The accumulation of metals (Cd, Cu, Pb, Zn and Ca) by two
ecologically contrasting earthworm species (
Lumbricus rubellus
and
Aporrectodea caliginosa
): implications for ecotoxicological testing
J.E. Morgan
a, A.J. Morgan
b,*aDepartment of Applied Science, Writtle College, Chelmsford Essex, CM1 3RR, UK
bSchool of Pure and Applied Biology, University of Wales College of Cardiff, P.O. Box 915, Cardiff CF1 3TL, UK Received 16 July 1998; received in revised form 12 February 1999; accepted 16 February 1999
Abstract
The metal (Cd, Cu, Pb, Zn and Ca) concentrations in the tissues, ingesta (crop contents) and egesta (faeces) were investigated in two physiologically contrasting earthworm species (Lumbricus rubellus and Aporrectodea caliginosa) inhabiting soils exhibiting various levels of heavy metal contamination. In addition, a complementary soil layering experiment, conducted under laboratory conditions, was undertaken to investigate whether the distribution of Pb within a soil vertical pro®le in¯uenced the relative metal accumulation patterns of these species. Generally, the Cd, Cu and Pb concentrations of ®eld populations ofA. caliginosawere signi®cantly greater than inL. rubellus, a pattern reversed for Ca. Concentrations of Zn were signi®cantly greater inA. caliginosain soils containing the lowest Zn concentrations, but no species differences were apparent at high soil concentrations of this metal. Comparisons of metal concentrations between ingesta and soils indicate that both species selectively ingest material from the soil matrix, although no signi®cant correlations were found between tissue metal and ingesta metal concentrations. Differences in concentrations of Cd, Pb and Zn between the ingesta of the species were, however, concomitant with observed differences in tissue concentrations of the respective metals, which cannot be explained by excretion via the egesta. The soil strati®cation experiment indicated that Pb distribution within a soil pro®le affected the pattern of species differences in tissue metal concentrations observed in ®eld populations. The evidence therefore suggests that the difference in dietary intakes of these metals is an important factor in contributing to observed differences between these species, although other factors are also contributory. The observations are discussed in the context of soil hazard assessment monitoring, and in particular, the role of concentration factors in such applied surveys.# 1999 Elsevier Science B.V. All rights reserved.
Keywords:Earthworms; Heavy metals; Species differences; Concentration factors; Ecotoxicology
1. Introduction
Different earthworm species inhabiting the same polluted microhabitat clearly exhibit different
dispo-sitions to accumulate essential and non-essential metals (Ireland, 1979; Ireland and Richards, 1977; Ash and Lee, 1980; Wright and Stringer, 1980; Ander-sen and LaurAnder-sen, 1982; Morris and Morgan, 1986; Morgan and Morgan, 1988, 1991, 1993; Morgan and Morris, 1982; Morgan et al., 1986; Terhivuo et al., 1994). It has been suggested that these observations *Corresponding author. Tel.: 24200; fax:
+44-1245-420456; e-mail: jem@writtle.ac.uk
are a result of inter-speci®c differences in dietary metal intakes and physiological utilisation (Andersen and Laursen, 1982; Ireland, 1983; Beyer et al., 1985; Morris and Morgan, 1986; Hopkin, 1989; Morgan and Morgan, 1992, 1993), although little direct evidence has been collected to substantiate these claims.
In addition, although there appears to be a degree of inter-species consistency in the accumulation of par-ticular metals from metalliferous mine soils (e.g. Morgan and Morgan, 1991), such patterns become obscured when soils polluted primarily by aerial deposition are considered (e.g. Ash and Lee, 1980). Such disparities appear to have received little atten-tion, but clearly have implications for biomonitoring and hazard assessment programmes.
The aims of this study were three-fold. First, to con®rm our provisional observations, restricted to one metalliferous site (Morgan and Morgan, 1993), that differences exist in tissue Cd, Cu, Pb, Zn and Ca concentrations between two ecophysiologically con-trasting earthworm species (Lumbricus rubellus and Aporrectodea caliginosa).L. rubellusis a litter dwell-ing (epigeic) species, which under normal conditions lives outside of the mineral substrata, whereas A. caliginosalives permanently in essentially horizontal burrows within the organo-mineral soil matrix, and is thus categorised as an endogeic species (see Sims and Gerard, 1985).
Second, to establish whether tissue metal concen-trations in L. rubellusand A. caliginosa re¯ect the respective metal concentrations of their ingested materials, since it appears that absorption of metals through the alimentary canal probably accounts for most of the metal burdens of earthworms (e.g. Piearce, 1972). Although a number of studies (e.g. Ma, 1982; Ma et al., 1983; Morgan and Morgan, 1988) have demonstrated statistically that, for a number of heavy metals, the concentration in the soil explains a large proportion of the variance in accumulated tissue metal concentrations of earthworms, there appears to be little published information on the metal concentra-tions of material ingested by earthworms. This is surprising since earthworms are known to ingest material selectively (Piearce, 1978; Edwards and Bohlen, 1996).
Third, to investigate, by means of a simple soil layering experiment, whether the differences in Pb accumulation by L. rubellus and A. caliginosa is
affected by the preferred vertical distribution of the species. The source of contamination may in¯uence metal distribution within the soil vertical pro®le (Martin and Coughtrey, 1982), which could be a factor determining the exposure and accumulation by biota, and the consequent metal transfer through food chains.
Recent laboratory exposure studies have suggested that species differences exist in their sensitivities to metals (Spurgeon and Hopkin, 1995) and organic residues (Edwards and Coulson, 1992). Under ®eld conditions the effects of metal contaminants are more complicated: the toxicological effects of a given metal on a given species should be coupled with the prob-ability of encountering the metal during the earth-worm's normal activities within the soil vertical pro®le (i.e. exposure). Clearly this aspect of bioavail-ability of toxicants to de®ned target species has important implications for toxicity testing pro-grammes under ®eld conditions.
2. Materials and methods
2.1. Earthworms and soils
Mature (clitellate) specimens of L. rubellus and A. caliginosa were collected from an abandoned, non-acidic Pb±Zn mine at Llantrisant, S. Wales (O.S. grid reference ST 048822) by formalin extrac-tion (0.55%, 20 l). Samples were taken from ®ve 5 m2 `stations' across the site and represented areas within the mine workings and adjacent pastures, which were used for cattle grazing. Both species were collected in a similar manner from a pasture adjacent to an aban-doned Pb±Zn mine at Halkyn, N. Wales (O.S. grid reference SJ 201704), and from an uncontaminated `control' site at Dinas Powis, S. Wales (O.S. grid reference ST 146723). All sampling occurred early in November.
further day without ®lter paper but with a few drops of deionised water to allow complete egestion of gut contents. Dissection of several specimens con®rmed that the guts were clean of consumed material.
The animals were quenched in liquid nitrogen, placed in individual pre-weighed glass beakers, and oven-dried (24 h at 708C) to constant weight. The samples were allowed to cool in desiccators and re-weighed in a moisture-free environment on a Mettler HK60 semi-microbalance. The samples were wet oxidised in 2 ml boiling concentrated `AnalaR' grade nitric acid (BDH Chemicals, Poole, Dorset, UK), and were made up to 10 ml with double de-ionised water. Blank digests were also prepared.
At least 15 soil samples from two depths (0±5 and 10±15 cm) were taken randomly from within the designated areas and pooled for each sampling area. Each pooled soil was dried at 258C in a fan-oven, gently crushed in an acid-washed porcelain pestle and mortar, and passed through a 2 mm2aperture stainless steel sieve. Subsequent analysis was undertaken on ®ve sub-samples of each soil.
`Total' soil metal concentrations were determined for both depths by extraction with boiling concen-trated `AnalaR' grade nitric acid (Morgan and Mor-gan, 1988). Extracted metal concentrations (0±5 cm) were determined after equilibration of 3 g of soil with 35 ml 0.5% acetic acid for 3 h (Morgan, 1987). Soil pH (de-ionised water) was measured in soil slurries (w/v, 1/2) following equilibration for 1 h (Peech, 1965). Soil organic carbon was determined by the Walkley±Black method (Allison, 1965).
2.2. Ingesta (crop contents) and egesta (faeces)
Fresh specimens of A. caliginosa collected from four Llantrisant stations were dissected and the con-tents of the crop and gizzard of each animal removed and placed on individual pre-weighed Millipore (Millipore S.A., Moshein, France) ®lter paper pieces. The contents of the crop/gizzard of two animals were combined for each sample. For comparative purposes, the crop/gizzard contents were removed fromL. rubel-lusat Station 5 of Llantrisant. Limited sample num-bers precluded the crop contents being collected from L. rubellusat all sites.
Egesta were collected from A. caliginosa taken from Llantrisant stations during the ®rst day of
starva-tion. For comparative purposes egesta were collected from L. rubellus from Station 5. Each sample was derived by pooling faeces from two individuals for each species, and prepared for analysis as described for the ingesta.
Samples were oven-dried at 708C for 24 h, cooled in a desiccator, and weighed on a Mettler HK60 semi-microbalance. Each sample was individually wet-oxi-dised in concentrated boiling `AnalaR' grade nitric acid. Excess acid was evaporated to leave approxi-mately 0.5 ml to which was added 1 ml double de-ionised water. Each sample was centrifuged at 2500 rpm for 2 min, and the supernatant removed. A further 1 ml aliquot of de-ionised water was added to the solid residue, the suspension again centrifuged, and the supernatant added to that collected previously. This procedure was repeated a third time. The ®nal volume of the pooled supernatant solution was made up to 5 ml with double de-ionised water. Blank digests were also prepared.
2.3. Soil stratification experiment
This laboratory experiment was designed to deter-mine whether the vertical distribution of Pb within strati®ed combinations of naturally contaminated and uncontaminated soils affects the relative concentra-tions of Pb accumulated by L. rubellusand A. cali-ginosa. Groups of between 12 and 16 similarly sized, clitellate animals of each species, collected from the uncontaminated Dinas Powis site, were exposed to different soil treatments for 60 days in the dark at 48C. Prior to their exposure to the appropriate soil, each worm group was starved for 2 days to remove most of their gut contents.
Surface soils (0±5 cm) used for the experiment were collected from Dinas Powis (uncontaminated site) and Llantrisant (Pb-contaminated site). The individual soils were thoroughly mixed before being placed in plastic containers to a depth of 8 cm, as follows:
Expt. Group 1 8 cm Dinas Powis soil Expt. Group 2 8 cm Llantrisant soil
Expt. Group 3 5 cm Llantrisant soil covered by 3 cm Dinas Powis soil
At the end of the exposure period, the animals were removed from the soils and processed as previously described. Soil samples were taken for analysis from 0±3 and 3±8 cm horizons in each experimental con-tainer. The soils were processed for `total' soil Pb as described above.
2.4. Atomic absorption spectrophotometry (AAS)
Digests were quanti®ed for Cd, Cu, Pb, Zn and Ca by ¯ame (air-acetylene) AAS in a Varian AA6 atomic absorption spectrophotometer. Samples from the soil layering experiment were analysed for Pb only. Back-ground correction for Cd, Cu, Pb and Zn analysis was made automatically with a hydrogen continuum lamp. Samples and standards for determination of Ca were made up containing 1% lanthanum.
2.5. Statistics
Statistical comparisons between the species for dry weight and metal concentrations of tissue, ingesta and egesta were made by the Mann±Whitney U-test. Sig-ni®cance between metal concentrations in ingesta and the whole worm was determined by Kendall's rank correlation. For all statistical tests, differences were regarded as signi®cant ifp< 0.05.
3. Results
3.1. Soils
Soil pH, percent organic carbon content and `total' (nitric acid-extractable) Cd, Cu, Pb, Zn and Ca con-centrations (0±5 cm depth) are presented in Table 1. Total concentrations of Cd, Pb and Zn were elevated in the vicinity of the abandoned mine sites compared to the uncontaminated control site (Dinas Powis), but the concentrations of Cu were broadly similar at all sites (Table 1). `Total' concentrations of Ca ranged from 1050 to 8300 mg/kg dw, and re¯ected the range of soil pH (5±6.8). `Total' concentrations of metals in the 10± 15 cm depth (data not shown) ranged from 0.8 to 16 mg/kg dw for Cd, 21±60 mg/kg dw for Cu, 158± 10 020 mg/kg dw for Pb, 185±1870 mg/kg dw for Zn and 980±8120 mg/kg dw for Ca. The data indicated that no surface (0±5 cm) enrichment of heavy metals
was prevalent in any of the soils investigated. The acetic acid-extractable fraction of the heavy metals (Table 1) was greatest for Cd (9±26%), followed by Zn (2±10%), Cu (0.7±1.4%) and Pb (0.4±1.0%). Extractability for Ca ranged from 11 to 41%. Organic carbon contents of the soils (0±5 cm) ranged from 3 to 9% (Table 1).
3.2. Whole-worm field observations
Concentrations of Cd, Cu, Pb, Zn and Ca in L. rubellusandA. caliginosafrom the different sites are presented in Table 2. The concentrations of Cd, Pb, and Zn were greater in both species inhabiting the mine soils than in the uncontaminated soil.
Mean concentrations of Cd in A. caliginosa were between 2 and 5 times greater than those ofL. rubellus in six of the seven soils examined, with all the differences statistically signi®cant. The trend for Pb was similar to that for Cd; greater concentrations were found inA. caliginosa, although the differences were not always statistically signi®cant. No consistent pat-tern for Zn accumulation was apparent for these species, although A. caliginosa accumulated signi®-cantly greater concentrations of Zn than L. rubellus from the soils containing the lowest (total and extrac-table) concentrations of Zn. Although concentrations of Cu were broadly similar for both species (range 11± 20 mg/kg dw) in all soils, the concentrations in A. caliginosa were signi®cantly greater in four of the soils studied. Concentrations of Ca were signi®-cantly greater in L. rubellusin all soils.
3.3. Metal concentrations in ingesta (crop contents) and egesta (faeces)
The Cd, Cu and Ca concentrations in the ingesta of both species were in general about 2±15 times greater than total concentrations in bulk soil (compare Tables 1 and 3). The pattern for Pb and Zn was less consistent, although marked differences between ingesta and soil were apparent for A. caliginosa. The results suggest that both species ingest material selectively.
the egesta. No signi®cant correlations (tested by Kendall's rank correlation) were found between the concentrations of metals in crop contents and earth-worm tissue.
The concentrations of Cd, Pb and Zn in the crop contents of A. caliginosa from Llantrisant 5 were signi®cantly greater than those of L. rubellus (Fig. 1, Table 3), and re¯ected the signi®cant species differences in the whole-worm concentrations of the respective metals. No clear pattern was evident for the Cu concentrations in the ingesta and whole-worm, and no signi®cant species differences were found. Concentrations of Ca were greater in the crop contents of L. rubellus, although the difference was not statistically signi®cant; signi®cant differences were apparent between the species in the whole-worm Ca concentration.
No signi®cant differences between the species were apparent in the concentrations of Cd, Cu and Pb of the egesta (Fig. 1). The concentration of Zn in the egesta ofA. caliginosawas signi®cantly greater than inL. rubellus, with the pattern being reversed for Ca.
3.4. Soil stratification experiment
The concentrations of Pb in A. caliginosa were signi®cantly greater than inL. rubellusfor Treatments 1 to 3 (Table 4) following the pattern found in the ®eld studies. However, in Treatment 4, where the surface soil was greatly contaminated with Pb, the concentra-tions in both species were similar. The results indicate that the pattern of Pb accumulation by the two earth-worm species is affected by the distribution of soil Pb within the vertical soil pro®le.
Table 1
Total (16N nitric acid-extractable) and extractable (0.5% acetic acid-extractable) concentrations of Cd, Cu, Pb, Zn and Ca (mg/kg dry wt.), percentage organic matter and pH of soils (0±5 cm depth) from an uncontaminated site (Dinas Powis) and metalliferous mine sites
Total metal concentration (mg/kg dry wt.)
Sitea Cd Cu Pb Zn Ca
DP 0.90.1 261 1667 1934 179048
Halk 3.60.1 403 135017 6759 608071
Ll1 14.70.3 621 10110280 155027 7560212
Ll2 6.50.1 271 237063 77021 5320110
Ll3 10.50.1 301 382038 83015 8300166
Ll4 170.5 311 673071 196033 512086
Ll5 2.70.1 231 5706 4606 105021
Extractable metal concentration (mg/kg dry wt.)
DP 0.080.01 0.40.1 10.2 4.50.1 68017
Halk 0.320.01 0.40.1 80.1 3812 229043
Ll1 3.00.1 0.30.1 1004 1554 173020
Ll2 1.20.1 0.40.1 181 31 161015
Ll3 1.60.1 0.30.1 211 351 192025
Ll4 2.40.1 0.20.1 291 972 211525
Ll5 0.70.1 0.20.1 31 241 3752
Organic carbon (%) pH (de-ionised water)
DP 4.70.3 5.190.01
Halk 9.00.5 6.520.03
Ll1 5.60.2 6.540.01
Ll2 4.10.1 6.490.01
Ll3 5.00.2 6.780.01
Ll4 6.50.3 6.720.01
Ll5 2.90.1 4.980.01
Values are meanSE;n5.
4. Discussion
Earthworms are known to be selective consumers (Edwards and Bohlen, 1996). Although relatively few studies have measured the concentrations of contami-nants within material ingested by earthworms, pub-lished ®ndings con®rm selective consumption because there are marked differences between residue concen-trations in bulk soil and ingesta, both in the case of trace organics (Diercxsens et al., 1985) and heavy metals (Ireland, 1976; Morgan and Morgan, 1992). The present study also demonstrated that there are signi®cant differences in concentrations of certain essential and non-essential metals between the ingesta and bulk soil.
Such observations pose questions in relation to the interpretation of concentration factors (the ratio of metal concentration of earthworm tissue and bulk soil), which have been cited as being representative
indicators of metal bioavailability in soils (Neuhauser et al., 1995). Although there exist clearly demon-strable relationships between the concentrations in earthworms and bulk soil for several metals (Morgan and Morgan, 1988; Neuhauser et al., 1995), the parti-tion coef®cients of a metal within bulk soil have been shown to be an important parameter in determining metal bioavailability (Janssen et al., 1997). An advan-tage of the concentration factor approach is that it integrates the bioavailable metal fraction and the kinetics of the transport system (Janssen et al., 1997) and thus may have practical application for the evaluation of the potential toxicological impact of metals in soils. However, the data from the present study would suggest that without careful considera-tion, the use of such indices could lead to spurious conclusions in relation to toxicological testing of xenobiotics on different earthworm species. Further-more, it is probable that earthworm exposure to resi-Table 2
Dry weights (mg) and metal concentrations (mg/kg dw) inA. caliginosa(Ac) andL. rubellus(Lr) from an uncontaminated soil (Dinas Powis) and 6 soils from abandoned metalliferous mines (see Table 1 for key) (meanSE;nnumber of samples analysed). Statistical differences were determined by the Mann±Whitney U-test
Site Species Dry wt. (mg) Cd Cu Pb Zn Ca
DP Ac (n21) 94.63.6 242 170.6 81 66032 161032
Lr (n11) 186.912.8 121 141.1 51 33025 3225225
**** **** NS NS **** ****
Halk Ac (n12) 86.44.3 908 170.6 2010306 1835169 3420302 Lr (n11) 105.28.5 425 140.4 23831 1495235 5853452
NS **** *** **** NS ****
Ll1 Ac (n9) 85.88.5 23349 201.2 1940320 1160174 2270168 Lr (n8) 128.816.7 12312 141.2 46037 137587 4815227
NS ** ** **** NS ****
Ll2 Ac (n17) 94.36.3 13511 141 1270208 1200129 225083 Lr (n5) 120.418.2 14431 101 33065 112078 4740374
NS NS * ** NS ****
Ll3 Ac (n8) 71.24.5 22945 121.1 310109 770127 300050 Lr (n8) 129.312.5 576 131.2 30044 1350214 8300661
**** **** NS NS NS ****
Ll4 Ac (n16) 88.06.1 34430 130.8 1460267 1505131 2000199 Lr (n13) 117.111.7 14716 110.5 900115 1745198 4810309
NS **** * NS NS ****
Ll5 Ac (n17) 70.61.6 21411 120.4 66249 131553 326073 Lr (n27) 100.54.3 463 161.5 716 77059 52801690
**** **** NS **** **** ****
NSNot significant. *p< 0.02.
dues via the dietary route is more accurately repre-sented by considering residue concentrations in ingested crop contents rather than those in bulk soil. The use of concentration factors needs to be developed to account for such species differences; ecophysiolo-gical factors pertaining to earthworms (and indeed other invertebrate groups) must form an integral part of any toxicity testing regime.
The present study extended and con®rmed earlier observations (Morgan and Morgan, 1993; Morgan et al., 1986) on the relative accumulation of metals by the two earthworm species inhabiting a metallifer-ous mine soil: in general, signi®cantly higher concen-trations of Cd and Pb were found inA. caliginosathan inL. rubellus. Similar differences between endogeics and epigeics have also been found from uncontami-nated soils (Beyer et al., 1987; Morgan, 1987), although no differences were apparent in a study of species representing these groups in Finnish soils (Terhivuo et al., 1994). Marino et al. (1992) found that concentrations of Cd and Pb were generally greater in epigeic species than in endogeics inhabiting soils contaminated by vehicular emissions, although a con¯icting trend was apparent for earthworms inha-biting a soil heavily contaminated with Pb from a smelter (Terhivuo et al., 1994). Clearly the
relation-ships between earthworms and aerially deposited metals are complicated (Kansanen and Venetvaara, 1991), but at least some of the apparent contradictions in published observations may in part be explained by differences in the vertical distribution of metals within the soil pro®le. Contamination of soil by aerial deposition of heavy metals not only results in high, but long-term persistently high (Davies et al., 1988; McGrath and Lane, 1989) concentrations in surface layers (Martin and Coughtrey, 1982). Thus epigeic earthworms preferentially feeding within the surface soil layers may become exposed under such pollu-tion events to greater concentrapollu-tions of metals than endogeics occupying deeper layers. This suggestion gains a measure of support from the laboratory ex-periment undertaken in the present study. Our ®ndings are also in agreement with the conclusions of Beyer (1992) on trace organics; he concluded that the distribution of residues within the vertical soil pro®le, coupled with the depth within soil in which most activity of an earthworm species occurs, are important factors to be considered in ecotoxicological studies.
The observed pattern for tissue Zn accumulation appears to be dependent on the soil concentration of the metal. At lower soil concentrations, the body Table 3
Dry weight (mg) and concentrations of metals (mg/kg dw) in ingesta and egesta ofA. caliginosa(Ac) andL. rubellus(Lr) from Llantrisant soils (meanSE;nnos of samples). Statistical differences in metal concentrations were determined by the Mann±Whitney U-test
Species Site Component Dry wt. (mg) Cd Cu Pb Zn Ca
Ac Ll1 Ingesta (n10) 4.50.6 334 696 4520306 100547 10 4001190 Egesta (n6) 43.73.1 110.7 392 3720301 110095 3430720
** ** NS NS **
Ac Ll2 Ingesta (n10) 4.90.4 354 414 2100117 64542 7340311 Egesta (n10) 26.23.1 90.8 311 3425255 99065 2500294
*** * ** *** ***
Ac Ll3 Ingesta (n9) 5.70.3 479 352 197077 56039 8250750 Egesta (n9) 48.95.3 131.1 272 3400238 92086 6940610
*** * *** ** NS
Ac Ll5 Ingesta (n10) 5.60.8 315 527 74051 55015 13 5601932 Egesta (n9) 24.33.2 60.4 241 94074 68521 1760132
** *** ** *** ***
Lr Ll5 Ingesta (n10) 6.40.7 91 373 57041 41522 11 9501883 Egesta (n15) 40.06.0 61 274 81035 60519 3320409
** * *** *** ***
NSnot significant. *p< 0.05.
burden inA. caliginosawas signi®cantly greater than inL. rubellus. However, no species differences were evident at elevated soil Zn concentrations, a pattern which has previously been reported for these same earthworm species (Morgan and Morgan, 1993; Mor-gan et al., 1986). The greater ef®ciency of Zn accu-mulation by A. caliginosa in uncontaminated soils may be related to a requirement of a labile pool of Zn to ful®l the physiological demands of diapause, a resting phase which L. rubellus does not undergo (Edwards and Bohlen, 1996). Zinc has been shown to be mobilized from the chloragosomal storage com-partment during diapause of Aporrectodea species (Morgan, 1984; Morgan and Morgan, 1993; Morgan and Winters, 1991), although the physiological sig-ni®cance of diapausal Zn ¯ux remains to be eluci-dated. The marked differences in the whole-worm concentrations of Ca between the species primarily
relates to Ca burdens within the calciferous glands (Morgan and Morgan, 1990). The calciferous glands ofL. rubellusare `active' and contain mineralised Ca; A. caliginosacontains no such mineralized products within its so-called `inactive' calciferous glands (Piearce, 1972; Morgan, 1982).
The ®ndings demonstrate that differences between the concentrations of Cd, Pb and Zn in the ingesta of these two earthworm species are accompanied by concomitant differences in the tissue burdens of the respective metals. This evidence suggests that the concentrations of these metals within the ingesta are important determinants of accumulated metal burdens. The ®ndings support the conclusion of Mor-gan and MorMor-gan (1992) that species differences in whole-metal concentrations largely re¯ect differences in food selectivity and niche separation. Exposure of L. rubellus and A. caliginosa to Cd-spiked arti®cial Fig. 1. Comparisons between the mean metal concentrations (mg/kg dw) of whole-worm, ingesta and egesta ofA. caliginosaandL. rubellus
substrates (Ireland and Richards, 1981), and the ®nd-ings for Pb from the soil strati®cation experiment in the present study, add further support to this important conclusion. However, it is recognised that metal accu-mulators must not only consume metal-contaminated materials, but that the metals must also be rendered available for transport across the absorptive epithelia (Morgan et al., 1993). `Exposure' and `bioavailability' are related but not interchangeable terms.
The concentrations of Cd, Cu and Pb in the egesta were similar for the two species, which reinforces the fact that the degree of contamination of ingested materials by these three metals is a determinant of tissue concentrations. In contrast, inter-species differences were found for faecal Zn and Ca; greater concentrations of Zn and Ca were found in the egesta of A. caliginosa than in the egesta of L. rubellus. These differences were concomitant with higher concentrations of these metals in tissues and ingesta ofA. caliginosa compared withL. rubellus.
The apparent anomaly of greater concentrations of Pb and Zn in the egesta than in the ingesta is a
counter-intuitive observation which cannot easily be inter-preted, although it is a phenomenon which has been documented for other earthworm species inhabiting metalliferrous soils (Ireland, 1976; Morgan and Mor-gan, 1992). The increase in metal concentration in egesta has been suggested to be a consequence of a marked reduction in the proportion of organic material within the ingested material as it passes along the alimentary canal, resulting in apparent metal enrich-ment (Morgan and Morgan, 1992).
The ®ndings of this study indicate that there is no simple overall relationship between concentrations of metals within ingested material and whole-worm metal burdens. However, the demonstrable ability of earthworms to selectively consume material from bulk soil is a factor which warrants further investigation in pollution studies, particularly since (arti®cial) soils are widely employed in toxicity studies to determine risk assessment parameters (Greig-Smith et al., 1992; Spurgeon and Hopkin, 1995; Spurgeon et al., 1994; Khalil et al., 1996a, b). Furthermore, species-speci®c burrowing and feeding behaviour patterns may render Table 4
Comparisons between concentrations of Pb (mg/kg dw) inA. caliginosaandL. rubellusmaintained on experimental soils under laboratory conditions for 60 days
Treatment Species, Soil depth No. Dry wt. (mg) Pb (mg/kg dw)
One A. caliginosa 13 74.14.0 233
L. rubellus 12 133.36.2 132
*
Soil (0±3 cm) 1535
Soil (3±8 cm) 1546
Two A. caliginosa 13 72.14.9 41648
L. rubellus 16 124.48.0 19727 **
Soil (0±3 cm) 6457107
Soil (3±8 cm) 648398
Three A. caliginosa 11 69.45.2 29238
L. rubellus 15 123.55.5 11116 *
Soil (0±3 cm) 1849
Soil (3±8 cm) 6519113
Four A. caliginosa 13 76.03.8 1810
L. rubellus 16 129.59.4 7511
NS
Soil (0±3 cm) 6497105
Soil (3±8 cm) 20315
Treatment: `One'8 cm DP soil; `Two'8 cm Ll soil; `Three'top 3 cms layer of DP soil5 cm Ll soil; `Four'top 3 cm Ll soil5 cm DP soil.
some earthworm species more exposed or susceptible to contaminants (Tomlin, 1992; Edwards and Coulson, 1992; Greig-Smith, 1992) under ®eld (Spurgeon and Hopkin, 1995) and laboratory (Edwards and Coulson, 1992) conditions. These aspects need to be addressed especially if legislative guidelines are to be an objec-tive of such testing programmes (Marinussen, 1997). The choice of species for ecotoxicological testing needs careful consideration, and must not only take into account the source of pollution (with consequent distribution within the soil pro®le), but identi®cation of the earthworm species likely to represent the most vulnerable receptor for the contaminant. Additional factors which are also worthy of consideration include the bioavailability of contaminants within consumed food material (Streit, 1984; Greville and Morgan, 1989), assimilation ef®ciencies of different species (Wieser et al., 1977; Joose et al., 1981; Russell et al., 1981; Janssen et al., 1991; Brown and Bell, 1995), and biomarkers of toxicant exposures and effects (Weeks, 1998). Thus orthodox toxicity testing alone may not be adequate to estimate the risk to earthworms of eco-toxicants (Tomlin, 1992).
Acknowledgements
We wish to thank Mrs Valerie Ives and Mrs Tessa Smith for typing the Tables, and Miss Elizabeth Morgan for producing the Figure. Part of this work was undertaken during the tenure of an award from the British Natural Environment Research Council.
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