Published online 31 August 2007
Published in J Environ Qual 36:1403-1411 (2007)
DOI: 10.2134/jeq2006.0471
© 2007 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
677 S. Segoe Rd., Madison, WI 53711 USA
Differential Responses of Eubacterial, Mycobacterium, and Sphingomonas Communities in Polycyclic Aromatic Hydrocarbon (PAH)-Contaminated Soil to Artificially Induced Changes in PAH Profile
Maarten Uyttebroeka,
Astrid Spodena,
Jose-Julio Ortega-Calvob,
Katinka Woutersa,
Pierre Wattiauc,
Leen Bastiaensd and
Dirk Springaela,*
a Division of Soil and Water Management, Catholic Univ. of Leuven, Kasteelpark Arenberg 20, B-3001 Leuven, Belgium
b Instituto de Recursos Naturales y Agrobiología de Sevilla, CSIC, Avenida Reina Mercedes 10, Apartado 1052, E-41080 Seville, Spain
c Bacteriology and Immunology, Veterinary and Agrochemical Research Centre (VAR), Groeselenberg 99, B-1180 Brussels, Belgium
c Environmental and Process Technology, Flemish Institute for Technological Research (VITO), Boeretang 200, B-2400 Mol, Belgium
* Corresponding author (dirk.springael{at}biw.kuleuven.be).
Received for publication October 31, 2006.
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ABSTRACT
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Recent reports suggest that Mycobacterium is better adapted to soils containing poorly bioavailable polycyclic aromatic hydrocarbons (PAHs) compared to Sphingomonas. To study this hypothesis, artificial conditions regarding PAH profile and PAH bioavailability were induced in two PAH-contaminated soils and the response of the eubacterial, Mycobacterium, and Sphingomonas communities to these changed conditions was monitored during laboratory incubation. Soil K3663 with a relatively high proportion of high molecular weight PAHs was amended with phenanthrene or pyrene to artificially change the soil into a soil with a relatively increased bioavailable PAH contamination. Soil AndE with a relatively high proportion of bioavailable low molecular weight PAHs was treated by a single-step Tenax extraction to remove the largest part of the easily bioavailable PAH contamination. In soil K3663, the added phenanthrene or pyrene compounds were rapidly degraded, concomitant with a significant increase in the number of phenanthrene and pyrene degraders, and minor and no changes in the Mycobacterium community and Sphingomonas community, respectively. However, a transient change in the eubacterial community related to the proliferation of several 
-proteobacteria was noted in the phenanthrene-amended soil. In the extracted AndE soil, the Sphingomonas community initially developed into a more diverse community but finally decreased in size below the detection limit. Mycobacterium in that soil never increased to a detectable size, while the eubacterial community became dominated by a -proteobacterial population. The results suggest that the relative bioavailability of PAH contamination in soil affects bacterial community structure but that the behavior of Mycobacterium and Sphingomonas in soil is more complex than prospected from studies on their ecology and physiology.
Abbreviations: HMW, high molecular weight • PAH, polycyclic aromatic hydrocarbon • LMW, low molecular weight • MPN, most probable number
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INTRODUCTION
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POLYCYCLIC AROMATIC HYDROCARBONS (PAHs) are hydrophobic environmental contaminants that are found in high concentration at sites with former manufactured gas plants (MGPs) and with wood treatment activity (Cerniglia, 1992). The recalcitrance of high molecular weight (HMW) PAHs, i.e., PAHs with more than three rings, in the environment is higher compared to low molecular weight (LMW) PAHs, i.e., PAHs with less than four rings (Cerniglia, 1992; Miller et al., 1985). After the introduction of PAHs in soil, the LMW PAH fraction is preferentially removed both by physicochemical and biological processes, while remaining PAHs (both LMW and HMW) become progressively less accessible due to sorption to organic matter, partitioning into nonaqeous phase liquids (NAPLs) and sequestration in soil micropores, resulting in poor bioavailability (Alexander, 2000). Thus, depending on the environmental conditions, a PAH-contaminated soil will display, in time, different weathering stages, characterized by dissimilar PAH profiles and bioavailability.
Despite the low PAH bioavailability, bacteria able to use PAHs as the sole source of carbon and energy can be isolated from such soils. These isolates belong to a limited number of genera such as Pseudomonas, Burkholderia, Sphingomonas, and Mycobacterium, in which especially Sphingomonas and Mycobacterium are of interest. These genera are frequently isolated from contaminated soils as the main culturable PAH degraders and can degrade various PAH compounds. However, some main differences exist between both genera. Sphingomonads grow on compounds like phenanthrene and fluoranthene while mycobacteria are, in addition, able to use pyrene as sole source of carbon and energy (Kanaly and Harayama, 2000; Wick et al., 2001). Furthermore, several studies indicate that Mycobacterium is well adapted to degrade sorbed PAHs, while Sphingomonas and Burkholderia are more adapted to degrade readily available PAHs (Bastiaens et al., 2000; Friedrich et al., 2000; Grosser et al., 2000; Uyttebroek et al., 2006b). In addition, Mycobacterium shows several physiological adaptations to low PAH bioavailability and oligotrophy (Wick et al., 2001; Wattiau et al., 2002; Wick et al., 2002, 2003). On the basis of these studies, it can be hypothesized that Mycobacterium might have selective advantages in a soil containing a less bioavailable PAH, enriched in more recalcitrant PAHs, while Sphingomonas is more competitive in case of a more recent contamination, i.e., in soils with high phenanthrene concentrations. A similar hypothesis can be deduced from recent studies on the Mycobacterium and Sphingomonas community structure in PAH-contaminated soils with different degrees of weathering. Leys et al. (2005), using a culture-independent approach, found that Mycobacterium was prominent in weathered soils containing mainly poorly bioavailable PAHs, while in soils with high concentrations of bioavailable PAHs like phenanthrene, Mycobacterium could not be detected. In some of the weathered soils, PAH-degrading Mycobacterium belonged indeed to the majority of culturable PAH-degrading bacteria, while this was not the case in the soils with high PAH concentrations (Uyttebroek et al., 2006a; Fredslund and Springael, unpublished results, 2006). Leys et al. (2004), using a culture-independent approach, showed the occurrence of Sphingomonas in all of those soils, but soils with mainly HMW PAHs displayed a higher Sphingomonas diversity than soils with high concentrations of bioavailable PAHs, indicating selection of specific Sphingomonas species in the latter soils. Using a culture-dependent approach, Vanbroekhoven et al. (2004) showed that actual phenanthrene-degrading Sphingomonas could only be isolated from the soils with high PAH concentrations, while Sphingomonas strains isolated from the weathered soils did not degrade phenanthrene. These studies indicate that the PAH profile and hence the stage of weathering of a PAH contamination in soil might influence bacterial community structure including the communities of PAH-degrading bacterial specialists such as Mycobacterium and Sphingomonas.
In this study, we assessed the hypothesis that Mycobacterium and Sphingomonas occupy different niches in PAH-contaminated soils and that these niches depend on the degree of weathering of the contamination, i.e., the bioavailability, composition, and concentration of the PAHs. Therefore, in two soils contaminated with PAH with apparent different degree of weathering, we artificially created conditions, opposite to the initial conditions regarding profile and bioavailability of PAHs. Soil K3663, a soil with a relatively high proportion of HMW PAHs was amended with phenanthrene or pyrene to artificially change the soil into a soil with a relatively more bioavailable PAH contamination. Soil AndE, a soil with a relatively high proportion of bioavailable LMW PAHs was treated by a single-step Tenax extraction to remove the largest part of the easily bioavailable PAH contamination. Subsequently, the soils were incubated and the eubacterial, Mycobacterium, and Sphingomonas communities were followed in time using specific polymerase chain reaction-denaturing gel electrophoresis (PCR-DGGE).
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Materials and Methods
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Soil Samples
Soil K3663, derived from a PAH-contaminated site of a former coal gasification plant in Denmark, is a loamy sandy soil with a relatively weathered PAH contamination. It contains mainly poorly bioavailable HMW PAHs as prolonged incubation did only remove 5 to 10% of the contamination (Uyttebroek et al., unpublished data, 2006). Soil K3663 contains a large number of PAH-degrading Mycobacterium (Uyttebroek et al., 2006a). Soil AndE, originating from a PAH-contaminated railway station site in Spain, is a heavy clay soil. It is characterized by high concentrations of bioavailable PAHs (especially phenanthrene) as incubation of the soil removed rapidly 60 to 80% of the contamination (Bueno-Montes, 2005). The soil harbors a large number of phenanthrene-degrading Sphingomonas (Vanbroekhoven et al., 2004) but low numbers of Mycobacterium (Leys et al., 2005). The soils were sieved at 2 mm and stored fresh at 4°C in the dark until use. The physicochemical soil characteristics and the PAH concentrations in soil (Table 1) were determined using published methods as described by Leys et al. (2005). The total PAH concentration is the sum of the concentrations of the 16 PAHs, listed by the USEPA.
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Table 1. Physicochemical characteristics of soil samples K3663 and AndE before artificial changes in concentration, structure and bioavailability of polycyclic aromatic hydrocarbons (PAHs).
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Setup of Soil Microcosm Experiments
Soil K3663 microcosms consisted of 100 g soil incubated in 250 mL Erlenmeyer flasks, placed into a closed jar with a few mL of distilled water to keep the air water-saturated. One set of duplicate microcosms contained soil spiked at a final concentration of 1000 mg kg–1 phenanthrene (dry weight), one set contained soil spiked with 500 mg kg–1 pyrene, and one set received only acetone. In addition, to examine abiotic losses of phenanthrene or pyrene, two sets of duplicate microcosms containing either phenanthrene- or pyrene-spiked sterilized soil K3663 were included. The soil was sterilized before spiking by autoclaving at 0.1 MPa and 121°C for 1 h with an in-between incubation period of 2 d at 25°C, as described by Wolf and Skipper (1994). The soils were spiked as described by Brinch et al. (2002). In a 500 mL Erlenmeyer flask, 10 mL of a PAH solution in acetone (20 g L–1 phenanthrene or 10 g L–1 pyrene) or 10 mL of acetone alone was added to 50 g (dry weight) of soil. The soil and the acetone solution were mixed for 5 min with a metal spatula, the flask was closed for 5 min, and opened for 16 h for evaporation of the acetone. Afterward, the spiked soil was mixed with 150 g (dry weight) of non-spiked K3663 soil for 5 min with a metal spatula. Finally, 100 g (dry weight) of the mixture was incubated in each soil microcosm. In all microcosms, the moisture content of the soils was set at 20% with sterile distilled water. Regularly, the jars were opened to aerate and to compensate for water loss with sterile distilled water. The microcosms were incubated for 132 d at 25°C in the dark. Separate individual samples for most-probable-number enumeration of PAH degraders (2 g), PAH analysis (6 g), or PCR-DGGE analysis (0.8 g) were taken from each microcosm after 0, 13, 29, 55, and 132 d of incubation.
Microcosms with soil AndE were set up in duplicate with 30 g (dry weight) of non-extracted or Tenax-extracted soil AndE in 250 mL Erlenmeyer flasks as described for the K3663 soil. Tenax extraction removes the fraction of a PAH contamination which corresponds to the rapidly desorbing and hence easily bioavailable PAH fraction in soil (Cornelissen et al., 2001) and was performed according to Cornelissen et al. (1998). Tenax TA (60–80 mesh, 177–250 µm), a porous polymer based on 2,6-diphenyl-p-phenylene oxide, was purchased from Teknokroma (Barcelona, Spain). Before use, the Tenax beads were washed as described by Cornelissen et al. (1998) and dried overnight at 70°C. A mixture of 1.5 g Tenax, 1 g soil (dry weight), and 70 mL milli-Q water was shaken for 24 h on a rotary shaker at 100 rpm at 20°C in a 250 mL separation funnel. The soil suspension was separated from the Tenax and remaining soil particles were washed twice from the Tenax beads with 15 mL milli-Q water. The soil was recovered by centrifugation at 16300 x g for 15 min. Soil from different extractions was pooled and stored at 4°C for 19 d. Before use, the extracted soil was dried at 25°C until it had a moisture content of 20%. Finally, the extracted soil was divided over two 250 mL Erlenmeyer flasks. The moisture content of the non-extracted and the Tenax-extracted soil in the microcosms flasks was finally set at 40% with distilled water and kept constant during the incubation experiment. Incubation was done at 25°C in the dark. One soil sample (0.4 g) for PCR-DGGE analysis was taken from each microcosm after 0, 21, 35, 63, 126, and 252 d of incubation.
Polycyclic Aromatic Hydrocarbon Analysis
Most samples from soils K3663 were extracted with hexane-acetone and the PAH concentration was measured by HPLC. Six grams of soil sampled from each microcosm was divided into six 1-g subsamples and each subsample was analyzed separately. The 1-g subsamples were extracted twice with 5 mL hexane-acetone (80:20 v/v) by vortexing at 2800 rpm for 2 min and centrifugation at 1700 x g for 3 min. Phenanthrene and pyrene concentrations in the acetone-hexane fraction were measured by high performance liquid chromatography (HPLC) (Merck Hitachi LaChrom D-7000, Tokyo, Japan), equipped with a 5 µm PurospherSTAR RP-18 endcapped column (Merck, LiChroCART 125 x 4 mm) and a UV-VIS detector (Merck Hitachi LaChrom L-7420, Tokyo, Japan) set at 254 nm. The mobile phase was acetonitrile-water (75:25 v/v) and the flow rate was 1 mL min–1. K3663 soil samples taken after 206 d of incubation and all AndE soil samples were extracted by accelerated solvent extraction (ASE 200, Dionex, Sunnyvale, CA, USA) with hexane-tetrahydrofuran (80:20 v/v), as described by Leys et al. (2005). The PAH concentration was measured by gas chromatography (GC, MFC 500, Carlo Erba Instruments, Milan, Italy), coupled to a mass-spectrophotometric detector (MS, QMD 100, Fisons Instruments, Loughborough, UK), as described by Leys et al. (2005).
Most Probable Number (MPN) Enumeration of Phenanthrene and Pyrene Degraders in Soil K3663
Bacteria were extracted from the soil samples by mixing 2 g of fresh sample and 18 mL of 10–2 M MgSO4 in a 100 mL Erlenmeyer flask and shaking the soil suspension on a rotary shaker at 125 rpm for 2 h at 25°C. The suspension was allowed to settle by gravity for 1 h at room temperature. A tenfold dilution series of the supernatant (to 10–8) was prepared in duplicate in 10–2 M MgSO4 and three drops (each 10 µL) of each dilution were put on a solid agar plate (three-row MPN) containing 15 g L–1 agar (Select Agar, Invitrogen, Merelbeke, Belgium). PAH degraders were counted on plates with phosphate-buffered minimal medium (MM) (pH 7.5), as described by Uyttebroek et al. (2006a). The plates with MM contained an opaque layer of PAH provided as sole source of carbon and energy. The PAH layer was prepared by spreading 1 mL acetone solution of phenanthrene (Fluka, Buchs, Switzerland; purity >97%) or pyrene (Fluka, Buchs, Switzerland; purity >97%) (4 mg mL–1) on the agar surface and evaporation of the acetone. MPN enumeration of phenanthrene or pyrene degraders was based on the formation of clearing zones in the opaque phenanthrene or pyrene layer, respectively. The numbers of positive drops were counted for each sample. Total heterotrophic bacteria were counted on 10–1 LB (Luria-Bertani) plates by the MPN technique as described above, but based on formation of colony forming units (CFU). The 10–1 LB medium (pH 7.0) contained (per liter) 1 g tryptone, 0.5 g yeast extract, and 5 g NaCl. The MPN was calculated by using Microsoft Excel, as described by Briones and Reichardt (1999).
Molecular Techniques
DNA was extracted from the soil samples as described previously (Uyttebroek et al., 2006a). Eubacterial 16S rRNA gene fragments were amplified by PCR from the DNA extract using the eubacterial primers GC40-63f and 518r (El Fantroussi et al., 1999), as described by Moreels et al. (2004). Sphingomonas and Mycobacterium 16S rRNA gene fragments were amplified using primer combinations Sphingo108f/GC40-Sphingo420r and MYCO66f/GC40-MYCO600r, respectively, as described by Leys et al. (2004, 2005). Limits of detection of the PCR amplification of Sphingomonas and Mycobacterium 16S rRNA gene in soil were previously determined as 104 and 106 CFU g–1 soil, respectively (Leys et al., 2004, 2005). The PCR amplification was performed on an Eppendorf Mastercycler and contained 1 µL soil DNA as a template. DGGE was performed on an INGENYphorU-2 system (Ingeny International, Goes, The Netherlands). DGGE analysis of the eubacterial PCR products was done as described by Moreels et al. (2004), but for 15 h at 120 V. DGGE analysis of the Sphingomonas and the Mycobacterium PCR products was done as described by Leys et al. (2004, 2005). Bands from eubacterial, Sphingomonas, and Mycobacterium 16S rRNA gene-based DGGE profiles were cloned into the plasmid vector pCR2.1-TOPO, using the TOPO cloning kit (Invitrogen, Merelbeke, Belgium) as described previously (Leys et al., 2004, 2005). DGGE patterns of cloned fragments were compared with the appropriate fingerprints of the soil eubacterial, Mycobacterium, or Sphingomonas community and appropriate clones were chosen for sequence analysis. Sequencing reactions were performed with the QuickStart DNA sequencing kit (Beckman, Boston, USA) using the eubacterial 16S rRNA gene primer 530r (Lane, 1991) for eubacteria and Mycobacterium, and the Sphingomonas-specific primer Sphingo108f for Sphingomonas, and analyzed on an automatic sequencer (CEQ 8000, Beckman-Coulter, Fullerton, CA, USA). Resulting partial 16S rRNA gene sequences (about 450 bp for eubacteria and Mycobacterium and about 300 bp for Sphingomonas) were compared to sequences deposited in GenBank by BLASTN search (Altschul et al., 1997). The partial 16S rRNA gene sequence data have been submitted to the EMBL database under accession numbers AM085776 to AM085799.
Statistical Analysis
The Tukey test (SAS 9.1; Cary, NC, USA) (P < 0.05) was used for all statistical comparisons of log-transformed MPN per g soil between soil treatments.
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Results
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K3663 Soil Experiment
The dynamics of phenanthrene and pyrene concentrations in the soils during 132 d of incubation are shown in Fig. 1
. Rapid degradation of freshly added PAHs was observed in the non-sterilized soils during the first 29 d, which was clearly due to biodegradation since the PAH concentration in the sterilized soils remained constant during that period. However, a decrease of the PAH concentrations in the sterilized soils was observed from Day 55. This was probably due to the development of a PAH-degrading community that had resisted the sterilization procedure since large numbers of phenanthrene and pyrene degraders were found in the sterilized soils after 132 d of incubation (data not shown). To avoid biases related to the extraction protocol, all soils were extracted after 206 d of incubation with the accelerated solvent extraction method. The non-sterilized phenanthrene- and pyrene-spiked soils contained 16 mg kg–1 phenanthrene and 27 mg kg–1 pyrene, respectively, while the sterilized phenanthrene and pyrene spiked soils contained 28 mg kg–1 phenanthrene and 65 mg kg–1 pyrene, respectively. These results indicate that the observed PAH loss in all soils was indeed due to PAH biodegradation. It was also observed that the incubation resulted in a decrease in the concentration of original PAHs in the K3663 soil, as deduced from the PAH analysis results from the control soil and the acetone-treated soil after incubation. In both soils phenanthrene, pyrene, LMW PAH, and HMW PAHs were around 1.5, 7.2, 2.2, and 57.1 mg kg–1 instead of 2, 10, 4, and 69 mg kg–1 in the original soil, indicating that there was still a fraction of the original PAHs in the K3663 soil that was available for biodegradation.
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Fig. 1. Time lapse phenanthrene or pyrene concentrations in non-sterile or sterile soil K3663, spiked with phenanthrene (phe) or pyrene (pyr), respectively, during the K3663 incubation experiment. Error bars represent one standard error from six replicates.
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The dynamics of total heterotrophic bacteria and phenanthrene and pyrene degraders during 132 d of incubation are shown in Table 2. Both the number of total heterotrophs and PAH degraders were not significantly different between the non-treated and the acetone-treated control soils at each sampling time, except for the number of pyrene degraders at Day 29. It shows that the application of acetone as spiking solvent did not have a significant effect on culturable community structure. At Day 13, compared to the acetone control, the number of both phenanthrene and pyrene degraders was significantly increased as well in the soil spiked with phenanthrene (1.8 log units), as in the soil spiked with pyrene (1.3 log units). The increase of PAH degraders was also reflected in a significant increase (1.4 log units) of total heterotrophs, compared to the acetone control, in the soil spiked with phenanthrene at Day 13, and in the soil spiked with pyrene at Day 29. In all cases, the increases in number were followed by a gradual decrease during further incubation down to numbers similar to those in the control treatments after 132 d of incubation.
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Table 2. Most probable number (MPN) enumeration of heterotrophic bacteria, phenanthrene degraders, and pyrene degraders after 0, 13, 29, 55, and 132 d of incubation for the K3663 soil experiment.
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The community dynamics during the incubation are shown in Fig. 2
for eubacteria and Mycobacterium. The Sphingomonas community composition in the control K3663 soils and the K3773 soils spiked with PAHs were identical and did not change during the incubation experiment (data not shown). The eubacterial community in the control soil and the acetone control soil did not change during the incubation of 132 d, indicating no detectable influence of incubation conditions and of the acetone spiking solvent on the eubacterial populations (data not shown). The eubacterial community in the soil spiked with pyrene remained identical to the community in the control soils and did not change during the 132 d of incubation (data not shown). On the other hand, the eubacterial community in the soil spiked with phenanthrene showed a transient change, in contrast with the control soils. At Days 13 and 29, additional dominant bands such as K1, K2, and especially K8 appeared in the DGGE fingerprints of the phenanthrene-spiked soil, but disappeared again at Day 132. The rest of the banding pattern did not change during the incubation. 16S rRNA gene sequences corresponding to bands K1, K2, and K8 were most related to the 16S rRNA gene sequences of a Pseudomonas sp. and two uncultured -proteobacteria, respectively (Table 3). The Mycobacterium community structures in the control soils and even in the soils spiked with phenanthrene or pyrene did not show major changes during the incubation (data not shown). A small but clear transient shift in Mycobacterium community was detected in both soils spiked with phenanthrene and pyrene. At Day 13, band K11 decreased in intensity, relative to the other bands, in the soil spiked with phenanthrene and even completely disappeared in the soil spiked with pyrene, compared to the control soils at Day 13. After 132 d of incubation, the composition of the Mycobacterium community in the spiked soils was the same as its composition after 0, 29, and 55 d of incubation (data not shown).
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Fig. 2. Fig. 3. 16S rRNA gene-based denaturing gradient gel electrophoresis community analysis of indigenous eubacteria (lanes 1–7) and indigenous Mycobacterium spp. (lanes 8–11) in soil for the K3663 soil experiment. Lanes 1–2: eubacterial fingerprints at day zero for the control soil and the acetone control soil, respectively. Lanes 3–7: eubacterial fingerprints for the non-sterilized soil spiked with phenanthrene after 0, 13, 29, 55, and 132 d of incubation, respectively. Lanes 8–11: Mycobacterium fingerprints after 13 d of incubation for the control soil, the acetone control soil and the non-sterilized soils spiked with phenanthrene and pyrene, respectively. Phe, phenanthrene; pyr, pyrene.
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Table 3. Closest GenBank match and sequence similarity of the denaturing gradient gel electrophoresis (DGGE) bands for the K3663 experiment.
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AndE Soil Experiment
After Tenax extraction, the PAH concentration in the extracted AndE soil was decreased from 314 to 32 mg kg–1 for phenanthrene, from 435 to 90 mg kg–1 for the sum of LMW PAHs, from 299 to 112 mg kg–1 for pyrene, and from 969 to 377 mg kg–1 for the sum of HMW PAHs (Fig. 3
). Since the 24 h single-step Tenax extraction primarily targets the fast desorbing PAH fraction, the resulting soil was relatively enriched in slowly desorbing (and therefore less bioavailable) fraction of the original sample (Gomez-Lahoz and Ortega-Calvo, 2005). After 252 d of incubation, PAHs in the control soil were degraded by the indigenous microbial community to 7 mg kg–1 phenanthrene, 21 mg kg–1 LMW PAHs, 34 mg kg–1 pyrene, and 174 mg kg–1 HMW PAHs. Similar concentrations of phenanthrene (11 mg kg–1), LMW PAHs (29 mg kg–1), pyrene (20 mg kg–1), and HMW PAHs (121 mg kg–1) were determined in the Tenax-extracted soil after 252 d of incubation.
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Fig. 3. Polycyclic aromatic hydrocarbon (PAH) concentration in the control soil and Tenax-extracted soil during the AndE soil incubation experiment. Error bars represent one standard error from four replicates. PAHs: phe, phenanthrene; pyr, pyrene; LMW, low-molecular-weight (PAHs with <4 rings of 16 EPA PAHs); HMW, high-molecular-weight (PAHs with >3 rings of 16 EPA PAHs).
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The community dynamics during the incubation were investigated by 16S rRNA gene-based PCR-DGGE and are shown in Fig. 4
for eubacteria and Sphingomonas. At the start of the experiment, the eubacterial communities in the control soil and the extracted soil were not significantly different, indicating that the Tenax extraction did not have a strong influence on the eubacterial community. After 21 d of incubation, one additional band (A1) appeared in the eubacterial fingerprint of the extracted soil and became more intense at 35 and 126 d of incubation. The 16S rRNA gene sequence corresponding with band A1 was closest related to the 16S rRNA gene sequence of a Xanthomonas sp. (Table 4). After 252 d of incubation, the eubacterial community in both the control soil and the extracted soil showed a highly similar profile consisting of one dominant band, i.e., band A2. The corresponding 16S rRNA gene sequence of band A2 matched best with the 16S rRNA gene sequence of an uncultured -proteobacterium. The 16S rRNA gene sequences corresponding to bands A3, A4, A5, and A6 were closest related to the 16S rRNA gene sequences of strains of Bradyrhizobium, Cupriavidus, Pseudomonas, and Caulobacter, respectively.
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Fig. 4. 16S rRNA gene-based denaturing gradient gel electrophoresis community analysis of indigenous eubacteria (lanes 1–10) and indigenous Sphingomonas spp. (lanes 11–18) in soil for the AndE soil experiment. Lanes 1, 3, 5, 7, 9: eubacterial fingerprints for control soil (C) after 0, 21, 35, 126, and 252 d of incubation, respectively. Lanes 2, 4, 6, 8, 10: eubacterial fingerprints for extracted soil (E) after 0, 21, 35, 126, and 252 d of incubation, respectively. Lanes 11–14: Sphingomonas fingerprints for control soil after 0, 21, 35, 126 d of incubation, respectively. Lanes 15–18: Sphingomonas fingerprints for extracted soil after 0, 21, 35, 126 d of incubation, respectively.
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Table 4. Closest GenBank match and sequence similarity of the denaturing gradient gel electrophoresis (DGGE) bands for the AndE experiment.
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The composition of the Sphingomonas community in the extracted soil changed gradually from the start of the experiment, compared to the Sphingomonas community in the control soil (Fig. 4). Bands A7, A8, and A9 became more intense (relative to other bands in the same lane) in the extracted soil compared to the control soil. These bands remained dominant during 126 d of incubation. The 16S rRNA gene sequences corresponding to bands A7, A8, and A9 were most related to the 16S rRNA gene sequences of Sphingopyxis witflariensis, Sphingopyxis chilensis, and Sphingobium herbicidivorans. In contrast, band A11 disappeared in the extracted soil and became gradually less dominant in the control soil, while band A10 remained dominant in both the control soil and the extracted soil. The 16S rRNA gene sequences corresponding to bands A10 and A11 were closest related to the 16S rRNA gene sequences of uncultured Sphingomonas spp. Surprisingly, Sphingomonas was no longer detected in the control soil and the extracted soil after 252 d of incubation. During the incubation experiment, Mycobacterium was never detected in the control soil and the extracted soil using a direct Mycobacterium-specific PCR (data not shown).
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Discussion
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In soil K3663, 92% of the freshly added phenanthrene (1000 mg kg–1) and 67% of the freshly added pyrene (500 mg kg–1) was degraded after 2 wk of incubation. After 150 d most of the freshly added phenanthrene and pyrene was degraded, showing that the added phenanthrene and pyrene was highly bioavailable. The biodegradation of phenanthrene and pyrene during the first 2 wk of incubation was concomitant with an increase (1.2–1.8 log units) in the number of phenanthrene and pyrene degraders in the soils spiked with PAHs, indicating growth-associated biodegradation. Since the community fingerprints in the control and acetone-treated K3663 soil remained similar during the incubation, we conclude that the recorded community changes in the phenanthrene- and pyrene-amended K3663 soils are indeed due to the artificial changes in PAH profile and not due to the overall conditions created by the treatment and incubation procedure. In contrast, in the case of the AndE soil, the PAH concentrations in the control soil itself did rapidly decrease on laboratory incubation with concomitant changes in community profile. The degradation of the intrinsic phenanthrene in the control AndE soil is in accordance with recent data reported by Bueno-Montes (2005). Apparently, the conditions created in the microcosms, such as addition of oxygen by mixing during sampling and increase in moisture content, rapidly induced degradation of the PAHs in soil AndE. Therefore, it cannot be excluded that community changes in the Tenax-extracted soil are rather due to the overall incubation procedure than to the change in the PAH profile. On the other hand, at the end of the experiment, both AndE soils showed similar PAH profiles and it would be expected that similar communities would develop in both control and treated AndE soil which was indeed the case. In addition, although the Tenax extraction removed most of the PAHs (90% of phenanthrene and 63% of pyrene), subsequent incubation resulted in a further (but slower) decrease in PAH content, as a result of degradation of these relatively enriched slowly desorbing fractions (Pignatello and Xing, 1996; Cornelissen et al., 1997). As such, the Tenax extraction did not have an abrupt change of the bioavailability and still a slow degradable fraction was present.
According to our hypothesis, the addition of phenanthrene to soil K3663 would have negatively affected the Mycobacterium community, either in size or in composition. In contrast, the removal of phenanthrene in the AndE soil was expected to result in an increase of Mycobacterium as a specialist in degradation of less available PAHs. In both cases, the results did not follow our hypothesis. In the phenanthrene-amended K3663 soil, the intensity of the Mycobacterium PCR signals did not decrease. In addition, a dominant band (K7) in the eubacterial DGGE community fingerprint, previously shown to correspond to a M. pyrenivorans species (Uyttebroek et al., 2006a), did not decrease in relative intensity throughout the experiment. Since the applied PCR method has a detection limit of 106 CFU g–1 soil (Leys et al., 2005), we therefore assume that no substantial decreases in Mycobacterium community sizes occurred. On the other hand, a minor shift was observed in the Mycobacterium community. A particular Mycobacterium population, corresponding to band K11, decreased in relative intensity. This population is related to M. tusciae, a species frequently found in PAH-contaminated soils (Uyttebroek et al., 2006a; Leys et al., 2005), but never isolated as a PAH degrader. Those data indicate that most of the Mycobacterium populations, except for the M. tusciae population, were not affected by the added high concentrations of bioavailable phenanthrene. However, mycobacteria probably also did not have a major contribution to degradation of the added phenanthrene, as the PCR signal did not increase in intensity during soil incubation and none of the eubacterial bands transiently appearing at Days 13 and 25 were identified as Mycobacterium. Although not all those bands could be sequenced, mycobacteria are not expected to be present among them, as Mycobacterium bands migrate at the lower side of the gel. Our data indicate that, in contrast with our hypothesis, Mycobacterium is not outcompeted by other bacteria in a soil containing high concentrations of bioavailable phenanthrene. Alternatively, well-established Mycobacterium populations might be difficult to outcompete by other organisms.
In contrast with phenanthrene that can be utilized by many different types of bacteria, bacterial isolates that can grow on pyrene as their sole source of carbon and energy are almost exclusively Mycobacterium (Kanaly and Harayama, 2000; Wick et al., 2001). Therefore, it was expected that the addition of pyrene to the K3663 soil would favor the growth of Mycobacterium and possibly induce changes in Mycobacterium community composition. However, the Mycobacterium PCR signal did not increase in intensity and only a minor transient shift in Mycobacterium community composition, similar to the one observed for phenanthrene addition, was apparent. This result might indicate that other bacteria different from Mycobacterium were responsible for degradation of the added pyrene. However, as discussed below, addition of pyrene, in contrast with phenanthrene, did not induce any change in eubacterial community profile and therefore, it cannot be excluded that resident bacteria, inclusive Mycobacterium were involved.
Mycobacterium was detected neither in the Tenax-extracted AndE soil or in the AndE control soil, indicating that the Mycobacterium populations in the soil never exceeded a size larger than 106 CFU g–1 soil. Hence, the artificial creation of a less available PAH contamination in the soil did not induce the growth of a Mycobacterium population exceeding that size, a size frequently found in weathered PAH-contaminated soils (Leys et al., 2005). Since mycobacteria are relatively slow-growing bacteria, Mycobacterium might eventually develop to a dominant community in the AndE soil with a weathered PAH contamination. Therefore, we cannot exclude the appearance of a Mycobacterium population on the longer term in the soil. Alternatively, it can be discussed whether the initial non-detectable Mycobacterium population was not affected by the Tenax treatment. However, Mycobacterium was also not detected in the control AndE soil in which, at the end of the experiment, PAH concentrations were similar to those in the Tenax-extracted soil.
In contrast with our hypothesis, no major changes were found in the Sphingomonas community composition upon contamination of K3663 with fresh phenanthrene. We expected that the Sphingomonas community might have evolved to a less diverse community by selection for particular phenanthrene-degrading Sphingomonas strains. This was not the case. Changes in Sphingomonas community were never found and Sphingomonas was not identified among the eubacterial bands. Soil K3663 may not contain phenanthrene-degrading Sphingomonas phenotypes and apparently, as shown by the eubacterial DGGE data, other bacteria like Pseudomonas were far more competitive under the changed PAH conditions. In the pyrene-amended K3663 soil, no shift in the Sphingomonas community was found. Pyrene has never been reported as a sole carbon source for Sphingomonas and was apparently not an additional selective force for inducing major Sphingomonas community changes.
It was expected that in the extracted AndE soil and eventually also in the control AndE soil, the Sphingomonas community would change as PAH-degrading Sphingomonas populations, might lose their ecological advantage on removal of the most accessible phenanthrene, and eventually decrease in size. This was indeed the case. In the Tenax-extracted soil, novel Sphingomonas populations gradually became dominant while others lost dominance. Finally, in both soils, the Sphingomonas community was not detectable (detection limit of 104 CFU g–1 soil) after 252 d of incubation indicating that Sphingomonas was completely out-competed by other bacteria. The latter is contradictory to observations with soils with a weathered PAH contamination in which Sphingomonas communities were present, but apparently without actual PAH degraders (Vanbroekhoven et al., 2004).
Interestingly, the addition of phenanthrene or removal of bioavailable phenanthrene had clear effects on eubacterial community in the soils, indicating that PAH profile indeed affects community structure. In the K3663 soil, phenanthrene addition had a clear but transient effect on the eubacterial community composition, showing various novel bands after 13 and 29 d of incubation. Additional bands corresponded to 16S rRNA gene sequences of a Pseudomonas sp. and uncultured -proteobacteria. Since those populations appeared and disappeared concomitant with phenanthrene addition and removal, we suggest that they are linked to the degradation of the added phenanthrene or at least to degradation of metabolites produced from phenanthrene by other organisms and to the biomass change. Pseudomonads are well known for their ability to degrade LMW PAHs such as naphthalene and phenanthrene (Cerniglia, 1992; García-Junco et al., 2001; Wick et al., 2001). Based on 16S rRNA gene sequence, band K2 is related to an organism linked to a hydrocarbon contamination (Kaplan and Kitts, 2004). In contrast with phenanthrene, the addition of pyrene did not result in recorded changes in the eubacterial community composition. These results show that in contrast with phenanthrene, added pyrene did not exert a major additional selective effect on the existing dominant eubacterial populations, although pyrene was efficiently consumed and resulted in an increase in biomass of the pyrene degraders. These results indicate that either (i) the indigenous eubacterial community was well adapted to pyrene degradation and all populations were in one way or another equally profiting from it, (ii) other carbon sources remained the selective force despite the high pyrene concentration added, or (iii) pyrene was consumed by populations that did not increase to dominance in the eubacterial community patterns. The latter is unlikely regarding the large increase in population density of culturable phenanthrene and pyrene degraders but cannot be excluded.
In the AndE soil, a -proteobacterium clearly dominated both the control and the Tenax-extracted soil at the end of our experimental incubation time. We measured the PAH mineralization activity of the AndE soil before and after the incubation period by incubating the soil in a slurry in the presence of either 14C-phenanthrene or 14C-pyrene and following 14CO2 production (Uyttebroek, unpublished data, 2006). Mineralization rates for pyrene and for phenanthrene were equally high for the Tenax-extracted soil and the non-extracted soil before and after incubation. This shows that the observed community change did not affect the PAH-degrading potential in the soil. Further research has to show if the uncultured -proteobacterium represents a PAH degrader in the soil and if the condition of reduced bioaivalability was the true reason for its appearance.
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Conclusions
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Our data indicate that the profile of the PAH contamination indeed influences overall community structure and that in a particular PAH-contaminated soil different PAH-degrading bacteria will become dominant during the course of weathering. However, the results regarding the dynamics of the Mycobacterium and Sphingomonas communities did not always follow our hypotheses. Obviously, the ecology and behavior of Mycobacterium and Sphingomonas in soil is more complex than prospected from studies on their physiological characteristics or abundance in PAH-contaminated soils, and that other genera might be involved. On the other hand, the history of the soil might have affected community structure in such a way that certain population changes were not possible or take a much longer time. In addition, certain treatment steps (ex. the Tenax treatment) might have impacted specific populations. Finally, the artificial and abrupt nature of the change in PAH profile is quite different from the change in PAH profile of a PAH contamination in soil during weathering. Clearly, more research and additional inventive experimental designs are needed to describe the ecological niche of PAH-degrading specialists in PAH-contaminated soil.
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ACKNOWLEDGMENTS
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We thank M. Maesen for ACE extraction and GC-MS analysis of soil samples. This work was supported by Grant G.0371.06 of the Fonds voor Wetenschappelijk Onderzoek Vlaanderen (FWO).
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NOTES
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All rights reserved. No part of this periodical may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher.
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TOP
ABSTRACT
INTRODUCTION
