- Email: [email protected]
Functional resilience of microbial communities from perturbed upland grassland soils to further persistent or transient stresses
ARTICLE IN PRESS
Soil Biology & Biochemistry 38 (2006) 2300–2306 www.elsevier.com/locate/soilbio
Functional resilience of microbial communities from perturbed upland grassland soils to further persistent or transient stresses H.L. Kuana, C. Fenwickb,1, L.A. Gloverb, B.S. Griffithsa,, K. Ritzc a
Environment–Plant Interactions Programme, Scottish Crop Research Institute, Invergowrie, Dundee DD2 5DA, UK b School of Medical Sciences, Institute of Medical Sciences, University of Aberdeen, Aberdeen AB25 2ZD, UK c National Soil Resources Institute, Cranfield University, Silsoe, Bedfordshire MK45 4DT, UK Received 21 October 2005; received in revised form 17 February 2006; accepted 21 February 2006 Available online 3 April 2006
Abstract The microbial functioning of soils following perturbation was assessed at a temperate upland grassland site, maintained by the Soil Biodiversity and Ecosystem Function Programme at Sourhope Research Station, Scotland. Published results indicated that the soil microbial communities were resilient to these initial perturbations; in this paper we tested whether they were equally resilient to a subsequent perturbation. Soil samples were taken from field plots receiving treatments that represented different forms of perturbation, viz. reseeding, application of sewage-sludge, biocide or nitrogen plus lime, and a non-perturbed control. Functional resilience following further perturbation comprising a transient heat or persistent copper perturbation was assessed over 28 days, by monitoring the shortterm decomposition of added plant residues. Bacterial community structure was assessed by DGGE separation of eubacterial 16S rDNA PCR products. PCR-DGGE did not distinguish any significant difference ðP40:05Þ between the bacterial communities of soils under different treatments, showing differences only between treated soils and the untreated, control soils. Two days after the application of stresses, functional capability differed in soils under different treatments. Soil samples from all the treated plots were less resilient to heat stress than soil from control plots. The initial reduction in decomposition following the addition of copper differed between treatments, but function had not recovered in any of the Cu-amended soils within 28 days. Soil resilience varied according to the type and duration of stress applied, microbial activity, soil characteristics and treatment regimes. The initial resistance of function to stress was not predictive of recovery of function over time. r 2006 Elsevier Ltd. All rights reserved. Keywords: Function; Grassland; Resilience; Soil microbial community; Stability
1. Introduction Research published to date from the major research programme, Soil Biodiversity and Ecosystem Function (SBP), funded by the UK Natural Environment Research Council has recently been reviewed (Fitter et al., 2005). To achieve the SBP’s objective, a grassland ecosystem was perturbed by removing taxa using a biocide and by subjecting discrete plots to various treatments representative of anthropogenic inputs and designed to affect Corresponding author. Tel.: +44 1382 562731; fax: +44 1382 568502.
E-mail address: bryan.griffi[email protected] (B.S. Griffiths). Present address: The Natural Environment Research Council, Polaris House, North Star Avenue, Swindon SN2 1EU, UK. 1
0038-0717/$ - see front matter r 2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.soilbio.2006.02.013
biodiversity. Many of the treatments had the expected substantial impacts on various groups of soil biota but some broad-scale processes were apparently unaffected by treatment. For example, liming had no detectable effect on the rate of soil respiration, suggesting that at this broad process-level the system was resistant to perturbation (Gray et al., 2003). Similar results were obtained by imposing specific stresses on the system, including drought, heat and the application of sewage sludge (Griffiths et al., 2003). Even in communities reconstructed by body-size exclusion techniques there were no differences in overall productivity or ecosystem carbon (C) exchange rate, suggesting that these systems are extremely resilient (Bradford et al., 2002). Although these soils appear able to retain function even when their biological structure has
ARTICLE IN PRESS H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
been radically altered, how long they can sustain this and how far we can push them remains to be seen (Fitter et al., 2005). Studies in which the biodiversity of soil from the SBP site was experimentally manipulated by perturbation showed that the resilience of organic matter decomposition in the soil was related to a specific microbial community structure rather than biodiversity per se (Griffiths et al., 2004). Another interpretation of these results is that the decomposition of organic matter is less resilient in soils that have been previously stressed (de Ruiter et al., 2002). Samples from soil amended experimentally with heavy metal contaminated sewage sludge showed that the resilience of organic matter decomposition to subsequent perturbation was affected by the previous land use management (Griffiths et al., 2005). While the evidence is that functions of soil communities at the SBP site are apparently resilient to a range of perturbations (Fitter et al., 2005; Gray et al., 2003; Griffiths et al., 2003), it is unclear whether such communities also show resilience to further perturbations. We tested the hypothesis that soil communities already subjected to an environmental stress would be less resilient to a subsequent stress. The resilience of short-term decomposition of plant material in SBP soils further subjected to heating and Cu stresses was taken as being representative of a transient and a persistent stress as used in previous studies (Griffiths et al., 2005). 2. Materials and methods
2301
were approximately at field-capacity when sampled, so were spread out and aired at 18 1C for 24 h to facilitate sieving. The litter layer, visible fauna and vegetation were removed. Soils were passed through a 4 mm metal sieve and stored field moist at 4 1C prior to analysis or assay. The soil moisture content (105 1C) and soil pH for each plot were determined. The ratios of soil volumes from different horizons present in each core were maintained in the homogenised bulk samples, e.g. if a sample consisted of 30% soil from the H-horizon and 70% from the Ahhorizon, the soils from each horizon were sieved separately and mixed together at a ratio of 3:7. Soil samples of 450 g were packed into nylon mesh-based polyethylene cylinders at a density of 1 g cm3. The packed soils were placed on a tension table (Ball and Hunter, 1988) and equilibrated to a matric potential of 50 cm water (5 kPa) for 6 days. 2.2. Soil properties Soil properties were measured at the start of the experiment, using bulked soil. Microbial biomass C was determined as the flush of C following chloroformfumigation-extraction (Vance et al., 1987). NO 3 –N, NH+ 4 –N and dissolved organic C (DOC) were measured following extraction in 1 M KCl (Wheatley et al., 1989). + NO 3 –N and NH4 –N and DOC were measured concurrently by colorimetric analysis using a segmented-flow autoanalyser (Skalar Analytical, Breda, Netherlands). Soil pH was measured in a 1:5 soil–0.01 M CaCl2 solution (w/v).
2.1. Experimental site and sampling 2.3. Soil respiration and functional stability Soil was collected from the Rigg Foot experimental site of the Soil Biodiversity and Ecosystem Function Programme at Sourhope Research Station (21 15 W, 551 30 N, UK Ordnance Survey Grid Reference NGR NT854196), in the Cheviot hills, southern Scotland. Rigg Foot is an upland grassland site, 309 m above sea level, comprising a freely drained Brown Forest soil of the Sourhope Series (Muir, 1956) with some gleying at the lower reaches. Vegetation at the site was predominantly Festuca ovinaAgrostis capillaris-Galium saxitale grassland (NVC community U4d); (Rodwell, 1992). A randomised block design was used, with three of the five replicate plots of five treatments being sampled in November 2001: untreated (control); biocide application, N and lime application; sewage sludge application; and reseeding. The N and lime plots received 600 g CaCO3 m2 y1 and 24 g NH4NO3 m2 y1. Anaerobically digested domestic sewage sludge was applied to the sewage plots at 10 l m2 in August 1999 and 20 l m2 in May 2000. Biocide-treated plots received Dursban 4 (chlorpyriphos: Dow AgroSciences, Indianapolis, USA) 1.5 l ha1 y1. The indigenous turf was removed from the reseeded plots in May 1999 and replaced with Lolium perenne L. Three replicate plots for each treatment were sampled with a metal trowel to 100 mm depth. The soils
The rate of carbon mineralisation (as measured by the short-term decomposition of plant residues) was selected as a measure of soil functional capability, based on the procedure of Griffiths et al. (2001). Composite grass cuttings (C:N ratio 20:1) were obtained from an untreated section of the Sourhope site under permanent pasture. The cuttings were dried at 60 1C and ball-milled to a fine, uniform consistency. Thirty replicate aliquots of 10 g equilibrated soil from each sample under consideration were transferred to 25 ml polypropylene scintillation vials. These soils were apportioned in batches of ten to either a heat or Cu perturbation, or a non-stressed control. The 10 heat-perturbed soils were amended with 100 ml sterile deionised water, sealed with screw-capped lids and heated at 40 1C for 18 h, cooled to 15 1C and uncapped. The ten Cu-perturbed soils were amended with CuSO4.5H2O to a concentration of 1 mg Cu g1. The ten control soils were amended with 100 ml sterile deionised water. After perturbation, all vials were incubated in a moist atmosphere at 15 1C. CO2 evolution was measured using a capillary gas chromatograph (Hewlett Packard HP5890) with a TCD detector at a GC column temperature of 90 1C. Peak areas were calculated using a SP4270 integrator (Spectra-Physics, CA, USA). Basal respiration was determined from
ARTICLE IN PRESS 2302
H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
cumulative 24 h CO2 evolution in duplicate 10 g unamended soil samples. Microbial biomass-specific respiration was calculated as the evolution of CO2 per unit of biomass C-flush (mg CO2-C mg microbial C1 h1) (after Anderson and Domsch, 1993).
the resulting data matrices analysed by principal coordinate analysis (PCO) and the resultant PCO scores analysed by ANOVA. To overcome potential bias in the results due to inconsistencies in the DGGE gels, samples were randomised in lanes across a gel, and only samples run on the same gel were compared statistically.
2.4. Bioassay of soil pore water and acute Cu toxicity An assay using the lux-marked luminescent biosensor Escherichia coli HB101 (pUCD607) (Shaw and Kado, 1986) was carried out on soil pore water extracts from the SPB plots, experimentally spiked with Cu. The assay was used to determine whether observed differences in functional stability to the copper stress were due to soil effects on copper bioavailability, or to the capacity of the microbial community to respond to added copper. Triplicate soil samples of 20 g from each of the sampled plots were either amended to 1000 mg kg1 Cu or unamended, then left for 28 d at 15 1C. The soils were diluted to 50% w/v in sterile deionised water, left to equilibrate for 24 h at 15 1C and centrifuged for 10 min at 5000 g. The supernatant was filtered through 0.7 mm GF/F glass microfibre syringe filters (Whatman International, Maidstone, UK) and frozen at 20 1C until use. The soil pore water extracts from the Cu-amended soils were diluted to 5% v/v in double-deionised water at pH 5.5 to bring the Cu load within the optimal range of microbial biosensor sensitivity. Cu standards of 0.1 mg l1 to 2 mg l1 were prepared in double deionised water from a 1111 mg l1 stock solution of Cu (3.929 g l1 CuSO4.5H2O). The biosensor was resuscitated from a freeze-dried culture and exposed to the soil solutions following the method of Paton et al. (1997). Sample luminescence was determined using a Lucy Anthos 1 microplate-reading luminometer and Stingray software (Anthos Labtec, Salzburg, Austria). Luminescence was measured after 15 minutes and expressed as relative light units (RLU), with 1 RLU being equivalent to 100 mV s1.
3. Results 3.1. Soil properties Soil pH was significantly ðPo0:05Þ elevated in the N and lime plots compared to the other treatments (Table 1). Variation within the treatments was larger than between treatments for NH+ 4 –N, NO3 –N or DOC and there were no significant differences ðP40:05Þ between the treatments (Table 1). Microbial biomass C-flush (chloroform-labile C) was significantly different ðPo0:01Þ between treatments (Table 1). Microbial biomass showed a 3-fold variation between treatments, being greatest in the sewage-amended plots and smallest in the control and reseeded plots (Po0:05; Table 1). 3.2. Soil respiration and functional stability There was no significant difference ðP40:05Þ in basal respiration or grass-induced respiration rates between treatments, but values for respiration per unit microbial biomass were significantly different ðPo0:05Þ and highest in soils from control and reseeded plots (Table 2). One day after Cu-amendment, functional stability to Cu differed significantly ðPo0:05Þ between treatments. N and lime treated soils were the most resistant to Cu, with 94% functional capability relative to a non-perturbed control and reseeded soils were least resistant to Cu, with 66% functional capability relative to a non-perturbed control (Fig. 1). Function did not recover in any of the soils 28 days after Cu application but there was a temporal
2.5. Microbial community structure DNA was extracted from a 0.5 g (fresh weight) subsample and PCR amplified as previously described by Griffiths et al. (2004). Denaturing gradient gel electrophoresis (DGGE) gels were prepared as described by McCaig et al. (2001) and samples run for 15 h at 60 V, 60 1C.
Table 1 Soil properties for Rigg Foot plots at Sourhope under different treatment regimes. Data represent mean +/ bracketed figures representing standard error ðn ¼ 3Þ
2.6. Data analysis
Control Biocide N and lime Sewage Reseed P LSD
Data were analysed by ANOVA using Genstat for Windows (version 7). For each sampling event at 1, 3, 8, 14 and 28 days after perturbation, the functional capability of the perturbed soil was determined to be the CO2 evolution of grass-amended, perturbed soils as a percentage of grassamended, non-perturbed soil. DGGE patterns were converted into presence/absence scores at each band position;
Treatment
NH+ 4 –N
pH
4.2 4.1 5.4 4.3 4.4 * 0.6
(0.1) (0.1) (0.2) (0.1) (0.1)
NO 3 –N
DOC
C-flush
mg N g1 dry wt soil
mg C g1 dry wt soil
24.1 12.2 6.6 16.6 5.7 NS 20.5
1.0 0.2 8.2 6.9 0.05 NS 12.1
(6.0) (4.6) (0.6) (3.6) (0.2)
0.9 o0.01 23.0 14.6 o0.01 NS 34.0
(0.3) (0.0) (12.9) (5.2) (0.0)
(0.6) (0.1) (2.9) (4.0) (0.01)
165 362 338 497 150 ** 155
(27) (12) (21) (43) (29)
Significant differences within a property are shown as *,**, for Po0:05 and Po0:01, respectively. LSD represents least significant differences of means at 5% level.
ARTICLE IN PRESS H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
2303
Table 2 Basal respiration, respiration induced by addition of 10 mg powdered grass g1 soil and respiration per unit of biomass C-flush in Sourhope soils under different treatment regimes Treatment
Basal mg CO2–C g1 h1
Grass-induced
Respiration rate per unit biomass C-flush mg CO2–C g1 C-flush h1
Control Biocide N and lime Sewage Reseed P LSD
1.27 1.06 1.22 1.03 1.68 NS 0.67
12.12 9.68 17.11 13.14 11.67 NS 5.46
0.0085 0.0030 0.0036 0.0022 0.0143 * 0.0082
(0.14) (0.06) (0.07) (0.05) (0.21)
(1.83) (0.57) (0.72) (0.61) (0.68)
(0.0034) (0.0007) (0.0001) (0.0005) (0.0095)
Data represent mean +/ bracketed figures representing standard error ðn ¼ 3Þ. Significant differences are shown as *, for Po0:05. LSD represents least significant differences of means at 5% level.
Fig. 1. The percentage change in respiration (CO2 evolution over 24 h) from powdered grass added to Sourhope soils under different land management regimes, 1, 3, 8, 14, 28 days after application of 1 mg Cu g1. Bars represent means +/SE ðn ¼ 3Þ.
Fig. 2. The percentage change in respiration (CO2 evolution over 24 h) from powdered grass added to Sourhope soils under different treatments, 1, 3, 8, 14 and 28 days after perturbation by 40 1C for 18 h. Bars represent means +/SE ðn ¼ 3Þ.
convergence in the relative respiration rates of Cu-amended soils. At day 28 there was no significant difference ðP40:05Þ in the function of Cu-amended soils from different treatments, with a mean functional capability of 71%. The functional stability of soils after heat perturba-
tion differed according to soil treatment ðPo0:05Þ. One day after heating, soils from the control plots displayed the initial greatest resistance to heat perturbation with 71% of original capability (Fig. 2) and soils from the Reseeded plots were least resistant to heat with 42% of the original
ARTICLE IN PRESS H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
2304
capability. Recovery of function occurred in all soils over time, with a mean of 84% original capability at day 28. No significant difference ðPo0:05Þ was seen in the functional capability of heat-perturbed soils from plots under different treatments at Day 28. 3.3. Soil pore water and acute Cu toxicity E. coli HB101 (pUCD607) luminescence after exposure to unamended soil pore water extracts did not differ ðP40:05Þ between plot treatments (data not shown). Exposure to Cu-amended soil pore water from biocide, control, reseed and sewage plots resulted in a significant ðPo0:01Þ inhibition of biosensor luminescence relative to unamended samples (Fig. 3). There was a significant ðPo0:05Þ interaction between plot treatments and Cu amendment such that extracts from N and lime soils amended with Cu did not result in further inhibition of
% luminescence (RLU)
125 100 75 50 25
Control
Biocide
Reseeded
Sewage
N + lime
0 0
5 10 Exposure time (minutes)
15
Fig. 3. The inhibition of E.coli HB101 (pUCD607) luminescence over a 15 minute exposure to soil pore water extracts from soils amended with 1 mg g1 Cu from Sourhope plots, as % of luminescence relative to a deionised water control. Data represent mean +/SE Figures in brackets represent SE ðn ¼ 6Þ.
PCO 2 (13% of variation)
0.50
Control N+Lime Biocide Reseeded Sewage
0.25
0.00
-0.25
-0.50 -0.4
-0.2
0.0 0.2 PCO 1 (17% of variation)
0.4
Fig. 4. Plot of first and second principal co-ordinates of PCR-DGGE profiles obtained from Sourhope soils under different treatments, using eubacterial primers. Points show means (n ¼ 3; bars denote +/SE).
biosensor luminescence relative to unamended soil from the same plots. 3.4. Microbial community structure PCO analysis of PCR-DGGE profiles revealed an intrinsically high degree of similarity between eubacterial community structures in all soils. The first principal coordinate did not distinguish between any treatments, and the second co-ordinate significantly distinguished the N and lime, reseeded and sewage treatments from the control (Po0:05; Fig. 4). The community structure in the biocidetreated soils was indistinguishable from the other treatments (Fig. 4). 4. Discussion Microbial biomass C-flush and soil pH differed according to land management practice. The largest microbial biomass C-flush was observed in sewage plots, suggesting that inputs of organic C and N from sewage application supported a growth in the soil microbial biomass. Microbial biomass C-flush was significantly higher in the sewage plots, N and lime plots and biocide-treated plots than in soils from control plots. This contrasts with the results of long-term studies of temperate grasslands, where soil microbial biomass size and activity were lower in Namended areas than in unfertilised areas (Lovell et al., 1995;Yeates et al., 1997). Rates of basal respiration did not differ in soils from plots under different management regimes. This was consistent with other studies at the site (Gray et al., 2003). Similarly, rates of grass-induced respiration did not differ in soils from plots under different management regimes. Also, no significant differences were observed in concentrations of NH+ 4 –N, NO3 –N or DOC in the soils under different treatments. However, variation between plots under the same treatment was larger than variation between treatments, reflecting the heterogeneity of the site (Ritz et al., 2004). Analysis of the genetic structure of the eubacterial bacterial communities in these soils showed a significant but weak trend that the community structures from treated plots differed from that of untreated soils. This agrees closely with the results of Gray et al. (2003) who analysed the eubacterial community using TGGE and found significant but only slight, effects of the treatments. Whiteley et al. (2003) could not detect changes in eubacterial community structure in the same soil following wet-dry cycles. Studies of a different field at the same site also indicated that DGGE profiling of eubacterial DNA was generally unaffected by environmental heterogeneity (Ritz et al., 2004). Analysis of more defined taxonomic and functional groups, such as ammonia oxidisers (Gray et al., 2003), methanogens (Sheppard et al., 2005) and chitin degraders (Metcalfe et al., 2002) were able to detect changes in microbial community structure induced by the field treatments. One of the reasons for choosing the
ARTICLE IN PRESS H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
decomposition of plant residues in the assay was that it is a general function carried out by a wide range of organisms (Griffiths et al., 2000), so it would be difficult to relate changes in functionally defined microbial groups to decomposition. Our results suggest that the observed variations in functional resilience to heat and copper perturbation in soils under different treatments were not associated with changes in broad-scale microbial community structure. The microbial biosensor E. coli HB101 (pUCD607) did not distinguish between soil pore water extracts from the plots under different treatments. This suggests that differences in functional resilience were not due to soil treatments at the SBP site (e.g. sewage sludge) exerting a direct toxic effect on the soil microbial community. However, biosensor exposure to Cu-amended soil pore water extracts resulted in inhibition of luminescence relative to the unamended extracts, with the exception of Cu-amended soil pore water from the N and lime plots. The reduced Cu bioavailability in the extracts from the N and lime soils would be associated with the elevated pH due to liming treatments (Ro¨mkens and Salomons, 1998). Soils under different land management regimes varied in their functional response to Cu perturbation, but were not resilient to Cu, as measured by recovery of grassdecomposition capability over 28 days. Initially, grassdecomposition capability in soil from the N and lime and control plots was most resistant to Cu; this corresponds with the results of the Cu bioavailability study. Functional capability decreased over time in Cu-perturbed N and lime and control soils, so that by day 28 there was no significant difference (P40.05) between the functional capability in any of the soils. This temporal convergence in functional capability suggests that certain soil properties have an initial protective effect on soil microbial community function in the presence of Cu, but these factors may only offer short-term protection. Measuring the long-term resilience of microbial function may therefore be more informative than acute toxicity assays in assessing the effects of persistent perturbations such as heavy metals. A study by Renella et al. (2002) confirmed that fresh additions of heavy metals to experimental systems are not predictive of long-term effects on microbial activity. Functional responses to heat perturbation differed according to the land management practices applied to the plots. The initial resistance to heat (as measured by decreases in functional capability one day after heating) was highest in soils from the control plots and lowest in the reseeded soils. The reseeded soils were most resilient to heat (i.e. function recovered most over 28 days, but this is also related to these soils having the greatest initial reduction in function after heat) and the sewage-amended soils were least resilient to heat over a 28-day period. The underlying mechanisms of these differences are unclear and may be related to differences in soil conditions under different treatments (e.g. pH, organic matter content, physical structure, plant community composition and productivity).
2305
5. Conclusions The results indicate that although soils from the SBP site are resistant to an initial perturbation, the nature of the perturbation can make them more susceptible (i.e. have a lower functional stability) to subsequent perturbations. Similar conclusions were recently drawn from a gradient of agricultural intensification, in which intensification (initial perturbation) decreased the resistance of microbial community structure to change after a subsequent perturbation (rewetting) thereby reducing the stability of the system (Steenworth et al., 2005). Acknowledgements This work was supported by the Soil Biodiversity thematic programme from the Natural Environment Research Council. The Scottish Crop Research Institute receives grant-in-aid from the Scottish Executive Environment and Rural Affairs Department. We thank A. Horsborough (University of Aberdeen) for supplying the E. coli biosensor. References Anderson, T.H., Domsch, K.H., 1993. The metabolic quotient for CO2 (qCO2) as a specific activity parameter to assess the effects of environmental-conditions, such as pH, on the microbial biomass of forest soils. Soil Biology & Biochemistry 25, 393–395. Ball, B.C., Hunter, R., 1988. The determination of water release characteristics of soil cores at low suctions. Geoderma 43, 195–212. Bradford, M.A., Tordoff, G.M., Eggers, T., Jones, T.H., Newington, J.E., 2002. Microbiota, fauna, and mesh size interactions in litter decomposition. Oikos 99, 317–323. de Ruiter, P.C., Griffiths, B.S., Moore, J.C., 2002. Biodiversity and stability in soil ecosystems: patterns, processes and the effects of disturbance. In: Loreau, M., Naeem, S., Inchausti, P. (Eds.), Biodiversity and Ecosystem Functioning: Synthesis and Perspectives. Oxford University Press, Oxford, UK, pp. 102–114. Fitter, A.H., Gilligan, C.A., Hollingworth, K., Kleczkowski, A., Twyman, R.M., Pitchford, J.W., The members of the NERC Soil Biodiversity Programme, 2005. Biodiversity and ecosystem function in soil. Functional Ecology 19, 369. Gray, N.D., Hastings, R.C., Sheppard, S.K., Loughnane, P., Lloyd, D., McCarthy, A.J., Head, I.M., 2003. Effects of soil improvement treatments on bacterial community structure and soil processes in an upland grassland soil. FEMS Microbiology Ecology 46, 11–22. Griffiths, B.S., Hallett, P.D., Kuan, H.L., Pitkin, Y., Aitken, M.N., 2005. Biological and physical resilience of soil amended with heavy metalcontaminated sewage sludge. European Journal of Soil Science 56, 197–206. Griffiths, B.S., Kuan, H.L., Ritz, K., Glover, L.A., Fenwick, C., 2004. The Relationship between microbial community structure and functional stability, tested experimentally in an upland pasture soil. Microbial Ecology 47, 104–113. Griffiths, B.S., Ritz, K., Wheatley, R., Kuan, H.L., Boag, B., Christensen, S., Ekelund, F., Sorensen, S.J., Muller, S., Bloem, J., 2001. An examination of the biodiversity-ecosystem function relationship in arable soil microbial communities. Soil Biology & Biochemistry 33, 1713–1722. Griffiths, R.I., Whiteley, A.S., O’Donnell, A.G., Bailey, M.J., 2003. Physiological and community responses of established grassland
ARTICLE IN PRESS 2306
H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
bacterial populations to water stress. Applied and Environmental Microbiology 69, 6961–6968. Lovell, R.D., Jarvis, S.C., Bardgett, R.D., 1995. Soil microbial biomass and activity in long-term grassland: effects of management changes. Soil Biology & Biochemistry 27, 969–975. McCaig, A.E., Glover, L.A., Prosser, J.I., 2001. Numerical analysis of grassland bacterial community structure under different land management regimens by using 16S ribosomal DNA sequence data and denaturing gradient gel electrophoresis banding patterns. Applied & Environmental Microbiology 67, 4554–4559. Metcalfe, A.C., Krsek, M., Gooday, G.W., Prosser, J.I., Wellington, E.M.H., 2002. Molecular analysis of a bacterial chitinolytic community in an upland pasture. Applied & Environmental Microbiology 68, 5042–5050. Muir, J.W., 1956. The soils around Jedburgh and Morebattle. Her Majesty’s Stationery Office, Edinburgh, UK. Paton, G.I., Rattray, E.A.S., Campbell, C.D., Cresser, M.S., Glover, L.A., Meeussen, J.C.L., Killham, K., 1997. Use of genetically modified microbial biosensors for soil ecotoxicity testing. In: Pankhurst, C., Doube, B., Gupta, V. (Eds.), Biological Indicators of Soil Health and Sustainable Productivity. CAB International, pp. 397–418. Renella, G., Chaudri, A.M., Brookes, P.C., 2002. Fresh additions of heavy metals do not model long-term effects on microbial biomass and activity. Soil Biology & Biochemistry 34, 121–124. Ritz, K., McNicol, G.W., Nunan, N., Grayston, S., Millard, P., Atkinson, D., Gollotte, A., Habeshaw, D., Boag, B., Clegg, C.D., Griffiths, B.S., Wheatley, R.E., Glover, L.A., McCaig, A.E., Prosser, J.I., 2004. Spatial structure in soil chemical and microbiological properties in an upland grassland. FEMS Microbiology and Ecology 49, 191–205.
Rodwell, J.S., 1992. British Plant Communities. Cambridge University Press, Cambridge, UK. Ro¨mkens, P.F.A.M., Salomons, W., 1998. Cd, Cu and Zn solubility in arable and forest soils: Consequences of land use changes for metal mobility and risk assessment. Soil Science 163, 859–871. Shaw, J.J., Kado, C.I., 1986. Development of a Vibrio bioluminescence gene set to monitor phytopathogenic bacteria during the ongoing disease process in a nondisruptive manner. Bio-Technology 4, 560–564. Sheppard, S.K., McCarthy, A.J., Loughnane, J.P., Gray, N.D., Head, I.M., Lloyd, D., 2005. The impact of sludge amendment on methanogen community structure in an upland soil. Applied Soil Ecology 28, 147–162. Steenworth, K.L., Jackson, L.E., Caldero´n, F.J., Scow, K.M., Rolston, D.E., 2005. Response of microbial community composition and activity in agricultural and grassland soils after a simulated rainfall. Soil Biology & Biochemistry 37, 2249–2262. Vance, E.D., Brookes, P.C., Jenkinson, D.S., 1987. An extraction method for measuring soil microbial biomass-C. Soil Biology & Biochemistry 19, 703–707. Wheatley, R.E., Macdonald, R., Smith, A.M., 1989. Extraction of nitrogen from soils. Biology & Fertility of Soils 8, 189–190. Whiteley, A.S., Griffiths, R.I., Bailey, M.J., 2003. Analysis of the microbial functional diversity within water-stressed soil communities by flow cytometric analysis and CTC+ cell sorting. Journal of Microbiological Methods 54, 257–267. Yeates, G.W., Bardgett, R.D., Cook, R., Hobbs, P.J., Bowling, P.J., Potter, J.F., 1997. Faunal and microbial diversity in three Welsh grassland soils under conventional and organic management regimes. Journal of Applied Ecology 34, 453–470.
Soil Biology & Biochemistry 38 (2006) 2300–2306 www.elsevier.com/locate/soilbio
Functional resilience of microbial communities from perturbed upland grassland soils to further persistent or transient stresses H.L. Kuana, C. Fenwickb,1, L.A. Gloverb, B.S. Griffithsa,, K. Ritzc a
Environment–Plant Interactions Programme, Scottish Crop Research Institute, Invergowrie, Dundee DD2 5DA, UK b School of Medical Sciences, Institute of Medical Sciences, University of Aberdeen, Aberdeen AB25 2ZD, UK c National Soil Resources Institute, Cranfield University, Silsoe, Bedfordshire MK45 4DT, UK Received 21 October 2005; received in revised form 17 February 2006; accepted 21 February 2006 Available online 3 April 2006
Abstract The microbial functioning of soils following perturbation was assessed at a temperate upland grassland site, maintained by the Soil Biodiversity and Ecosystem Function Programme at Sourhope Research Station, Scotland. Published results indicated that the soil microbial communities were resilient to these initial perturbations; in this paper we tested whether they were equally resilient to a subsequent perturbation. Soil samples were taken from field plots receiving treatments that represented different forms of perturbation, viz. reseeding, application of sewage-sludge, biocide or nitrogen plus lime, and a non-perturbed control. Functional resilience following further perturbation comprising a transient heat or persistent copper perturbation was assessed over 28 days, by monitoring the shortterm decomposition of added plant residues. Bacterial community structure was assessed by DGGE separation of eubacterial 16S rDNA PCR products. PCR-DGGE did not distinguish any significant difference ðP40:05Þ between the bacterial communities of soils under different treatments, showing differences only between treated soils and the untreated, control soils. Two days after the application of stresses, functional capability differed in soils under different treatments. Soil samples from all the treated plots were less resilient to heat stress than soil from control plots. The initial reduction in decomposition following the addition of copper differed between treatments, but function had not recovered in any of the Cu-amended soils within 28 days. Soil resilience varied according to the type and duration of stress applied, microbial activity, soil characteristics and treatment regimes. The initial resistance of function to stress was not predictive of recovery of function over time. r 2006 Elsevier Ltd. All rights reserved. Keywords: Function; Grassland; Resilience; Soil microbial community; Stability
1. Introduction Research published to date from the major research programme, Soil Biodiversity and Ecosystem Function (SBP), funded by the UK Natural Environment Research Council has recently been reviewed (Fitter et al., 2005). To achieve the SBP’s objective, a grassland ecosystem was perturbed by removing taxa using a biocide and by subjecting discrete plots to various treatments representative of anthropogenic inputs and designed to affect Corresponding author. Tel.: +44 1382 562731; fax: +44 1382 568502.
E-mail address: bryan.griffi[email protected] (B.S. Griffiths). Present address: The Natural Environment Research Council, Polaris House, North Star Avenue, Swindon SN2 1EU, UK. 1
0038-0717/$ - see front matter r 2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.soilbio.2006.02.013
biodiversity. Many of the treatments had the expected substantial impacts on various groups of soil biota but some broad-scale processes were apparently unaffected by treatment. For example, liming had no detectable effect on the rate of soil respiration, suggesting that at this broad process-level the system was resistant to perturbation (Gray et al., 2003). Similar results were obtained by imposing specific stresses on the system, including drought, heat and the application of sewage sludge (Griffiths et al., 2003). Even in communities reconstructed by body-size exclusion techniques there were no differences in overall productivity or ecosystem carbon (C) exchange rate, suggesting that these systems are extremely resilient (Bradford et al., 2002). Although these soils appear able to retain function even when their biological structure has
ARTICLE IN PRESS H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
been radically altered, how long they can sustain this and how far we can push them remains to be seen (Fitter et al., 2005). Studies in which the biodiversity of soil from the SBP site was experimentally manipulated by perturbation showed that the resilience of organic matter decomposition in the soil was related to a specific microbial community structure rather than biodiversity per se (Griffiths et al., 2004). Another interpretation of these results is that the decomposition of organic matter is less resilient in soils that have been previously stressed (de Ruiter et al., 2002). Samples from soil amended experimentally with heavy metal contaminated sewage sludge showed that the resilience of organic matter decomposition to subsequent perturbation was affected by the previous land use management (Griffiths et al., 2005). While the evidence is that functions of soil communities at the SBP site are apparently resilient to a range of perturbations (Fitter et al., 2005; Gray et al., 2003; Griffiths et al., 2003), it is unclear whether such communities also show resilience to further perturbations. We tested the hypothesis that soil communities already subjected to an environmental stress would be less resilient to a subsequent stress. The resilience of short-term decomposition of plant material in SBP soils further subjected to heating and Cu stresses was taken as being representative of a transient and a persistent stress as used in previous studies (Griffiths et al., 2005). 2. Materials and methods
2301
were approximately at field-capacity when sampled, so were spread out and aired at 18 1C for 24 h to facilitate sieving. The litter layer, visible fauna and vegetation were removed. Soils were passed through a 4 mm metal sieve and stored field moist at 4 1C prior to analysis or assay. The soil moisture content (105 1C) and soil pH for each plot were determined. The ratios of soil volumes from different horizons present in each core were maintained in the homogenised bulk samples, e.g. if a sample consisted of 30% soil from the H-horizon and 70% from the Ahhorizon, the soils from each horizon were sieved separately and mixed together at a ratio of 3:7. Soil samples of 450 g were packed into nylon mesh-based polyethylene cylinders at a density of 1 g cm3. The packed soils were placed on a tension table (Ball and Hunter, 1988) and equilibrated to a matric potential of 50 cm water (5 kPa) for 6 days. 2.2. Soil properties Soil properties were measured at the start of the experiment, using bulked soil. Microbial biomass C was determined as the flush of C following chloroformfumigation-extraction (Vance et al., 1987). NO 3 –N, NH+ 4 –N and dissolved organic C (DOC) were measured following extraction in 1 M KCl (Wheatley et al., 1989). + NO 3 –N and NH4 –N and DOC were measured concurrently by colorimetric analysis using a segmented-flow autoanalyser (Skalar Analytical, Breda, Netherlands). Soil pH was measured in a 1:5 soil–0.01 M CaCl2 solution (w/v).
2.1. Experimental site and sampling 2.3. Soil respiration and functional stability Soil was collected from the Rigg Foot experimental site of the Soil Biodiversity and Ecosystem Function Programme at Sourhope Research Station (21 15 W, 551 30 N, UK Ordnance Survey Grid Reference NGR NT854196), in the Cheviot hills, southern Scotland. Rigg Foot is an upland grassland site, 309 m above sea level, comprising a freely drained Brown Forest soil of the Sourhope Series (Muir, 1956) with some gleying at the lower reaches. Vegetation at the site was predominantly Festuca ovinaAgrostis capillaris-Galium saxitale grassland (NVC community U4d); (Rodwell, 1992). A randomised block design was used, with three of the five replicate plots of five treatments being sampled in November 2001: untreated (control); biocide application, N and lime application; sewage sludge application; and reseeding. The N and lime plots received 600 g CaCO3 m2 y1 and 24 g NH4NO3 m2 y1. Anaerobically digested domestic sewage sludge was applied to the sewage plots at 10 l m2 in August 1999 and 20 l m2 in May 2000. Biocide-treated plots received Dursban 4 (chlorpyriphos: Dow AgroSciences, Indianapolis, USA) 1.5 l ha1 y1. The indigenous turf was removed from the reseeded plots in May 1999 and replaced with Lolium perenne L. Three replicate plots for each treatment were sampled with a metal trowel to 100 mm depth. The soils
The rate of carbon mineralisation (as measured by the short-term decomposition of plant residues) was selected as a measure of soil functional capability, based on the procedure of Griffiths et al. (2001). Composite grass cuttings (C:N ratio 20:1) were obtained from an untreated section of the Sourhope site under permanent pasture. The cuttings were dried at 60 1C and ball-milled to a fine, uniform consistency. Thirty replicate aliquots of 10 g equilibrated soil from each sample under consideration were transferred to 25 ml polypropylene scintillation vials. These soils were apportioned in batches of ten to either a heat or Cu perturbation, or a non-stressed control. The 10 heat-perturbed soils were amended with 100 ml sterile deionised water, sealed with screw-capped lids and heated at 40 1C for 18 h, cooled to 15 1C and uncapped. The ten Cu-perturbed soils were amended with CuSO4.5H2O to a concentration of 1 mg Cu g1. The ten control soils were amended with 100 ml sterile deionised water. After perturbation, all vials were incubated in a moist atmosphere at 15 1C. CO2 evolution was measured using a capillary gas chromatograph (Hewlett Packard HP5890) with a TCD detector at a GC column temperature of 90 1C. Peak areas were calculated using a SP4270 integrator (Spectra-Physics, CA, USA). Basal respiration was determined from
ARTICLE IN PRESS 2302
H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
cumulative 24 h CO2 evolution in duplicate 10 g unamended soil samples. Microbial biomass-specific respiration was calculated as the evolution of CO2 per unit of biomass C-flush (mg CO2-C mg microbial C1 h1) (after Anderson and Domsch, 1993).
the resulting data matrices analysed by principal coordinate analysis (PCO) and the resultant PCO scores analysed by ANOVA. To overcome potential bias in the results due to inconsistencies in the DGGE gels, samples were randomised in lanes across a gel, and only samples run on the same gel were compared statistically.
2.4. Bioassay of soil pore water and acute Cu toxicity An assay using the lux-marked luminescent biosensor Escherichia coli HB101 (pUCD607) (Shaw and Kado, 1986) was carried out on soil pore water extracts from the SPB plots, experimentally spiked with Cu. The assay was used to determine whether observed differences in functional stability to the copper stress were due to soil effects on copper bioavailability, or to the capacity of the microbial community to respond to added copper. Triplicate soil samples of 20 g from each of the sampled plots were either amended to 1000 mg kg1 Cu or unamended, then left for 28 d at 15 1C. The soils were diluted to 50% w/v in sterile deionised water, left to equilibrate for 24 h at 15 1C and centrifuged for 10 min at 5000 g. The supernatant was filtered through 0.7 mm GF/F glass microfibre syringe filters (Whatman International, Maidstone, UK) and frozen at 20 1C until use. The soil pore water extracts from the Cu-amended soils were diluted to 5% v/v in double-deionised water at pH 5.5 to bring the Cu load within the optimal range of microbial biosensor sensitivity. Cu standards of 0.1 mg l1 to 2 mg l1 were prepared in double deionised water from a 1111 mg l1 stock solution of Cu (3.929 g l1 CuSO4.5H2O). The biosensor was resuscitated from a freeze-dried culture and exposed to the soil solutions following the method of Paton et al. (1997). Sample luminescence was determined using a Lucy Anthos 1 microplate-reading luminometer and Stingray software (Anthos Labtec, Salzburg, Austria). Luminescence was measured after 15 minutes and expressed as relative light units (RLU), with 1 RLU being equivalent to 100 mV s1.
3. Results 3.1. Soil properties Soil pH was significantly ðPo0:05Þ elevated in the N and lime plots compared to the other treatments (Table 1). Variation within the treatments was larger than between treatments for NH+ 4 –N, NO3 –N or DOC and there were no significant differences ðP40:05Þ between the treatments (Table 1). Microbial biomass C-flush (chloroform-labile C) was significantly different ðPo0:01Þ between treatments (Table 1). Microbial biomass showed a 3-fold variation between treatments, being greatest in the sewage-amended plots and smallest in the control and reseeded plots (Po0:05; Table 1). 3.2. Soil respiration and functional stability There was no significant difference ðP40:05Þ in basal respiration or grass-induced respiration rates between treatments, but values for respiration per unit microbial biomass were significantly different ðPo0:05Þ and highest in soils from control and reseeded plots (Table 2). One day after Cu-amendment, functional stability to Cu differed significantly ðPo0:05Þ between treatments. N and lime treated soils were the most resistant to Cu, with 94% functional capability relative to a non-perturbed control and reseeded soils were least resistant to Cu, with 66% functional capability relative to a non-perturbed control (Fig. 1). Function did not recover in any of the soils 28 days after Cu application but there was a temporal
2.5. Microbial community structure DNA was extracted from a 0.5 g (fresh weight) subsample and PCR amplified as previously described by Griffiths et al. (2004). Denaturing gradient gel electrophoresis (DGGE) gels were prepared as described by McCaig et al. (2001) and samples run for 15 h at 60 V, 60 1C.
Table 1 Soil properties for Rigg Foot plots at Sourhope under different treatment regimes. Data represent mean +/ bracketed figures representing standard error ðn ¼ 3Þ
2.6. Data analysis
Control Biocide N and lime Sewage Reseed P LSD
Data were analysed by ANOVA using Genstat for Windows (version 7). For each sampling event at 1, 3, 8, 14 and 28 days after perturbation, the functional capability of the perturbed soil was determined to be the CO2 evolution of grass-amended, perturbed soils as a percentage of grassamended, non-perturbed soil. DGGE patterns were converted into presence/absence scores at each band position;
Treatment
NH+ 4 –N
pH
4.2 4.1 5.4 4.3 4.4 * 0.6
(0.1) (0.1) (0.2) (0.1) (0.1)
NO 3 –N
DOC
C-flush
mg N g1 dry wt soil
mg C g1 dry wt soil
24.1 12.2 6.6 16.6 5.7 NS 20.5
1.0 0.2 8.2 6.9 0.05 NS 12.1
(6.0) (4.6) (0.6) (3.6) (0.2)
0.9 o0.01 23.0 14.6 o0.01 NS 34.0
(0.3) (0.0) (12.9) (5.2) (0.0)
(0.6) (0.1) (2.9) (4.0) (0.01)
165 362 338 497 150 ** 155
(27) (12) (21) (43) (29)
Significant differences within a property are shown as *,**, for Po0:05 and Po0:01, respectively. LSD represents least significant differences of means at 5% level.
ARTICLE IN PRESS H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
2303
Table 2 Basal respiration, respiration induced by addition of 10 mg powdered grass g1 soil and respiration per unit of biomass C-flush in Sourhope soils under different treatment regimes Treatment
Basal mg CO2–C g1 h1
Grass-induced
Respiration rate per unit biomass C-flush mg CO2–C g1 C-flush h1
Control Biocide N and lime Sewage Reseed P LSD
1.27 1.06 1.22 1.03 1.68 NS 0.67
12.12 9.68 17.11 13.14 11.67 NS 5.46
0.0085 0.0030 0.0036 0.0022 0.0143 * 0.0082
(0.14) (0.06) (0.07) (0.05) (0.21)
(1.83) (0.57) (0.72) (0.61) (0.68)
(0.0034) (0.0007) (0.0001) (0.0005) (0.0095)
Data represent mean +/ bracketed figures representing standard error ðn ¼ 3Þ. Significant differences are shown as *, for Po0:05. LSD represents least significant differences of means at 5% level.
Fig. 1. The percentage change in respiration (CO2 evolution over 24 h) from powdered grass added to Sourhope soils under different land management regimes, 1, 3, 8, 14, 28 days after application of 1 mg Cu g1. Bars represent means +/SE ðn ¼ 3Þ.
Fig. 2. The percentage change in respiration (CO2 evolution over 24 h) from powdered grass added to Sourhope soils under different treatments, 1, 3, 8, 14 and 28 days after perturbation by 40 1C for 18 h. Bars represent means +/SE ðn ¼ 3Þ.
convergence in the relative respiration rates of Cu-amended soils. At day 28 there was no significant difference ðP40:05Þ in the function of Cu-amended soils from different treatments, with a mean functional capability of 71%. The functional stability of soils after heat perturba-
tion differed according to soil treatment ðPo0:05Þ. One day after heating, soils from the control plots displayed the initial greatest resistance to heat perturbation with 71% of original capability (Fig. 2) and soils from the Reseeded plots were least resistant to heat with 42% of the original
ARTICLE IN PRESS H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
2304
capability. Recovery of function occurred in all soils over time, with a mean of 84% original capability at day 28. No significant difference ðPo0:05Þ was seen in the functional capability of heat-perturbed soils from plots under different treatments at Day 28. 3.3. Soil pore water and acute Cu toxicity E. coli HB101 (pUCD607) luminescence after exposure to unamended soil pore water extracts did not differ ðP40:05Þ between plot treatments (data not shown). Exposure to Cu-amended soil pore water from biocide, control, reseed and sewage plots resulted in a significant ðPo0:01Þ inhibition of biosensor luminescence relative to unamended samples (Fig. 3). There was a significant ðPo0:05Þ interaction between plot treatments and Cu amendment such that extracts from N and lime soils amended with Cu did not result in further inhibition of
% luminescence (RLU)
125 100 75 50 25
Control
Biocide
Reseeded
Sewage
N + lime
0 0
5 10 Exposure time (minutes)
15
Fig. 3. The inhibition of E.coli HB101 (pUCD607) luminescence over a 15 minute exposure to soil pore water extracts from soils amended with 1 mg g1 Cu from Sourhope plots, as % of luminescence relative to a deionised water control. Data represent mean +/SE Figures in brackets represent SE ðn ¼ 6Þ.
PCO 2 (13% of variation)
0.50
Control N+Lime Biocide Reseeded Sewage
0.25
0.00
-0.25
-0.50 -0.4
-0.2
0.0 0.2 PCO 1 (17% of variation)
0.4
Fig. 4. Plot of first and second principal co-ordinates of PCR-DGGE profiles obtained from Sourhope soils under different treatments, using eubacterial primers. Points show means (n ¼ 3; bars denote +/SE).
biosensor luminescence relative to unamended soil from the same plots. 3.4. Microbial community structure PCO analysis of PCR-DGGE profiles revealed an intrinsically high degree of similarity between eubacterial community structures in all soils. The first principal coordinate did not distinguish between any treatments, and the second co-ordinate significantly distinguished the N and lime, reseeded and sewage treatments from the control (Po0:05; Fig. 4). The community structure in the biocidetreated soils was indistinguishable from the other treatments (Fig. 4). 4. Discussion Microbial biomass C-flush and soil pH differed according to land management practice. The largest microbial biomass C-flush was observed in sewage plots, suggesting that inputs of organic C and N from sewage application supported a growth in the soil microbial biomass. Microbial biomass C-flush was significantly higher in the sewage plots, N and lime plots and biocide-treated plots than in soils from control plots. This contrasts with the results of long-term studies of temperate grasslands, where soil microbial biomass size and activity were lower in Namended areas than in unfertilised areas (Lovell et al., 1995;Yeates et al., 1997). Rates of basal respiration did not differ in soils from plots under different management regimes. This was consistent with other studies at the site (Gray et al., 2003). Similarly, rates of grass-induced respiration did not differ in soils from plots under different management regimes. Also, no significant differences were observed in concentrations of NH+ 4 –N, NO3 –N or DOC in the soils under different treatments. However, variation between plots under the same treatment was larger than variation between treatments, reflecting the heterogeneity of the site (Ritz et al., 2004). Analysis of the genetic structure of the eubacterial bacterial communities in these soils showed a significant but weak trend that the community structures from treated plots differed from that of untreated soils. This agrees closely with the results of Gray et al. (2003) who analysed the eubacterial community using TGGE and found significant but only slight, effects of the treatments. Whiteley et al. (2003) could not detect changes in eubacterial community structure in the same soil following wet-dry cycles. Studies of a different field at the same site also indicated that DGGE profiling of eubacterial DNA was generally unaffected by environmental heterogeneity (Ritz et al., 2004). Analysis of more defined taxonomic and functional groups, such as ammonia oxidisers (Gray et al., 2003), methanogens (Sheppard et al., 2005) and chitin degraders (Metcalfe et al., 2002) were able to detect changes in microbial community structure induced by the field treatments. One of the reasons for choosing the
ARTICLE IN PRESS H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
decomposition of plant residues in the assay was that it is a general function carried out by a wide range of organisms (Griffiths et al., 2000), so it would be difficult to relate changes in functionally defined microbial groups to decomposition. Our results suggest that the observed variations in functional resilience to heat and copper perturbation in soils under different treatments were not associated with changes in broad-scale microbial community structure. The microbial biosensor E. coli HB101 (pUCD607) did not distinguish between soil pore water extracts from the plots under different treatments. This suggests that differences in functional resilience were not due to soil treatments at the SBP site (e.g. sewage sludge) exerting a direct toxic effect on the soil microbial community. However, biosensor exposure to Cu-amended soil pore water extracts resulted in inhibition of luminescence relative to the unamended extracts, with the exception of Cu-amended soil pore water from the N and lime plots. The reduced Cu bioavailability in the extracts from the N and lime soils would be associated with the elevated pH due to liming treatments (Ro¨mkens and Salomons, 1998). Soils under different land management regimes varied in their functional response to Cu perturbation, but were not resilient to Cu, as measured by recovery of grassdecomposition capability over 28 days. Initially, grassdecomposition capability in soil from the N and lime and control plots was most resistant to Cu; this corresponds with the results of the Cu bioavailability study. Functional capability decreased over time in Cu-perturbed N and lime and control soils, so that by day 28 there was no significant difference (P40.05) between the functional capability in any of the soils. This temporal convergence in functional capability suggests that certain soil properties have an initial protective effect on soil microbial community function in the presence of Cu, but these factors may only offer short-term protection. Measuring the long-term resilience of microbial function may therefore be more informative than acute toxicity assays in assessing the effects of persistent perturbations such as heavy metals. A study by Renella et al. (2002) confirmed that fresh additions of heavy metals to experimental systems are not predictive of long-term effects on microbial activity. Functional responses to heat perturbation differed according to the land management practices applied to the plots. The initial resistance to heat (as measured by decreases in functional capability one day after heating) was highest in soils from the control plots and lowest in the reseeded soils. The reseeded soils were most resilient to heat (i.e. function recovered most over 28 days, but this is also related to these soils having the greatest initial reduction in function after heat) and the sewage-amended soils were least resilient to heat over a 28-day period. The underlying mechanisms of these differences are unclear and may be related to differences in soil conditions under different treatments (e.g. pH, organic matter content, physical structure, plant community composition and productivity).
2305
5. Conclusions The results indicate that although soils from the SBP site are resistant to an initial perturbation, the nature of the perturbation can make them more susceptible (i.e. have a lower functional stability) to subsequent perturbations. Similar conclusions were recently drawn from a gradient of agricultural intensification, in which intensification (initial perturbation) decreased the resistance of microbial community structure to change after a subsequent perturbation (rewetting) thereby reducing the stability of the system (Steenworth et al., 2005). Acknowledgements This work was supported by the Soil Biodiversity thematic programme from the Natural Environment Research Council. The Scottish Crop Research Institute receives grant-in-aid from the Scottish Executive Environment and Rural Affairs Department. We thank A. Horsborough (University of Aberdeen) for supplying the E. coli biosensor. References Anderson, T.H., Domsch, K.H., 1993. The metabolic quotient for CO2 (qCO2) as a specific activity parameter to assess the effects of environmental-conditions, such as pH, on the microbial biomass of forest soils. Soil Biology & Biochemistry 25, 393–395. Ball, B.C., Hunter, R., 1988. The determination of water release characteristics of soil cores at low suctions. Geoderma 43, 195–212. Bradford, M.A., Tordoff, G.M., Eggers, T., Jones, T.H., Newington, J.E., 2002. Microbiota, fauna, and mesh size interactions in litter decomposition. Oikos 99, 317–323. de Ruiter, P.C., Griffiths, B.S., Moore, J.C., 2002. Biodiversity and stability in soil ecosystems: patterns, processes and the effects of disturbance. In: Loreau, M., Naeem, S., Inchausti, P. (Eds.), Biodiversity and Ecosystem Functioning: Synthesis and Perspectives. Oxford University Press, Oxford, UK, pp. 102–114. Fitter, A.H., Gilligan, C.A., Hollingworth, K., Kleczkowski, A., Twyman, R.M., Pitchford, J.W., The members of the NERC Soil Biodiversity Programme, 2005. Biodiversity and ecosystem function in soil. Functional Ecology 19, 369. Gray, N.D., Hastings, R.C., Sheppard, S.K., Loughnane, P., Lloyd, D., McCarthy, A.J., Head, I.M., 2003. Effects of soil improvement treatments on bacterial community structure and soil processes in an upland grassland soil. FEMS Microbiology Ecology 46, 11–22. Griffiths, B.S., Hallett, P.D., Kuan, H.L., Pitkin, Y., Aitken, M.N., 2005. Biological and physical resilience of soil amended with heavy metalcontaminated sewage sludge. European Journal of Soil Science 56, 197–206. Griffiths, B.S., Kuan, H.L., Ritz, K., Glover, L.A., Fenwick, C., 2004. The Relationship between microbial community structure and functional stability, tested experimentally in an upland pasture soil. Microbial Ecology 47, 104–113. Griffiths, B.S., Ritz, K., Wheatley, R., Kuan, H.L., Boag, B., Christensen, S., Ekelund, F., Sorensen, S.J., Muller, S., Bloem, J., 2001. An examination of the biodiversity-ecosystem function relationship in arable soil microbial communities. Soil Biology & Biochemistry 33, 1713–1722. Griffiths, R.I., Whiteley, A.S., O’Donnell, A.G., Bailey, M.J., 2003. Physiological and community responses of established grassland
ARTICLE IN PRESS 2306
H.L. Kuan et al. / Soil Biology & Biochemistry 38 (2006) 2300–2306
bacterial populations to water stress. Applied and Environmental Microbiology 69, 6961–6968. Lovell, R.D., Jarvis, S.C., Bardgett, R.D., 1995. Soil microbial biomass and activity in long-term grassland: effects of management changes. Soil Biology & Biochemistry 27, 969–975. McCaig, A.E., Glover, L.A., Prosser, J.I., 2001. Numerical analysis of grassland bacterial community structure under different land management regimens by using 16S ribosomal DNA sequence data and denaturing gradient gel electrophoresis banding patterns. Applied & Environmental Microbiology 67, 4554–4559. Metcalfe, A.C., Krsek, M., Gooday, G.W., Prosser, J.I., Wellington, E.M.H., 2002. Molecular analysis of a bacterial chitinolytic community in an upland pasture. Applied & Environmental Microbiology 68, 5042–5050. Muir, J.W., 1956. The soils around Jedburgh and Morebattle. Her Majesty’s Stationery Office, Edinburgh, UK. Paton, G.I., Rattray, E.A.S., Campbell, C.D., Cresser, M.S., Glover, L.A., Meeussen, J.C.L., Killham, K., 1997. Use of genetically modified microbial biosensors for soil ecotoxicity testing. In: Pankhurst, C., Doube, B., Gupta, V. (Eds.), Biological Indicators of Soil Health and Sustainable Productivity. CAB International, pp. 397–418. Renella, G., Chaudri, A.M., Brookes, P.C., 2002. Fresh additions of heavy metals do not model long-term effects on microbial biomass and activity. Soil Biology & Biochemistry 34, 121–124. Ritz, K., McNicol, G.W., Nunan, N., Grayston, S., Millard, P., Atkinson, D., Gollotte, A., Habeshaw, D., Boag, B., Clegg, C.D., Griffiths, B.S., Wheatley, R.E., Glover, L.A., McCaig, A.E., Prosser, J.I., 2004. Spatial structure in soil chemical and microbiological properties in an upland grassland. FEMS Microbiology and Ecology 49, 191–205.
Rodwell, J.S., 1992. British Plant Communities. Cambridge University Press, Cambridge, UK. Ro¨mkens, P.F.A.M., Salomons, W., 1998. Cd, Cu and Zn solubility in arable and forest soils: Consequences of land use changes for metal mobility and risk assessment. Soil Science 163, 859–871. Shaw, J.J., Kado, C.I., 1986. Development of a Vibrio bioluminescence gene set to monitor phytopathogenic bacteria during the ongoing disease process in a nondisruptive manner. Bio-Technology 4, 560–564. Sheppard, S.K., McCarthy, A.J., Loughnane, J.P., Gray, N.D., Head, I.M., Lloyd, D., 2005. The impact of sludge amendment on methanogen community structure in an upland soil. Applied Soil Ecology 28, 147–162. Steenworth, K.L., Jackson, L.E., Caldero´n, F.J., Scow, K.M., Rolston, D.E., 2005. Response of microbial community composition and activity in agricultural and grassland soils after a simulated rainfall. Soil Biology & Biochemistry 37, 2249–2262. Vance, E.D., Brookes, P.C., Jenkinson, D.S., 1987. An extraction method for measuring soil microbial biomass-C. Soil Biology & Biochemistry 19, 703–707. Wheatley, R.E., Macdonald, R., Smith, A.M., 1989. Extraction of nitrogen from soils. Biology & Fertility of Soils 8, 189–190. Whiteley, A.S., Griffiths, R.I., Bailey, M.J., 2003. Analysis of the microbial functional diversity within water-stressed soil communities by flow cytometric analysis and CTC+ cell sorting. Journal of Microbiological Methods 54, 257–267. Yeates, G.W., Bardgett, R.D., Cook, R., Hobbs, P.J., Bowling, P.J., Potter, J.F., 1997. Faunal and microbial diversity in three Welsh grassland soils under conventional and organic management regimes. Journal of Applied Ecology 34, 453–470.