.8 to 21.4 g kg−1. Soil pH was usually between 5.3 and 6.7, and CEC varied from 15.4 to 20.1 cmol kg−1. TN contents ranged from 2.1 to 3.2 g kg−1 (Table 1).

3.3. Relationship between earthworm community structure and soil properties Relationships between earthworm community data (the total number of individuals, proportion of adults and juveniles, and diversity and evenness indices) and specific soil properties (CEC, pH, TN, SOC, and soil Cd concentrations) were tested through co-inertia analysis (Fig. 2). Significant co-variation (co-inertia analysis: RV = 0.51, pvalue < 0.05) was observed among earthworm community descriptors and soil data set (Fig. 2). The first axis of the co-inertia analysis (92% variance explained) exhibited the descriptors of earthworms and soils (Fig. 2a). Earthworm community descriptors showed that the total numbers of individuals and adults were positively related to SOC content (R2 = 0.40, F = 6.54, p < 0.05; R2 = 0.40, F = 6.70, p < 0.05, respectively). The number of juveniles was influenced by TN contents (R2 = 0.68, F = 21.7, p < 0.01). In contrast, no significant correlations were found between earthworm diversity and evenness (Shannon and Pielou) and soil properties (TN, SOC, and CEC) or soil Cd concentrations (p > 0.05) (Fig. 2b, c).

3.2. Earthworm community Earthworm species collected at each site are tabulated in Table 2. A total of 2956 individuals were collected from the twelve sites, and comprised 3 families, 6 genera and 14 species that occurred in varying proportions across the different sites. Nine species belonged to the genus Amynthas, 3 species belonged to Metaphire, and the remaining species belonged to the genera Drawida, Bimastus and Aporrectodea. The most widespread and abundant species was Metaphire californica (Kinberg, 1867) (31.7%, of the total number of individuals worms), followed by Amynthas pecteniferus (Michaelsen, 1931) (4.7%), Amynthas heterochaetus (Michaelsen, 1869) (3.9%), and Amynthas morrisi (Beddard, 1892) (3.0%). The species richness was relatively high in S1, S6 and S10 (which contained 6 species), while it was low in S5, S7, S11 and S12 (which had only 3 species). Averaged across all sites, the population density and biomass of M. californica was 16 ind m−2 and 23.7 g m−2, respectively (Table 2). This dominant species was present at all sites including the most highly contaminated soil at S12 (17.8 mg kg−1 Cd). The rare species were Metaphire bucculenta, Metaphire hesperidum, Amynthas obscuritoporus, Aporrectodea trapezoids, and Bimastus parvus, which had relatively low densities (less than 1 ind m−2) (Table 2). Analysis of the distributions of the different earthworm species showed that members of the family Megascolecidae (including the genera Amynthas and Metaphire) were dominant, while the Lumbricidae (including the genera Aporrectodea and Bimastus) were represented by only one species per site or were completely absent. Earthworm communities were less diverse at highly contaminated sites compared to the sites with low levels of Cd (Table 2). The density of earthworms differed significantly among the field sites. Density was relatively high at S1 (80 ind m−2) and at S4 (72 ind m−2) compared to S2 or S3 where the soil Cd concentrations were lower than 3 mg kg−1. One exception occurred where earthworm density was elevated at site S10 (104 ind m−2) (Fig. 1a). Earthworm biomass decreased with increasing Cd concentration, and was the largest at S3 (6.6 g m−2), but biomass values were not significantly different between sites (Fig. 1b). The relationships between the proportion of adults (or juveniles) and soil Cd concentrations showed that the earthworm population age structure was significantly influenced by soil total Cd concentrations (R2 = 0.34, F = 5.2, p < 0.05). The proportion of adults (or juveniles) was not correlated with the DTPA-Cd concentrations (R2 = 0.30, F = 4.34, p > 0.05); the juveniles were mainly found in soils with low Cd concentrations. Earthworm diversity and evenness, expressed as Shannon, Simpson,

3.4. Cadmium accumulation in Metaphire californica Cd concentrations in M. californica were significantly different between sampling sites (Fig. 3), ranging from 17.8 to 259 mg kg−1. The earthworms with highest Cd concentration were collected from the most contaminated soil (Cd > 17.8 mg kg−1 in soil). The relationship between Cd concentrations in M. californica and soil (shown in Fig. 4) was positive and significant (R2 = 0.84, F = 53.36, p < 0.01) (Fig. 4a), while earthworm Cd concentrations also significantly correlated with DTPA-extractable Cd concentration (R2 = 0.90, F = 90.4, p < 0.01, Fig. 4b). 4. Discussion 4.1. Earthworm community characteristics and relationship with soil properties M. californica was the most widespread of the 14 species present, and had the highest percent density (31.7%) across all sites. This implies a broad tolerance to diverse habitats and Cd pollution levels as high as 17.8 mg kg−1 in soil (S12). Earthworm community composition tended to be simplified, which was reflected by a decrease in species richness, especially at sites S11 and S12. Some species (such as A. homochaetus and A. hupeiensis) decreased or were absent in soil having high levels of Cd as compared to these with low levels of metal contamination. Along with the decrease in species diversity, the total density of earthworms seemed to be related to soil Cd concentrations (Fig. 1). One exception was S10, where a relatively high average earthworm density (80 inds m−2) was found in soil having 5.11 mg Cd kg−1. This can probably be explained by the high SOC content of this soil, which positively influences earthworm density. The SOC content was the highest 4

5

Metaphire californica (Kinberg, 1867) Metaphire bucculenta (Gates, 1935) Metaphire hesperidum (Beddard, 1892) Amynthas homochaetus (Chen, 1938) Amynthas hupeiensis (Michaelsen, 1895) Amynthas heterochaetus (Michaelsen, 1869) Amynthas pecteniferus (Michaelsen, 1931) Amynthas morrisi (Beddard, 1892) Amynthas obscuritoporus (Chen, 1930) Amynthas zhongi (Wang, 1991) Amynthas bellatulus (Gates, 1932)

Aporrectodea trapezoids (Dugès, 1828) Bimastus parvus (Eisen, 1874)

Drawida japonica (Michaelsen, 1892)

Megascolecidae

Lumbricidae

Moniligastridae

1.6 ± 1.6 1.6 ± 1.6 2.4 ± 1.6

1.6 ± 1

0.8 ± 0.8

0.8 ± 0.8

4 ± 3.1

2.4 ± 1.6

12 ± 5.1

S2

20.8 ± 5.3

S1

0.8 ± 0.8

0.8 ± 0.8

2.4 ± 1.6

17.6 ± 3.2

S3

0.8 ± 0.8

15.2 ± 7

4±4

0.8 ± 0.8

12 ± 2.5

S4

1.6 ± 1.6

4 ± 2.5

3.2 ± 1.5

S5

4±4

4.8 ± 2.9

2.4 ± 2.4

2.4 ± 2.4

0.8 ± 0.8

15.2 ± 7

S6

3.2 ± 3.2

0.8 ± 0.8

8.8 ± 2

S7

1.6 ± 1.6

1.6 ± 1

1.6 ± 1

1.6 ± 1.6

15.2 ± 5.6

S8

1.6 ± 1.6

5.6 ± 1

4 ± 3.1

8 ± 4.6

S9

0.8 ± 0.8

15.2 ± 5.1

2.4 ± 1.6

0.8 ± 0.8

3.2 ± 3.2

36 ± 14

S10

4 ± 1.8

0.8 ± 0.8

12 ± 3.1

S11

Note: Density (Inds m−2) and Biomass (g m−2) represent the average for individuals across all sampling sites and the average biomass of each earthworm species per site, respectively.

Species

Family

Table 2 Earthworm species densities (individuals m−2) and biomass (g m−2) (n = 5, ± SE) in the sampled sites in Hunan province, south China.

0.8 ± 0.8

4 ± 2.5

26.4 ± 5.5

S12

0.52 ± 0.35

0.08 ± 0.07

0.08 ± 0.07

0.28 ± 0.27

0.32 ± 0.33

0.12 ± 0.13

1.48 ± 1.26

2.32 ± 1.27

1.92 ± 0.56

0.72 ± 0.23

1.4 ± 0.41

0.12 ± 0.13

0.08 ± 0.07

15.6 ± 2.55

Density (Inds m−2)

0.85 ± 0.04

0.01 ± 0.0

0.42 ± 0.03

2.46 ± 0.20

0.45 ± 0.04

0.70 ± 0.06

3.78 ± 0.19

5.55 ± 0.19

8.49 ± 0.20

11.4 ± 0.77

3.75 ± 0.11

2.93 ± 0.24

1.79 ± 0.15

23.7 ± 0.47

Biomass (g m−2)

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Fig. 1. (a) Population density (individuals m−2) and (b) biomass (g m−2) of earthworms in Cd contaminated soils from different sites in Hunan province, south China, including unidentified species and juveniles (n = 5). Different letters indicate significant differences in earthworm counts and biomass between sites (p < 0.05).

(21.0 g kg−1) in S10, providing a more favorable habitat to earthworms. Results of the present study are in good agreement with previous studies that have reported that the densities and biomass of earthworms decrease in response to metals, and that sensitive species are absent from highly contaminated soils, leading to reduced species richness (Spurgeon and Hopkin, 1999; Kamitani and Kaneko, 2007; Lévêque et al., 2015). Spurgeon and Hopkin (1999) reported that the abundance and biomass of earthworms was decreased with increasing Pb, Zn, and Cd concentrations in the proximity of a smelting factory. In contrast, Kamitani and Kaneko (2007) showed that soil contamination might not influence the abundance and biomass of earthworms, although in that study the highest soil metal concentrations were 2.51 mg Cd kg−1, 386.7 mg Cu kg−1, 260.6 mg Zn kg−1, and 95.1 mg Pb kg−1, which approaches those measured at S4 in our study. Lévêque et al. (2015) studied the influence of metals on earthworm communities near a lead recycling factory at which soil Pb concentrations ranged from 5000 to 30,000 mg kg−1. Apporectodea longa was the main species present just near the smelter. Earthworm density and species richness were significantly reduced, with the density decreasing progressively from 135 inds m−2 to 5 inds m−2 as soil Pb concentrations increased from 480 to 5060 mg kg−1. In the present study, we observed that the proportion of adults was positively correlated with total soil Cd concentrations (p < 0.05), indicating that adult earthworms may be more tolerant to Cd than juveniles. These findings are in agreement with Lévêque et al. (2015), who found a greater proportion of adults in highly polluted soils containing Pb concentrations of 30,000–5000 mg kg−1 along a metal pollution gradient close to a lead-recycling factory. Žaltauskaitė and Sodienė (2014) exposed juveniles of Eisenia fetida to Cd (1–500 µg Cd g−1) and Pb (20–2500 µg Pb g−1) for fourteen weeks. Both metals significantly affected the survival of the juveniles (fourteen-week LC50 296 ± 125 µg Cd g−1 and 911 ± 164 µg Pb g−1) and prolonged the time to sexual maturation. The survival probability for the juveniles decreased significantly with time and soil metal concentration, and the juveniles were more sensitive to metals than adults. Moreover, the proportion of adults (or juveniles) was not associated with DTPA-extractable Cd, suggesting that it is not the available Cd but rather the total concentration that predicts metal uptake by the earthworms (Hobbelen et al., 2006). Earthworm community diversity and evenness characterized by Shannon and Pielou indices were not significantly different among the sites in this study (p > 0.05, Table 3), while the Simpson index values decreased significantly with increasing soil Cd concentration (R2 = 0.47, p < 0.05). This suggests that Cd might be the priority contaminant influencing the density of dominant earthworm species in our study areas. This disparity between different measures of diversity has been observed in previous studies. The Simpson index is considered

to be a dominance index as the formula gives more weight to dominant species as compared to the Shannon index. Grześ (2009) also reported that the Simpson index decreased, indicating a decrease of species dominance with increasing soil contamination. While the main focus of this study was on the effects of Cd, other environmental factors should also be taken into account, such as the properties of the soil that affect habitat quality and the bioavailability of metals to earthworms. Here, the study sites were mainly agricultural soils with typical loam-based red soil. Time of year is also a factor; here soils were sampled during the dry season in early November, when soil water contents ranged from 10% to 20% and were suitable for earthworm activity, whereas at other times of the year, earthworm abundance may vary with soil moisture, temperature, and carbon inputs from plant growth. In this study, the correlation between the earthworm parameters (density, adult or juvenile numbers, diversity) and soil Pb (7.4-fold variances) or Cu (11-fold variances) concentrations were not significant (p > 0.05), indicating that earthworm species composition might not impacted by Pb or Cu in the study areas. Based on these results, earthworm parameters appeared to be only related to Cd concentrations (22-fold variances). Descriptive approaches based on CoIA (co-inertia test) provide further information on the relationships between earthworm density, diversity index, species evenness with selected soil properties and Cd concentrations (Fig. 2). Among the factors examined, the density of earthworms was significantly and positively correlated with SOC, a finding that has been observed in some other studies (Wang et al., 2009; Yatso and Lilleskov, 2016; Lee, 1985; Nannoni et al., 2014). Likewise, earthworm density is also known to be correlated with TN contents, similar results were reported by Umiker et al. (2009). These soil factors can strongly affect the population response to metal contamination. Kamitani and Kaneko (2007) found that the earthworm community was influenced to a greater extent by soil properties (pH and clay fraction) rather than the level of soil contamination by metals. In agreement with their results, the present study showed that the size of the earthworm populations was strongly correlated with SOC levels, which ranged from 15.8 to 21.4 g kg−1. This highlights the importance of considering differences in SOC when evaluating the potential effects of Cd toxicity on earthworm communities in soils that vary in organic matter content.

4.2. Cadmium concentrations in Metaphire californica M. californica belongs to epi-endogeic ecological category of earthworms. Members of this category feed on a mixture of soil and surface litter, and are mainly found in humid tropical soils (Fragoso et al., 1999). According to our results, M. californica showed a high capacity for bioaccumulation of Cd. This is in agreement with the findings of other authors in studies that have compared different earthworm 6

0.49 ± 0.05a 0.93 ± 0.02a 0.62 ± 0.16a 4

0.59 ± 0.09a 0.90 ± 0.02ab 0.53 ± 0.14a 6

H D P S

0.52 ± 0.05a 0.89 ± 0.02ab 0.65 ± 0.17a 4

S3 0.63 ± 0.09a 0.89 ± 0.03ab 0.60 ± 0.16a 5

S4 0.32 ± 0.10a 0.72 ± 0.18ab 0.18 ± 0.18a 3

S5 0.55 ± 0.17a 0.56 ± 0.13ab 0.45 ± 0.20a 6

S6 0.39 ± 0.08a 0.58 ± 0.11ab 0.31 ± 0.20a 3

S7

S10 0.65 ± 0.09a 0.71 ± 0.08ab 0.56 ± 0.15a 6

S9 0.65 ± 0.08a 0.86 ± 0.05ab 0.84 ± 0.04a 3

S8 0.59 ± 0.11a 0.78 ± 0.11ab 0.58 ± 0.15a 5

0.51 ± 0.11a 0.86 ± 0.08ab 0.48 ± 0.19a 2

S11

0.39 ± 0.15a 0.45 ± 0.14b 0.34 ± 0.21a 3

S12

Note: S1 to S12 represent the soil sampling sites and H, D, P and S correspond to the values of Shannon index (H′), Simpson index (D′), Pielou index (P′), species richness (S) in each site, respectively. The same letter indicate significantly differences at p < 0.05, using a post-hoc Tukey HSD test.

S2

S1

Index

Table 3 Mean values of Shannon index (H′), Simpson index (D′), Pielou index (P′), and species richness (S) for the earthworm community at each site in Hunan province, south China (n = 5, ± SE).

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d=1

S3 S8

S2

D

7

S5

S9

JU

TE

(b)

S7

S6

S4

S1 S11 S12

S10

(a)

= 0.2 d = d0.2

TN = 0.2 d =d0.2

P H CEC

SOC Cd pH

(c)

AD

Fig. 2. Results of co-inertia analysis between earthworm community descriptors and soil properties for different Cd contaminated sites in Hunan province, south China, set in a plane defined by axis 1 and 2 (RV = 0.51, p-value < 0.05 = 0.007, RV is a multivariate generalization of the Pearson correlation coefficient). (a) Projection of 12 sites. (b) Projection of earthworm descriptors. H, D, P correspond to the values of Shannon index (H′), Simpson index (D′), Pielou index (P′) in each site. TE, AD and JU correspond to abundance of earthworm individuals, adults and juveniles. (c) Projection of soil properties parameters.

Fig. 3. Cd concentrations in Metaphire california from different sites in Hunan province, south China. Different letters indicate significant differences in Cd concentrations among sampling sites.

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study, M californica, is a member of the Pheretima group that is widely distributed in China (Blakemore, 2002; Blakemore, 2004; Easton, 1979, 1982; James, 2005a,b; Sims and Easton, 1972; Chang et al., 2008), which is mainly associated with the soils that are rich in decaying organic matter. The high values of BAFCd for M. californica as compared to those for other metals may be attributed mainly to the greater mobility of Cd in soil (Nannoni et al., 2011; Wang et al., 2018). In the present study, the BAFCd had an average value of 11, and showed considerable variability, ranging from 5.07 to 24.8 times over the levels of total soil Cd. Its ability to bioaccumulate Cd suggests that this species has particular mechanisms for storage and detoxification of Cd that should be further studied. Cd concentrations in the earthworms were significantly correlated with both total and DTPA-extractable soil Cd concentrations. Similar results in previous work showed that Cd concentrations in earthworms depend not only on readily available Cd, but also on the total soil Cd concentration (Beaumelle et al., 2016). It is noteworthy that the accumulation of Cd in earthworms might also linked to DTPA-Cd levels as the latter are also significantly correlated with total Cd contents (p < 0.01) (Li and Thornton, 2001; Dai et al., 2004; Nannoni et al., 2011). Among metals of concern, Cd has a relatively high mobility and overall bioavailability in soil compared to other elements such as Pb, which is more strongly bound to organic matter and clay particles in soil (Chlopecka et al., 1996; Kabala and Singh, 2001; Nannoni et al., 2011; Nannoni et al., 2014; Acosta et al., 2015; Rodr Guez-Seijo et al., 2017). The bioaccumulation of Cd in the abundant earthworm species M. californica suggest that this species may be particularly useful for risk assessment and applications that use earthworms to mix amendments into the soil (Hobbelen et al., 2006; Nahmani et al., 2007). 5. Conclusions Fourteen species of earthworm were identified in Cd-contaminated fields in Hunan Province, a subtropical area of Southern China. Among the species that were found, M. californica was the most widespread across the study area and was the dominant species occurring even in the most contaminated soil (Cd: 17.8 mg kg−1). In agreement with prior studies, the earthworm community structure tended to simplify with increasing levels of soil Cd contamination, which was manifested both by decreased earthworm density and species diversity (Simpson index). Further, Cd contamination altered the age structure of the earthworm community and was associated with increases in the proportion of adults to juveniles. The dominant species, M. californica, was found to bioaccumulate high concentrations of Cd in its tissues, which reached concentrations that were 5.07 to 24.8 times those measured for total soil Cd. The ability to bioaccumulate Cd and survive at high population numbers indicates this species is relatively tolerant to Cd contamination. Cd concentrations in the earthworms were correlated with both the total and extractable soil Cd concentrations. This study highlights the importance of considering differences in the adaptability of different species to metal contamination when earthworms are used for risk assessment of Cd-polluted soils.

Fig. 4. Relationship between Cd concentrations of Metaphire californica and total (a) or DTPA-Cd concentration (b) (p < 0.01) in soils from Hunan province, South China.

species (Dai et al., 2004; Hobbelen et al., 2006). Dai et al. (2004) studied the concentrations of Cd in earthworms (A. caliginosa and Lumbricus rubellus) compared with the total and DTPA-extractable soil concentrations near a former metallurgic industrial site. In that study, total soil concentrations ranged from 2.7 to 5.2 mg Cd kg−1. Concentrations of Cd were higher in A. caliginosa than in L. rubellus, with values ranging from 11.6 to 102.9 mg kg−1 in A. caliginosa, and from 7.7 to 26.3 mg kg−1 in L. rubellus, respectively. Hobbelen et al. (2006) reported that variations in the Cd concentrations in the tissues of L. rubellus (48.1–160 mg kg−1) were best predicted by total soil Cd concentrations, whereas the Cd concentrations in the body tissues of A. caliginosa (61–256 mg kg−1) were most strongly related to pore water Cd concentrations. These findings imply that in floodplain soils, earthworm Cd concentrations can be elevated even at low levels of available Cd. This further suggest that some uptake also occurs from the less available Cd pool that is included in measurements of total soil Cd. Prior studies have examined possible differences in the tolerances of different earthworm species to metals (e.g. L. rubellus, A. caliginosa, Nicodrilus caliginosus) (Dai et al., 2004; Nannoni et al., 2014), some of which tolerate high metal concentrations through use of mechanisms involving sequestration (Lanno et al., 2004; Andre et al., 2009), excretion and detoxification (Morgan and Morgan, 1998; Morgan and Morgan, 1999; Suthar et al., 2008). The ecological roles of earthworm species may differ in relation to the pattern of metal bioaccumulation from soil fractions, which are impacted by complex factors especially in field contaminated soils. The dominant species identified in the present

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