1.3 Composition and biomass of earthworms

Four different types of organic waste sludge used in this study were as follows: municipal sewage sludge (MSS) collected from sewage treatment plants on Gwacheon (Gyeonggi Province, South Korea); industrial sewage sludge (ISS) collected from industrial complex on Ansan (Gyeonggi Province); alcohol fermentation processing sludge (AFPS) collected from Ansan industrial complex; and leather processing sludge (LPS) collected from sewage treatment plant on Cheongju (Chungbuk Province, South Korea). Pig manure compost (PMC) was purchased from Anjung Nong-hyup, Anjung (Gyeonggi Province). These materials were collected in early March 1994 and kept in deep freezers (-60°C) to be applied annually from 1994 to 2001.

Lysimeters which composed of 45 concrete plots (1.0 m length, 1.0 m width and 1.1 m depth) (Fig. 2) were made in the upland field of Suwon (Gyeonggi Province) in March 1993. Each plot was uniformly filled with the same sandy loam soil without earthworms up to the ground surface in mid-May 1993. Three levels (12.5, 25 and 50 tons of dry matter ha-1 year-1) of test materials were applied to each plot twice annually for 8 consecutive years (mid­March 1994 to mid-March 2001) and mixed into the soil of a depth of 15 cm. PMC served as a standard for comparison in lysimeter tests. A randomized complete block design with three replicates was used. Two radish, Raphanus sativus, cultivars (jinmialtari and backkyoung) were cultivated in every spring and autumn, respectively. Planting densities were 12 x 15 cm in spring and 25 x 30 cm in autumn with one plant. Other practices followed standard Raphanus culture methods without application of any mineral fertilizer and pesticide. The lysimeters were covered with a nylon net to prevent any access by birds or animals.

Fig. 2. Field lysimeters

Earthworms were collected from each of the 45 lysimeter plots from an area of 1 m2 up to 0.3 m depth by hand sorting in mid-October 1997 and mid-October 2001 as described previously (Callaham & Hendrix, 1997). They were immediately transported to the laboratory in plastic containers and separated into juveniles and adults with a clitellum. The earthworm numbers, composition and biomass were investigated before they were fixed in a 10% formalin solution. Earthworm species identification followed Hong & James (2001), Kobayashi (1941) and Song & Paik (1969).

Pollution index (PI) was determined according to the method of Jung et al. (2005), PI = [^(heavy metal concentration in soil tolerable level-1) number of heavy metal-1]. Tolerable level of Cu, Zn, Cr, Cd, Pb and Ni were 125, 700, 10, 4, 300 and 100 mg kg-1 in Korean soil, respectively (Anon., 2007). PI values are employed to assess metal pollution in soil and indicate the average on ratios of metal concentration over tolerable level. A soil sample is judged as contaminated by heavy metal when PI value is greater than 1. Total toxic unit of PTEs was calculated by threshold level described under the Soil Environmental Conservation Act (Anon., 2007) in South Korea as follows: £ (Cu 50 + Zn 300 + Cr 4 + Cd 115 + Pb 100 + Ni 40). Bonferroni multiple-comparison method was used to test for significant differences among treatments in the fresh biomass of earthworms and pollution indices (SAS Institute, 2004). Correlations between accumulated pollutant contents and observed earthworm numbers and biomass were estimated from the Pearson correlation coefficients using SAS. pH values, heavy-metal contents and pollution indices of 8 consecutive yearly applications of three levels of four different organic waste materials and PMC in field lysimeters were reported previously (Na et al., 2011).

Effects on earthworm composition of 8 consecutive yearly applications of four organic waste materials and PMC were investigated using field lysimeters (Table 2). Earthworm composition in all treatments varied according to waste material examined, treatment level and application duration. Of 390 adults collected from 45 plots, earthworms were classified into 2 families (Megascolecidae and Moniligastridae), 2 genera (Amynthas and Drawida) and 5 species (Amynthas agrestis, Amynthas hupeiensis, Amynthas sangyeoli, Drawida koreana and Drawida japonica). The number of earthworm species in MSS-, ISS-, LPS-, AFPS — and PMC-treated soils was 2, 2, 2, 3 and 5, respectively. The dominant species were A. agrestis, A. hupeiensis, A. sangyeoli and D. japonica in the sludge treatments 4 years after treatment but was replaced with A. hupeiensis in all the plots 8 years after treatment. This finding indicates that A. hupeiensis was more tolerant to toxic heavy metals than other earthworm species. In ISS — and LPS-treated soils, the proportion of juveniles appeared was 67-100% 4 years after treatment, but no juveniles was observed 8 years after treatment.

At 4 years after treatment, effect of test waste material (F = 16.91; df = 4,44; P < 0.0001) and treatment level (F = 4.09; df = 2,44; P = 0.0268) on the number of earthworms was significant (Table 2). The material by level interaction was also significant (F = 2.63; df = 8,44; P = 0.0258). At 8 years after treatment, effect of test waste material (F = 17.33; df = 4,15; P < 0.001) and treatment level (F = 11.00; df = 3,29; P < 0.001) on the number of earthworms was significant. The material by level interaction was also significant (F = 20.53; df = 8,44; P < 0.001). The number of earthworms was significantly reduced in 25 and 50 ton MSS treatments, 25 and 50 ton AFPS treatments and 12.5 and 25 ton PMC treatments 4 years after treatments than 8 years of treatments. The total number of earthworms collected 4 and 8 years after treatment was as follows: MSS-treated soil, 66/29; ISS-treated soil, 4/2; LPS — treated soil, 15/1; AFPS-treated soil, 30/11; and PMC-treated soil, 127/439.

Earthworm biomass collected from 45 plots during the 8-year-investigation period is given in Fig. 3. The biomass in all treatments was dependent upon waste material examined, treatment level and application duration. At 4 years after treatment, effect of test waste material (F = 49.45; df = 4,44; P < 0.0001) and treatment level (F = 5.80; df = 2,44; P = 0.0074) on the earthworm biomass was significant. The material by level interaction was also significant (F = 3.88; df = 8,44; P = 0.0031). At 8 years after treatment, effect of test waste material (F = 165.13; df = 4,44; P < 0.0001) and treatment level (F = 14.39; df = 2,44; P < 0.0001) on the earthworm biomass was significant. The material by level interaction was also significant (F = 19.77; df = 8,44; P < 0.0001). Significant increase in biomass of soil treated with 50 ton PMC ha-1 year-1 was observed 8 years after treatment.

Table 2. Earthworm numbers and composition of 4 and 8 consecutive yearly applications (twice annually) of three levels of four different organic waste materials and pig manure compost using field lysimeters

To evaluate potential toxic effects of residual heavy metals, total toxic units of PTEs were determined (Fig. 4). The total toxic units in all treatments varied with waste material examined, treatment level and application duration. At 4 years after treatment, effect of test waste material (F = 34872.4; df = 4,44; P < 0.0001) and treatment level (F = 60.24; df = 2,44; P < 0.0001) on the the total toxic units of PTEs was significant. The material by level interaction was also significant (F = 2601.2; df = 8,44; P < 0.0001). At 8 years after treatment,
effect of test waste material (F = 52439.5; df = 4,44; P < 0.0001) and treatment level (F = 28451.0; df = 2,44; P < 0.0001) on the the total toxic unit of PTEs was significant. The material by level interaction was also significant (F = 13057.2; df = 8,44; P < 0.0001).

Fig. 4. Total toxic units of potentially toxic elements (PTEs) of 4 (■) and 8 ( ) consecutive yearly applications (twice annually) of three levels of four different organic waste materials and pig manure compost using field lysimeters. Abbreviations are same as in the text

Fig. 5. Pollution indices of 4 (■) and 8 ( ) consecutive yearly applications (twice annually) of three levels of four different organic waste materials and pig manure compost using field lysimeters. Abbreviations are same as in the text

PI values of lysimeter soils sampled during the 8-year-investigation period are reported in Fig. 5. At 4 years after treatment, effect of test waste material (F = 34047.6; df = 4,44; P < 0.0001) and treatment level (F = 5957.3; df = 2,44; P < 0.0001) on the the total toxic unit of PTEs was significant. The material by level interaction was also significant (F = 2505.3; df = 8,44; P < 0.0001). At 8 years after treatment, effect of test waste material (F = 48793.6; df = 4,44; P < 0.0001) and treatment level (F = 26515.1; df = 2,44; P < 0.0001) on the the total toxic unit of PTEs was significant. The material by level interaction was also significant (F = 12190.9; df = 8,44; P < 0.0001). There was significant difference in PI values between the treatment duration. Particularly, PI value of ISS-treated soil was higher 8 years after treatment than 4 years after treatment, while PI value of LPS-treated soil was higher 4 years after treatment than 8 years after treatment.

Correlation between total toxic unit of PTEs and PI and earthworm individuals and biomass was determined (Table 3). At 4 years after treatment, earthworm individuals were correlated negatively with the total toxic unit of PTEs (r = -0.509) and PI (r = -0.508). At 8 years after treatment, earthworm individuals were correlated negatively with the total toxic unit of PTEs (r = -0.265), but were not correlated negatively with PI.

At 4 years after treatment, earthworm biomass was correlated negatively with the total toxic unit of PTEs (r = -0.673) and PI (r = -0.672) (Table 3). At 8 years after treatment, earthworm biomass was correlated negatively with the total toxic unit of PTEs (r = -0.308), but were not correlated negatively with PI.

	Correlation coefficient (r)
Parameter	Earthworm individuals	Earthworm biomass
	4 YAT	8 YAT	4 YAT	8 YAT
Total toxic unit of PTEs	-0.509	-0.265	-0.673*	-0.308
PI	-0.508*b	-0.265	-0.672*	-0.280

a Years after treatmen b *0.001<P<0.05 treatment Table 3. Correlation between total toxic unit of potentially toxic elements (PTEs) and pollution indicies (PI) and earthworm individuals and biomass 4 and 8 years after treatment

The impact of heavy metals and sludge on lumbricid earthworms, particularly E. fetida and L. terrestris, has been well noted. Heavy metals cause mortality and reduce fertility, cocoon production and viability, growth, composition and biomass, and bioaccumulation and bioavailability of earthworms. The toxic values of heavy metals to earthworms vary according to an earthworm acute toxicity test. Based upon an artificial soil test, Spurgeon et al. (1994) determined no observed-effect concentrations (NOECs) for E. fetida exposed to heavy metals. The estimated NOEC values were 39.2 mg Cd kg-1, 32 mg Cu kg-1, 1,810 mg Pb kg-1 and 199 mg Zn kg-1. In soil contaminated by effluent containing Cr, the rate of 10 mg kg-1 was fatal to Peretima posthuma and other species (Abbasi & Soni, 1983). Copper caused higher mortality than Pb or Zn against E. fetida at the same rate and the LC50 and NOEC values for Cd could not be determined since no significant mortality was observed at the highest test rate (300 pg g-1) (Spurgeon et al., 1994).

Although heavy metals did not show direct lethal effects to earthworms, they can sensitively cause their reproduction and sperm count reduction and low hatching success of cocoons. Lumbricus terrestris worms exposed in artificial soil to sublethal concentrations of technical chlordane (6.25, 12.5 and 25 ppm) and cadmium nitrate (100, 200 and 300 ppm) exhibited significant reduction in spermatozoa from testes and seminal vesicles (Cikutovic et al.,

1993) . Eisenia fetida worms grew well in the lead-contaminated environment and produced cocoons at the same rate as the control worms, but the hatchability of these cocoons was much lower, indicating that lead toxicity affects reproductive performance by major spermatozoa damage (Reinecke & Reinecke, 1996). In addition, Zn, Mn and Cu produced slower growth, later maturation and fewer or no cocoons. Reinecke and Reinecke (1997) have shown the structural damage of spermatozoa, including breakage and loss of nuclear and flagellar membranes, thickening of membranes, malformed acrosomes and loss of nuclear material, and the results are associated with heavy metals, such as Pb and Mn. The toxicity order of metals on reproduction in earthworms is Cd, Cu, Zn and Pb. Similar results have been found in E. fetida exposed to a geometric series of concentrations of Cd, Cu, Pb and Zn in artificial soil and the effects of Cd and Cu on the reproductive rate were particularly acute (Spurgeon et al., 1994).

It has been well known that earthworms are able to inhabit soils contaminated with heavy metals (Becquer et al., 2005; Li et al., 2010; Maity et al., 2008) and can accumulate undesirably high concentration of heavy metals (Cu, Zn, Pb and Cd) that may give adverse effects on livestock (Hobbelen et al., 2006; Oste et al., 2001). Earthworms (L. rubellus and Dendrodrilus rubidus) sampled from one uncontaminated and 15 metal-contaminated sites showed significant positive correlations between earthworm and total (conc. nitric acid — extractable) soil Cd, Cu, Pb and Zn concentrations (Morgan & Morgan, 1988). The important factor in the accumulation of heavy metals in earthworms is bioavailability by uptake (Dai et al., 2004; Spurgeon & Hopkin, 1996) because there are significant correlations between the concentrations of heavy metal accumulated in earthworms and bioavailable metal concentrations of field soils (Hobbelen et al., 2006). Earthworm metal bioaccumulation and bioavailability have been well reviewed by Nahmani et al. (2007). There were positive relationships between earthworm tissue and soil metal concentrations and also earthworm tissue and soil solution metal concentrations with slightly more significant relationships between earthworm tissue and soil metal concentrations 42 days after treatment. Recently, Li et al. (2010) reported the positive logarithmic relationship between the bioaccumulation factors of E. fetida to heavy metals and the exchangeable metal concentration of pig manure. The differences in these accumulation and availability among earthworms may, in part, play a role in affecting their population density and genetic adaptation living in metal — contaminated soils.

However, Lee (1985) suggested that the differences in the relative toxicity of compounds may explain some of the conflicting data in the literature on the concentrations which have deleterious effects on earthworms. For instance, very high concentrations of lead that influence growth and reproduction of earthworms may be attributable more to the very low solubility of lead compounds that are found in soils and the ability of earthworms to sequester absorbed lead than to any lower toxicity of lead compared with other heavy metals. It has been suggested that E. fetida may regulate the concentration of zinc in their body tissue through allowing rapid elimination by binding zinc using metallothioneins in their chloragogenous tissue (Cotter-Howells et al., 2005; Morgan & Morris, 1982; Morgan & Winters, 1982; Prento, 1979). High tolerance of earthworms to cadmium poisoning may also result from detoxification by metallothionein proteins in the posterior alimentary canal (Morgan et al., 1989). In addition, heavy metals have high affinity for glutathione, metallothioneines and enzymes of intermediary metabolism and heme synthesis (Montgomery et al., 1980). The metals Zn, Pb, Bi and Cd which are not consistently prevailing toxicants were most accessible to earthworms and Cu, Zn and Cr were also accumulated in earthworm tissue and the contaminated soils imparied earthworm reproduction and reduced adult growth, while elevated superoxide dismutase activity suggested that earthworms experienced oxidative stress (Berthelot et al., 2008).

Lead, copper and zinc may inhibit d-aminolevulinic acid dehydratase (d-ALAD) which is a key enzyme in heme synthesis by lowering haemoglobin concentration in earthworm blood. Replacement of zinc, a protector of the active site of d-ALAD, by lead may result in its inhibition.

Soil pH has been comprehensively identified as the single most important soil factor controlling the availability of heavy metals in sludge-treated soils (Alloway & Jackson,

1991) . Soil pH is also one of the most important factors that limit the species, numbers and distribution of earthworms (Dunger, 1989; Edwards & Bohlen, 1996; Satchell & Stone, 1972) because it may affect the survival of adults and thus production and avoidance behaviour of juveniles (Aorim et al., 1999, 2005). van Gestel et al. (2011) reported that soil pH and organic matter content determine molybdenum toxicity to enchytraeid worm, Enchytraeus crypticus A higher pH resulted in a decreased sorption of the molybdate anion, and it caused increased bioavailability and toxicity.

A lot of studies concerning the effects of heavy metals on earthworms in terms of mortality, loss of weight, fertility, cocoon production, cocoon viability and growth were carried out during short-term experiments (14 or 21 days) in artificial soils contaminated with metal solution containing a single metallic element. Recently, Na et al. (2011) studied the effects of long-term (8 years) application of four organic waste materials on earthworm numbers and biomass. They reported that earthworm individuals were correlated positively with pH (r = 0.37) and negatively with heavy metals (r = -0.36 to -0.55) with the exception of Zn 4 years after treatment, while earthworm individuals were correlated positively with pH (r = 0.46) and negatively with Pb (r = -0.41) but positively with Zn (r = 0.59) 8 years after treatment. Earthworm biomass was correlated negatively with heavy metals (r = -0.43 to -0.72) with the exception of Zn 4 years after treatment, while earthworm biomass was correlated positively with pH (r = 0.57) and negatively with Pb (r = -0.50) and Ni (r = -0.30) but positively with Zn (r = 0.68) 8 years after treatment.

1.4 Effects of hexane extractable material on composition and biomass of earthworm

United States Environmental Protection Agency [USEPA] 9071B method (1998) was used to extract relatively non-volatile hydrocarbons from 45 lysimeter soils treated twice annually with three levels of four different organic waste materials and pig manure compost tested for 8 consecutive years, as stated in section 4.1. The extracts were generally designated hexane extractable material (HEM) because the solvent used was hexane. Soils were acidified with 0.3 ml of concentrated HCl and dried over magnesium sulfate monohydrate. After drying in a fume hood, HEM was extracted for 4 hr using a Soxhlet apparatus which was attached a 125 ml boiling flask containing 90 ml of hexane. Solvent was then concentrated under vacuum for less than 30 min at 35°C. The extracts were cooled in a desiccator for 30 min, and HEM concentrations were calculated by the formula, HEM (mg kg of dry weight-1) = (A x 1000)/BC, where A is gain in weight of flask (mg), B is weight of wet solid (g) and C is dry weight fraction (g of dry sample g of sample-1).

HEM amounts varied with treatment level and organic waste examined (Fig. 6). At 8 years after treatment, effect of test waste material (F = 49.45; df = 4,14; P < 0.001) and treatment level (F = 4.09; df = 2,30; P = 0.028) on the HEM was significant. The material by level interaction was also significant (F = 2.63; df = 8,44; P = 0.0258). Particularly, the amount of HEM in PMC-treated soil was the lowest of any of test materils at all treatment levels.

Fig. 6. Hexane extractable material (HEM) contents of 8 consecutive yearly applications (twice annually) of three levels of four different organic waste materials and pig manure compost using field lysimeters. Abbreviations are same as in the text

Correlation between HEM content (Fig. 6) and earthworm individuals and biomass (Table 2) was determined. At 8 years after treatment, earthworm individuals were negatively correlated with HEM (r = -0.313) and earthworm biomass (r = -0.335).

In general, organic compounds existed in sewage sludge have been potentially transferred to sludge-amended agricultural soils, and most organic compounds have been solved in hexane solvent. HEMs from sewage sludges contain a variety of contaminants, such as hydrocarbons, grease, plant or animal oils, wax, soap, polychlorinated biphenyls (PCBs) and polycyclic aromatic hydrocarbons (PAHs) (Hua et al., 2008; Stevens et al., 2003). Drescher — Kaden et al. (1992) reported that 332 organic contaminants (e. g., pyrene, benzo(fl)pyrene, benzene and toluene) with potential to exert soil contamination were identified in German sewage sludges. Hembrock-Heger (1992) found that the concentrations of PAHs and PCBs appeared to be highest in soils treated with sewage sludge for 10 years. According to the United Kingdom Water Research Centre Report No. DoE 3625/1 on the occurrence, fate and behaviour of some of organic pollutants in sewage sludge (Sweetman et al., 1994), there was no evidence of any significant problems arising from organic contaminants in sludges applied to agricultural land.

Of some waste sludge and PMC applied into red pepper fields in South Korea from 2003 to 2004, the highest contents of HEM and PAHs were observed in cosmetic and pharmaceutical industy sludge, respectively, and the cosmetic industry sludge affected remarkedly growth of red pepper, which resulted in 25-60% of yield reduction (Lee, 2006). These results indicate that PMC may contain a lot of polar compounds with functional groups, such as COO-, O-, NR2H, COOH or OH, to be more easily metabolized by various soil-born organisms, including earthworm. Water drained from processing of ISS, LPS, MSS and AFPS may contain more non-soluble compounds than that of PMC. Considering the hexane fraction obtained from PMC containing plentiful P or N atom (Na, 2004), it may be biodegradable by long-term exposure to a variety of soil organisms owing to biological uses. In general, most hydrophobic compounds are accumulative and difficult to biodegrade them introducing into environments because most aliphatic hydrocarbons retain unfavorable large AG (minus value) with increase in chain length.