Functional and community-level soil microbial responses to zinc addition may depend on test system biocomplexity

Chemosphere 65 (2006) 1747–1754 www.elsevier.com/locate/chemosphere

Functional and community-level soil microbial responses to zinc addition may depend on test system biocomplexity Line E. Sverdrup a,b,*, Roar Linjordet a, Gjermund Strømman c, Snorre B. Hagen d, ˚ sa Frostega˚rd f, Roald Sørheim a Cornelis A.M. van Gestel e, A a Bioforsk–Norwegian Centre for Soil and Environmental Research, NO-1432 A˚s, Norway University of Oslo, Department of Biology, P.O. Box 1050, Blindern, NO-0316 Oslo, Norway c Bioforsk Lab., NO-1432 A˚s, Norway d University of Tromsø, Institute of Biology, NO-9037 Tromsø, Norway e Institute of Ecological Science, Vrije Universiteit, De Boelelaan 1085, 1081 HV Amsterdam, The Netherlands Norwegian University of Life Sciences, Department of Chemistry, Biotechnology and Food Science, P.O. Box 5003, NO-1432 A˚s, Norway b

f

Received 3 October 2005; received in revised form 21 April 2006; accepted 27 April 2006 Available online 13 June 2006

Abstract The effect of zinc on soil nitrification and composition of the microbial community in soil was investigated using a full factorial experiment with five zinc concentrations and four levels of biological complexity (microbes only, microbes and earthworms (Eisenia fetida), microbes and Italian ryegrass (Lolium multiflorum var. Macho), and microbes, ryegrass and earthworms). After 6 weeks of exposure, the activity of soil nitrifying bacteria was measured and the microbial community structure was characterized by phospholipid fatty acid (PLFA) analysis. Soil nitrification and several PLFA markers were significantly influenced by either zinc addition and/or the presence of earthworms or ryegrass, and one of the most pronounced changes was the increase of fungi and decrease of bacteria with increasing concentrations of zinc. Of particular interest, however, was the potential interaction between the presence of plants and/or earthworms and the effect of zinc, which the factorial study design allowed us to explore. Such an effect was observed in two cases: Earthworms reduced the positive effect of zinc on the fungal biomass (ANOVA, p = 0.03), and the effect of earthworms on the soil nitrification activity depended on zinc concentration (ANOVA, p < 0.05). The effect of earthworm presence was not very large, but it does show that multispecies tests might give information about metal toxicity or bioavailability that cannot be predicted from single-species tests.  2006 Elsevier Ltd. All rights reserved. Keywords: Zn; Terrestrial; Multispecies; Model ecosystem

1. Introduction The bioavailability of zinc in soil varies predictably with soil-specific factors, such as pH (Spurgeon and Hopkin, 1996; Van Beelen and Fleuren-Kemila¨, 1997; Smit and van Gestel, 1998; Lock et al., 2000), organic matter content (Spurgeon and Hopkin, 1996; Van Beelen and FleurenKemila¨, 1997; Smit and van Gestel, 1998), and/or cationic * Corresponding author. Present address: University of Oslo, Department of Biology, P.O. Box 1050, Blindern, NO-0316 Oslo, Norway. Tel.: +47 22 85 46 47; fax: +47 22 85 47 26. E-mail address: [email protected] (L.E. Sverdrup).

0045-6535/$ - see front matter  2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2006.04.075

exchange capacity (Smit and van Gestel, 1998; Lock et al., 2000), and this is reflected in models that are currently suggested for assessing the toxicity of zinc to terrestrial organisms exposed in single-species test systems (see e.g., Lock and Janssen, 2001a). However, recent studies show that aged zinc-contaminated soils are less toxic than predicted from these models (Smit and van Gestel, 1998; Lock and Janssen, 2001b), and that zinc toxicity to terrestrial organisms may be affected by the presence of other metals in the soil or soil solution (Van Gestel and Hensbergen, 1997; Sharma et al., 1999) and previous exposure to zinc or other metals (McLaughlin and Smolders, 2001; Salminen et al., 2001). Thus, even when considering a single species in a

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controlled laboratory setting, assessing metal bioavailability and toxicity is a complex issue. In nature, one rarely has complete knowledge of all the factors that are known to affect toxicity of metals in the laboratory. Furthermore, little is known about the possible confounding effect of other natural stressors not present under optimal laboratory conditions (e.g., temperature, drought, competition, predation, parasitism, diseases, or presence of other pollutants). Ecological risk assessment involves not only an extrapolation from laboratory data to field toxicity, taking all such potentially confounding factors into account, but also aims at predicting effects on the entire ecosystem based on short-term toxicity data from a limited number of test species. Most studies on toxicity (microbial bioassays excluded) have been carried out in single-species test systems, thus leaving largely unknown the possible changes in the bioavailability and toxicity of zinc in the presence of other organisms. However, recent studies show that earthworms can affect local zinc concentrations by relocating metals from hot spots in soils (Zorn et al., 2005), changing the soil pH (Yu et al., 2005), and influencing on metal extractability and bioavailability (Cheng and Wong, 2002; Ma et al., 2003; Zorn et al., 2005). It is also known that plant root uptake of nutrient metals may exceed supply by mass flow so that depletion of metals such as Zn (McLaughlin, 2003) may occur at root surfaces, which may potentially reduce metal exposure for organisms living in the rhizosphere. These studies suggest that the toxicity of zinc to one organism could be affected by the presence of other organisms. There is increasing evidence that some microorganisms are more sensitive to heavy metal stress than plants and animals inhabiting polluted soils (Giller et al., 1998). Microorganisms have a key role in nutrient cycling in soils, and their activity and functional diversity is therefore of prime importance. In the present study, we investigated the microbial responses to zinc additions both at the functional level (nitrification) and with respect to community composition as determined by phospholipids fatty acid (PLFA) patterns. Phospholipids are important components of all cell membranes, and since different groups of microorganisms have different PLFA compositions, the PLFA pattern of a soil sample will reflect the composition of the soil microbial community. Membrane lipids are also affected by environmental factors so physiological effects may be detected even if the community composition remains unaltered. Many previous studies have dealt with effects of zinc on soil microbial parameters, but none has focused on the potential influence of higher organisms in such experiments. We therefore designed the present study to investigate the effect of zinc and biocomplexity, alone and in combination, using a factorial experiment employing five concentration levels of zinc and four levels of biological complexity. From this, we aimed to determine: (i) effects of zinc on soil nitrification in the presence or absence of earthworms, ryegrass, and both, and (ii) effects of zinc on soil microbial community structure (PLFA pattern) in the

presence or absence of earthworms, ryegrass, and both. The zinc test concentrations were carefully selected within a range that would elicit a response in several microbial parameters, but not influence the activity of the higher organisms (earthworms and plants). This way, there was no confounding between effects of zinc and the effect of the presence of higher organisms in the test system. 2. Materials and methods 2.1. Test design We carried out a full factorial experiment with five concentrations of added zinc (0, 20, 60, 200 and 600 mg kg1) and four levels of biological complexity (microorganisms only, microbes and earthworms (Eisenia fetida), microbes and Italian ryegrass (Lolium multiflorum var. Macho), and microbes, ryegrass and earthworms. Three replicates were run per treatment. 2.2. Test soil As a test soil, we used a sandy loam to which 3% (dry wt/ dry wt) of sludge was added. Soil was air-dried at 20 C for about one week, and then sieved through a 2 mm mesh. Sludge was dried at 40 C for 72 h and then ground to ease homogeneous mixing into soil. The soil had the following particle size distribution: sand (63–2000 lm) 76.1%, silt (2–63 lm) 14.6%, clay (<2 lm) 9.3%. Total organic carbon content was 1.6%, and the soil pHH2 O and the total cation exchange capacity (CEC) were 6.2, and 120 mmol(+) kg1, respectively. The sludge was mixed thoroughly in with the soil by hand prior to addition of zinc. We added the sludge as food for the earthworms and nutrient source for microorganisms and plants. The sludge was sampled at a waste water treatment plant (‘‘RA2’’) three months before the experiment started. E. fetida is easily cultivated in this sludge type, and sludge was selected over more common types of earthworm food (i.e., horsedung) because of its relevance for metal polluted (sludge amended) agricultural land. The sludge had an organic matter content of 80%, and chemical analyses showed low levels of heavy metals (in mg kg1 dry wt; Cu 138, Zn 267, Pb 22, Cd < 2, Ni 25, Cr 25, Hg 0.3) and organic contaminants (e.g., R16 PAH was 0.7 mg kg1 dry wt). 2.3. Pilot study A pilot study was performed to check if the added sludge was sufficient as a carbon and nitrogen source for microbes and to determine appropriate Zn concentrations for the main experiment. The pilot study was performed with only microbes present, using the same soil/sludge mixture as in the main experiment, and with added zinc concentrations of 0, 10, 30, 100, 300, 1000 and 3000 mg kg1 dry wt. With the exception of using sludge as an N-source for bacteria, the test followed the procedure standardized

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by ISO (1997). Estimated EC10 and EC50-values (95% c.i.) for the effect of zinc on nitrification from the pilot study were 270 (190–310) mg kg1 and 927 (869–968) mg kg1, respectively, based on the added zinc concentrations (zinc concentrations were not measured). 2.4. Test substance ZnCl2 (p.a. quality), purchased from Sigma–Aldrich, was dissolved in water and added in a volume corresponding to 60% of the water holding capacity of the soil, and mixed thoroughly in with the soil/sludge mixture to obtain a homogeneous distribution. The zinc concentrations added to the test soil/sludge mixture in the main experiment were based on the results from the pilot study, and were as follows: 0 (control), 20, 60, 200 and 600 mg kg1 soil dry weight. 2.5. Higher organisms Earthworms (Eisenia fetida Savigny) were obtained from a laboratory culture kept at 20 C, with a light:dark cycle of 16:8 h, and fed horse dung. For the tests, clitellate, non-synchronized worms of similar size (0.2–0.3 g) were selected. They were allowed to empty their gut in petridishes containing a moist filter paper for 24 h, after which their total mass (10 worms) was determined to the nearest 0.01 g. Ten individuals were added to each test container in the treatments where earthworms were included. Untreated seeds of ryegrass (Lolium multiflorum var. Macho) were ˚ s, Norway). For the treatpurchased from Felleskjøpet (A ments where plants were included, 0.5 g of seeds was added to each test container. 2.6. Test conditions After 48 h of equilibration of the zinc-treated soil, earthworms and ryegrass seeds were added. Tests were conducted in climate rooms with air temperatures between 18 and 28 C (soil temperature was 22 ± 2 C), and the light:dark cycle was set to 16:8 h. Test containers used were 5 l plastic buckets containing approximately 3 kg (dry wt) of test soil. The weight of each container was determined at the start of the experiment, and watering was performed at 1–3 day intervals, according to weight loss. At each time of watering, containers were re-randomised for the position they occupied in the climate room. At the end of the test, each replicate (container) was handled separately. Ryegrass was harvested (above-ground biomass), dried to constant weight at 40 C, and the dry weight was recorded. Earthworms were collected by hand, allowed to empty their gut on moist filter-papers placed in petri-dishes for 24 h, weighed, frozen at 18 C, and subsequently dried to constant weight at 40 C. Data on earthworm reproduction was not collected. Soil samples were collected by mixing four sub-samples from each test container. In containers containing ryegrass, roots were carefully removed. Samples were taken for analysis of Zn (30 g), nitrification potential

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(50 g), and soil microbial community composition (5 g), and frozen immediately at 18 C. 2.7. Potential ammonium oxidation (PAO) The potential ammonium oxidation rate was assayed as accumulated nitrite according to the short incubation, chlorate inhibition technique described by Belser and Mays (1980) and modified by Hansson et al. (1991). This procedure was chosen because an analysis of nitrate in soils, as recommended by ISO (1997), failed to give useful information for containers with plants (i.e., no nitrate left). PAO, on the other hand, gives a direct measure of the ammonium oxidation capacity at the time of sampling. In short, ammonium oxidizing bacteria were exposed to ammonium sulphate (4 mM) and sodium chlorate (15 mM) in a soil slurry in phosphate buffer (1 mM) at pH 7.2 on a reciprocal shaker (160 rpm) at 20 C. Sampling started after 2 h, after which five samples of 2.5 ml were transferred on an hourly basis to centrifuge tubes containing 2.5 ml 4 M KCl. The samples were centrifuged for 5 min at 15 000 · G. The presence of chlorate inhibits the conversion of nitrite to nitrate. Nitrite was determined by diazotizing with sulphanilamide followed by coupling with N-(1-naphthyl) ethylenediamine dihydrochloride. The resulting water soluble dye has a magenta colour that was read at 545 nm, and the nitrite concentration was determined by comparing with a calibration curve. 2.8. Microbial community structure PLFA analyses were only performed at the 0, 60, 200 and 600 mg zinc/kg soil treatments. The phospholipid extraction and PLFA analysis were performed as previously described by Frostega˚rd et al. (1993) with some modifications. Briefly, lipids from 3 g (fresh weight) of soil were extracted with a mixture (1:2:0.8) of chloroform, methanol and citrate buffer (0.15 M, pH 4.0). The lipid-containing layer was retrieved and the lipids were fractionated on silicic acid columns (Varian, Bond Elut Si, Middelburg, the Netherlands) to obtain phospholipids. The phospholipids were subjected to mild-alkali methanolysis and the resulting fatty acid methyl esters were separated by gas chromatography. To quantify individual PLFAs in the samples, known amounts of methyl nonadecanoate (19:0; Larodan Fine Chemicals, Malmo¨, Sweden) were added as internal standard before methanolysis. The fatty acid methyl esters were analysed in a splitless mode with a Perkin–Elmer (Norwalk, US) gas chromatograph (GC) equipped with a flame ionization detector (FID). GC conditions were as described in Frostega˚rd et al. (1993) except that helium was used as a carrier gas in the present study. Detector response was calibrated by using commercially purchased mixtures (Larodane) of FAMEs (C12–C22). Individual fatty acid methyl esters were identified using a pre-established data-base and the software TurboChrom (version 6.1.1; Perkin–Elmer).

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Fatty acids are designated in terms of the total number of carbon atoms: number of double bonds, followed by the position of the double bond from the methyl end of the molecule. cis and trans Configurations are indicated by c and t, respectively. The prefixes a and i indicate anteiso and iso branching; br indicates an unknown methyl branching position; and cy refers to cyclopropane fatty acids. 10Me indicates a methyl group on the 10th carbon atom from the carboxyl end of the molecule. In total, 34 PLFAs were measured. The amount of bacterial biomass (bactPLFA) was estimated from the summed amount of the following PLFA: i15:0, a15:0, 15:0, i16:0, 16:1x9, 16:1x7t, i17:0, a17:0, cy17:0, 17:0, 18:1x7, and cy19:0, according to Frostega˚rd and Ba˚a˚th (1996). PLFA 18:2x6.9 was used as a marker for fungi (Federle, 1986) (fungPLFA), and arachidonic acid (PLFA 20:4) as a marker for protozoa (Lechevalier and Lechevalier, 1988; Frostega˚rd et al., 1997; Mauclaire et al., 2003). 2.9. Analysis of zinc content Soils were digested with hot aqua regia and analysed by ICP-AES (Optima 3000, Perkin–Elmer) using matrixmatched standards and appropriate quality assurance.

Table 1 Measured zinc concentrations in soil samples at the beginning and end of the tests. Analysis was performed on a mixture of samples from one replicate of each of the biocomplexity treatments Amount of Zn added (mg kg1) 0 20 60 200 600

Measured concentrations of Zn (mg kg1 dry wt) Start of test

End of test

Mean concentration

90 110 160 350 690

90 110 110 290 540

90 110 135 320 615

treatments both at the beginning and end of the experiment. During the 6-week exposure period, earthworm growth was not affected within the concentration range tested (p = 0.93, data not shown), and survival was similar in the control (80%) and 600 mg kg1 treatment level (77%) (data not shown). Ryegrass above-ground biomass averaged 2.2 g dry weight per container, but was significantly (p < 0.05) reduced by about 18% at 600 mg Zn/kg, compared to control values (data not shown). The potential ammonium oxidation differed significantly (p < 0.05) among treatment groups (Fig. 1A), and was

2.10. Statistical analysis The PLFA composition (individual PLFAs expressed as mol % of the total amount of identified PLFAs) of all samples was subjected to principal-component analyses (PCA) by use of the software S-PLUS 2000 (MathSoft Engineering & Education, Inc. Cambridge, MA). Prior to analysis, the data were log 10 transformed to stabilize variance. The estimated 10% effect concentrations (EC10) for zinc on soil nitrification activity were calculated by linear interpolation in the computer programme ICp (Norberg-King, 1993), based on added concentrations of zinc. The effect of the various treatments (zinc concentration, presence or absence of earthworms and/or ryegrass) and their interactions were investigated for the response variables soil nitrification activity and some specific PLFA markers (i.e., bactPLFA, fungPLFA, protozoan PLFA, and the 16:1x7 trans/cis ratio) by analysis of variance (ANOVA) in S-PLUS 2000. Little or no differences in earthworm/plant performance were observed between the various Zn-treatments, and this made it possible to use the presence or absence of earthworms and plants as factors in the analysis, and thus have the possibility of exploring interaction effects between zinc and the presence of higher organisms in the test system. 3. Results Measured zinc concentrations in soil were similar to estimated ones (i.e., concentration in soil plus added amount of zinc), and slightly lower at the end of the tests, as compared to values measured at the beginning of the tests (Table 1). Soil pHH2 O was in the range of 6.1–6.3 in all

Fig. 1. Soil nitrification potential (A) and total microbial biomass (B) in soils after 6 weeks of exposure. Figures show average value ± standard deviation (n = 3). B = pots with only microorganisms, B + E = pots with microbes and earthworms, B + R = pots with microbes and ryegrass, B + E + R = pots with microbes, earthworms, and ryegrass present.

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Table 2 Toxicity of zinc for soil nitrifying bacteria. The EC10-value equals the zinc concentration (mg kg1) causing an estimated 10% reduction in nitrification potential of soil microorganisms, alone and in the presence of plants and/or earthworms Treatment

EC10

95% c.i.

Only microorganisms + Ryegrass + Earthworms + Ryegrass and earthworms

300 255 314 316

NEa 17–318 267–344 226–335

a

NE = could not be estimated.

Axis 1

(A) 2

Loading

0

-2 B B+E B+E+R B+R

-4 0

60

200

600

Zinc (mg/kg) Axis 2

(B) 1

0

Loading

significantly reduced at the highest zinc concentration (ANOVA/Dunnetts’ test, p < 0.05), leaving a NOEC of 200 mg kg1. PAO increased in the presence of earthworms (ANOVA, p < 0.0001) and decreased in the presence of ryegrass (ANOVA, p < 0.0001). In the presence of both earthworms and plants, effects on PAO were additive (no interaction). There was a significant interaction between the effect of zinc and the presence of earthworms (ANOVA, p < 0.05) on PAO, resulting from an increased nitrification potential in all concentrations but the highest one, thus giving a more pronounced effect of zinc at the highest test concentration in the presence of earthworms (Fig. 1A). However, on the basis of EC10-values (Table 2), which range from 255 to 316 mg kg1, the net effect of zinc on soil nitrification activity was similar for all treatment groups. For measurements of the total microbial biomass, estimated as sum of the 34 PLFAs (Fig. 1B), there was a significant positive effect of ryegrass (ANOVA, p < 0.0001), but no clear effect of zinc (ANOVA, p = 0.089), or earthworms (ANOVA, p = 0.48). The PCA on the 34 PLFAs showed that the main feature of the microbial community structure (PC1 or Axis 1) changed in a monotonous fashion with zinc concentration (Fig. 2A). Axis 1 explained approximately 61% of the variance. Axis 2 explained an additional 13% of the variance and was mainly related to the presence/absence of earthworms and ryegrass (or both) (Fig. 2B). Thus, zinc was the main determinant of microbial community structure, although biological complexity seemed to have a certain effect as well. The effect of zinc and biocomplexity on the relative contribution (mol %) of some indicator PLFAs was also assessed (Fig. 3). Bacterial PLFA levels were negatively affected both by zinc (ANOVA, p < 0.0001), and by the presence of earthworms (ANOVA, p = 0.010), but increased in the presence of ryegrass (ANOVA, p = 0.062); there was a tendency of an interaction between earthworms and zinc (ANOVA, p = 0.070), with the negative effect of earthworms on bactPLFA levels decreasing with increasing zinc concentration. Fungal PLFA increased in the presence of zinc (ANOVA, p < 0.0001), and this effect of zinc was reduced in the presence of earthworms (ANOVA, p = 0.03 for the interaction effect of zinc and earthworms). Protozoan PLFA increased with increasing

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-1 B B+E B+E+R B+R

-2

-3 0

60

200

600

Zinc (mg/kg) Fig. 2. Scores of the two first PCA axes based on the 34 PLFAs as a function of zinc concentration and biological complexity. B = pots with only microorganisms, B + E = pots with microbes and earthworms, B + R = pots with microbes and ryegrass, B + E + R = pots with microbes, earthworms, and ryegrass present.

zinc concentrations (ANOVA, p < 0.0001), and was negatively affected by the presence of ryegrass (ANOVA, p < 0.001), but not by earthworms (ANOVA, p = 0.64). The trans/cis ratio of the fatty acid 16:1x7 decreased with increasing zinc concentrations (ANOVA, p < 0.0001), and it was also lower both in the presence of ryegrass (ANOVA, p < 0.001) and earthworms (ANOVA, p = 0.040). 4. Discussion Soil organisms may change the soil both physically (e.g., earthworm burrows, plant root intrusion, water uptake) and chemically (e.g., uptake of essential nutrients, introduction and decomposition of organic matter, changes in pH, O2-level, water content, etc.). Such changes may directly or indirectly affect the bioavailability of metals to other organisms.

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Fig. 3. Indicator PLFAs for various groups of microorganisms exposed to zinc for 6 weeks, in the presence or absence of higher organisms. Figures show average value ± standard deviation of the relative biomass (mol %) of each group (n = 3). B = pots with only microorganisms, B + E = pots with microbes and earthworms, B + R = pots with microbes and ryegrass, B + E + R = pots with microbes, earthworms, and ryegrass present.

4.1. Interaction between zinc and biocomplexity Of special interest in this study was whether or not nonadditive (interaction) effects could be observed for the presence of plants and/or earthworms, and the effect of zinc. Although not very pronounced, significant interaction effects were found in two cases: The effect of earthworms on the potential ammonium oxidation was significantly decreased in the highest concentration of zinc (Fig. 1A), and effect of zinc on the relative biomass of fungi (Fig. 3C) was significantly reduced in the presence of earthworms. Whether or not the zinc response of individual groups of microorganisms is affected by the presence of earthworms or plants depends on whether their local microhabitat is physically affected at all, and if so, whether the changes in zinc availability make their habitat more or less favourable for them. Even though our results do not unravel the mechanisms behind the observed interaction effects, some possible explanations can be found in the literature, as described in the Introduction. 4.2. Effects of earthworms and ryegrass By their introduction, higher organisms, such as earthworms and ryegrass, potentially introduce new species of

microorganisms into the test system, and they also change the habitats of microorganisms by changing the physicochemical properties of the test soil. The presence of earthworms has in many cases been found to stimulate soil microbial activity (Binet et al., 1998) and increase the number of bacteria (Devliegher and Verstraete, 1997). For soil nitrifying bacteria, the earthworm castings and burrow walls (drilosphere) have  been found to be enriched with NHþ 4 and NO3 , which adds local conditions with a high potential for microbial nitrification and denitrification (Parkin and Berry, 1999). Thus, the significant increase in soil nitrification that was observed in the presence of earthworms (Fig. 1A) was not surprising. Presence of ryegrass increased the total microbial biomass (Fig. 1B) in our study, which is in line with results of Olsson et al. (1996), who found increased bacterial numbers and activity when plants were present. However, we found that ryegrass had a negative effect on soil nitrification (Fig. 1A), which might possibly be caused by plants competing for ammonia (Schimel et al., 1989) and thereby decreasing the population of ammonia-oxidizing bacteria. There was also a clear shift in PLFA pattern in the presence of ryegrass (Fig. 2), and for the specific PLFA markers (Fig. 3) we observed increased relative (mol %)

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bactPLFA levels and decreased relative protozoan PLFA levels in the presence of ryegrass. Ryegrass was the only factor found to affect the trans/cis ratio of PLFA 16:1x7, which is used by microorganisms to adjust the fluidity of their membrane. This ratio has been observed to change in case of nutrient deprivation (Guckert et al., 1986), but the observed change could also be due to changes in microbial species composition. 4.3. Effects of zinc Zinc did not affect the total microbial biomass (Fig. 1B), but it had a significant (p < 0.05) effect on most of the other microbial parameters measured. The microbial community composition (PLFA pattern) was grouped differently from the controls, even at the lowest level of added zinc where PLFA composition was measured (60 mg kg1) (Fig. 2). Many previous studies have documented shifts in the microbial community structure under the influence of metals (e.g., Frostega˚rd et al., 1993; Frostega˚rd et al., 1996; Ba˚a˚th et al., 1998), at similar concentration levels. One of the most pronounced changes in our system was the increased fungal/bacterial biomass ratio with increasing concentrations of zinc. Our data on the total (PLFA) biomass in the test system (i.e., not the relative (mol %) data as shown in the figure) show that this shift is caused by fungal growth being stimulated by zinc, while the gross bacterial biomass was more or less unaffected. These findings are in agreement with the general view that fungi are more resistant to zinc than bacteria (Doelman, 1985), and supports the findings of Frostega˚rd et al. (1993) and Rajapaksha et al. (2004) who documented an increased fungal/ bacterial biomass ratio (estimated on the basis of PLFA analysis) upon exposure of soil communities to zinc. An increased ratio of the trans/cis isomers of monounsaturated PLFAs has been interpreted as a response to toxicity (Heipieper and De Bont, 1994), for example this ratio increased with increasing Zn concentrations when a strain of Pseudomonas putida was grown in pure culture (Heipieper et al., 1996). In the present study the trans/cis ratio of 16:1x7 instead decreased with increasing Zn concentrations (Fig. 3D). Frostega˚rd et al. (1996) found increased trans/cis ratio in response to Zn in a forest soil, but not in an arable soil. They concluded that changes in the trans/cis ratio for complex microbial communities cannot be taken as general toxicity indices, and that care must be taken since a change in this index can instead be due to a change in the species composition of the microbial community. The threshold values for effects of added zinc on soil nitrification found in this study (i.e., a NOEC of 200 mg kg1 and EC10-values of 255–316 mg kg1) were similar to NOEC-values of 400 and 200 mg kg1 found by Cheng and Wong (2002) and Yin et al. (2003), respectively. The effect of zinc on the survival and reproduction of the earthworm species E. fetida has been documented in several soil types, with effect concentrations within the

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concentration range used in this study. In a soil type similar to the one used here, Spurgeon and Hopkin (1996) found a LC50 value of 620 mg kg1 and an EC50-value for effects on reproduction of 274 mg kg1, and Lock and Janssen (2001a) found LC50 and EC50 values of approximately 1000 and 300 mg kg1, respectively. Thus, in the highest test concentrations, earthworms may have been affected even if it was not apparent from the growth data. A possible effect on earthworms in the highest test concentration is also supported by the fact that the positive effect on PAO was reduced in this treatment (see Fig. 1A). Although the present study does not unravel the mechanisms behind the observed effects of changes in the biocomplexity of a test system, it shows the possibility that the toxicity and bioavailability of zinc in soil may change in the presence of higher organisms such as invertebrates. Acknowledgements This study was supported by the Research Council of Norway. We greatly acknowledge Hege Bergheim and Geir ˚ sli for technical assistance and Knut Sørensen for sample A preparation and analysis of PLFA. References ˚ ., Campbell, C.D., 1998. Effect Ba˚a˚th, E., Dı`az-Ravin˜a, M., Frostega˚rd, A of metal-rich sludge amendments on the soil microbial community. Appl. Environ. Microbiol. 64, 238–245. Belser, L.W., Mays, E.L., 1980. Specific-inhibition of nitrite oxidation by chlorate and its use in assessing nitrification in soils and sediments. Appl. Environ. Microbiol. 39, 505–510. Binet, F., Fayolle, L., Pussard, M., Crawford, J.J., Traina, S.J., Tuovinen, O.H., 1998. Significance of earthworms in stimulating soil microbial activity. Biol. Fertil. Soils 27, 79–84. Cheng, J., Wong, M.H., 2002. Effects of earthworms on Zn fractionation in soils. Biol. Fertil. Soils 36, 72–78. Devliegher, W., Verstraete, W., 1997. Microorganisms and soil physicochemical conditions in the drilosphere of Lumbricus terrestris. Soil Biol. Biochem. 29, 1721–1729. Doelman, P., 1985. Resistance of soil microbial communities to heavy metals. In: Jensen, V., Kjøller, A., Sørensen, L.H. (Eds.), Microbial Communities In Soil. Elsevier, London, pp. 369–384. Federle, T.W., 1986. Microbial distribution in soil: new techniques. In: Meguar, F., Gantar, M., (Eds.), Perspectives in Microbial Biology. Slovene Society for Microbiology, Ljubljana, Slovenia, pp. 493–498. ˚ ., Ba˚a˚th, E., 1996. The use of phospholipid fatty acid Frostega˚rd, A analysis to estimate bacterial and fungal biomass in soil. Biol. Fertil. Soils 22, 59–65. ˚ ., Tunlid, A., Ba˚a˚th, E., 1993. Phospholipid fatty acid Frostega˚rd, A composition, biomass, and activity of soil microbial communities from two soil types experimentally exposed to different metals. Appl. Environ. Microbiol. 59, 3605–3617. ˚ ., Tunlid, A., Ba˚a˚th, E., 1996. Changes in microbial Frostega˚rd, A community structure in two metal contaminated soils during longterm incubation. Soil Biol. Biochem. 28, 55–64. ˚ ., Petersen, SO., Baath, E., Nielsen, T.H., 1997. Dynamics of Frostega˚rd, A a microbial community associated with manure hot spots as revealed by phospholipid fatty acid analyses. Appl. Environ. Microbiol. 63, 2224–2231. Giller, K.E., Witter, E., McGrath, S.P., 1998. Toxicity of heavy metals to microorganisms and microbial processes in agricultural soils: a review. Soil Biol. Biochem. 30, 1389–1414.

1754

L.E. Sverdrup et al. / Chemosphere 65 (2006) 1747–1754

Guckert, J.B., Hood, M.A., White, D.C., 1986. Phospholipid ester-linked fatty acid profile changes during nutrient deprivation of Vibrio cholerae: increases in the trans/cis ratio and proportion of cyclopropyl fatty acids. Appl. Environ. Microbiol. 52, 794–801. Hansson, G.B., Klemedtsson, L., Stenstro¨m, J., Torstensson, L., 1991. Testing the influence of chemicals on soil autotrophic ammonium oxidation. Environ. Toxicol. Water 6, 351–360. Heipieper, H.J., De Bont, J.A.M., 1994. Adaptation of Pseudomonas putida S12 to ethanol and toluene at the level of fatty-acid composition of membranes. Appl. Environ. Microbiol. 60, 4440–4444. Heipieper, H.J., Meulenbeld, G., VanOirschot, Q., De Bont, J.A.M., 1996. Effect of environmental factors on the trans/cis ratio of unsaturated fatty acids in Pseudomonas putida S12. Appl. Environ. Microbiol. 62, 2773–2777. ISO, 1997. Soil quality—biological methods—determination of nitrogen mineralization and nitrification in soils and the influence of chemicals on these processesISO Method 14238. International Organization for Standardization, Geneva, Switzerland. Lechevalier, H., Lechevalier, M.P., 1988. Chemotaxonomic use of lipids— an overview. In: Ratledge, C., Wilkinson, S.G. (Eds.), Microbial Lipids. Academic Press, London, UK, pp. 869–902. Lock, K., Janssen, C.R., 2001a. Ecotoxicity of zinc in spiked artificial soils versus contaminated field soils. Environ. Sci. Technol. 35, 4295–4300. Lock, K., Janssen, C.R., 2001b. Modelling zinc toxicity for terrestrial invertebrates. Environ. Toxicol. Chem. 20, 1901–1908. Lock, K., Janssen, C.R., de Coen, W.M., 2000. Multivariate test design to assess the influence of zinc and cadmium bioavailability in soils on the toxicity to Enchytraeus albidus. Environ. Toxicol. Chem. 19, 2666– 2671. Ma, Y., Dickinson, N.M., Wong, M.H., 2003. Interactions between earthworms, trees, soil nutrition and metal mobility in amended Pb/Zn mine tailings from Guangdong, China. Soil Biol. Biochem. 35, 1369– 1379. Mauclaire, L., Pelza, O., Thullnera, M., Abrahamb, W.R., Zeyera, J., 2003. Assimilation of toluene carbon along a bacteria–protist food chain determined by 13C-enrichment of biomarker fatty acids. J. Microbiol. Methods 55, 635–649. McLaughlin, M.J., 2003. Bioavailability of metals to terrestrial plants. In: Allen, H.E. (Ed.), Bioavailability of metals in terrestrial ecosystems: importance of partitioning for bioavailability to invertebrates, microbes and plants. SETAC Press, Pensacola, FL, USA, pp. 39–68. McLaughlin, M.J., Smolders, E., 2001. Background zinc concentrations in soil affect the zinc sensitivity of soil microbial processes—a rationale for a metalloregion approach to risk assessments. Environ. Toxicol. Chem. 20, 2639–2643.

Norberg-King, T.J., 1993. A linear interpolation method for sub-lethal toxicity: the inhibition concentration (ICp) approach. (Version 2.0). Software Package and User’s Guide. National Effluent Toxicity Assessment Center, Duluth, MN, USA. Olsson, P.A., Ba˚a˚th, E., Jakobsen, I., So¨derstro¨m, B., 1996. Soil bacteria respond to presence of roots but not to mycelium of arbuscular mycorrhizal fungi. Soil Biol. Biochem. 28, 463–470. Parkin, T.B., Berry, E.C., 1999. Microbial nitrogen transformations in earthworm burrows. Soil Biol. Biochem. 31, 1765–1771. Rajapaksha, R.M.C.P., Tobor-Kaplon, M.A., Ba˚a˚th, E., 2004. Metal toxicity affects fungal and bacterial activities in soil differently. Appl. Environ. Microbiol. 70, 2966–2973. Salminen, J., van Gestel, C.A.M., Oksanen, J., 2001. Pollution induced community tolerance and functional redundancy in a decomposer food web in metal stressed soil. Environ. Toxicol. Chem. 20, 2287–2295. Schimel, J.P., Jackson, L.E., Firestone, M.K., 1989. Spatial and temporal effects on plant-microbial competition for inorganic nitrogen in a California annual grassland. Soil Biol. Biochem. 21, 1059–1066. Sharma, S.S., Schat, H., Vooijs, R., van Heerwaarden, L.M., 1999. Combination toxicology of copper, zinc, and cadmium in binary mixtures: concentration dependent antagonistic, nonadditive, and synergistic effects on root growth in Silene vulgaris. Environ. Toxicol. Chem. 18, 348–355. Smit, C.E., van Gestel, C.A.M., 1998. Effects of soil type, prepercolation, and ageing on bioaccumulation and toxicity of zinc for the springtail Folsomia candida. Environ. Toxicol. Chem. 17, 1132–1141. Spurgeon, D.J., Hopkin, S.P., 1996. Effects of variations of the organic matter content and pH of soils on the availability and toxicity of zinc to the earthworm Eisenia fetida. Pedobiologia 40, 80–96. Van Beelen, P., Fleuren-Kemila¨, A.K., 1997. Influence of pH on the toxic effects of zinc, cadmium, and pentachlorophenol on pure cultures of soil microorganisms. Environ. Toxicol. Chem. 16, 146–153. Van Gestel, C.A.M., Hensbergen, P.J., 1997. Interaction of Cd and Zn toxicity for Folsomia candida Willem (Collembola, Isotomidae) in relation to bioavailability in soil. Environ. Toxicol. Chem. 16, 1177– 1186. Yin, S., Yang, L., Yin, B., Mei, L., 2003. Nitrification and denitrification activities of zinc-treated soils worked by the earthworm Pheretima sp. Biol. Fertil. Soils 38, 176–180. Yu, X., Cheng, J., Wong, M.H., 2005. Earthworm-mycorrhiza interaction on Cd uptake and growth of ryegrass. Soil Biol. Biochem. 37, 195–201. Zorn, M.I., Van Gestel, C.A.M., Eijsackers, H., 2005. Effect of Lumbricus rubellus and Lumbricus terrestris on zinc distribution and availability in artificial soil columns. Biol. Fertil. Soils 41, 212–215.