- Email: [email protected]
Determination of biodegradation potential by two culture-independent methods in PAH-contaminated soils
Environmental Pollution 140 (2006) 536e545 www.elsevier.com/locate/envpol
Determination of biodegradation potential by two culture-independent methods in PAH-contaminated soils H.S. Moon a, H.-Y. Kahng b, J.Y. Kim a, J.J. Kukor c, K. Nam a,* a
School of Civil, Urban & Geosystem Engineering, Seoul National University, Seoul 151-742, South Korea Department of Environmental Education, Sunchon National University, Sunchon 540-742, South Korea c Biotechnology Center for Agriculture and the Environments, Rutgers University, New Jersey, NJ 08901, USA b
Received 16 August 2004; accepted 30 June 2005
Biodegradation potential of PAHs in contaminated soils is determined by two culture-independent methods and recommendations made on best approaches. Abstract Biodegradation potentials of polycyclic aromatic hydrocarbons (PAHs) were determined with soil samples collected from various depths of a PAH-contaminated site and of a site nearby where PAHs were not found. Putative dioxygenase genes were amplified by a primer set specific for initial dioxygenases and identified by web-based database homology search. They were further categorized into several groups of which four dioxygenases were selected as probes for DNA hybridization. The hybridization signals according to the presence of putative dioxygenases were positively related to the extent of PAH contamination. However, the signal intensities varied depending on the probes hybridized and moreover were not consistent with PAH biodegradation activities determined by CO2 evolution. Despite widely accepted advantages of molecular biodegradation assessment, our data clearly present the variations of assessment results depending on the genetic information used and suggest that the methodology may tend to underestimate the real biodegradation capacity of a site probably due to the limited dioxygenase database available at the moment. Therefore, the molecular assessment of biodegradation potential should involve a very careful primer and probe design and an extensive microbiological examination of a site of interest to accurately delineate the biodegradation potential of the site. Ó 2005 Elsevier Ltd. All rights reserved. Keywords: Biodegradation potential; Polymerase chain reaction; Dioxygenase; DNA hybridization; Respirometric method; Polycyclic aromatic hydrocarbons; Former manufactured gas plant site
1. Introduction When organic pollutants are released into the environment, especially into soils and sediments, the function and structure of the indigenous microflora change. It is reasonable to assume that microbial species that can degrade contaminants will flourish in a site where contamination occurs, while those that cannot compete with the enriched species or withstand such pollution will diminish with * Corresponding author: Tel.: C82 2 880 1448; fax: C82 2 889 0032. E-mail address: [email protected] (K. Nam). 0269-7491/$ - see front matter Ó 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2005.06.028
time. To carry out a successful bioremediation of a site, it is helpful to understand the microbiological characteristics of the area such as the degradability of the contaminant and the composition of the indigenous microbial population. Recently, the methodology consisting mainly of direct extraction of DNA from soil and hybridization of metabolic genes of interests with specific probes has been proposed to assess the biodegradation potential of a site. In addition to its specificity, simplicity, and speediness (Hamann et al., 1999; Lloyd-Jones et al., 1999; Sei et al., 1999) this type of genetic approach has a great advantage over the conventional cultivation
537
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
method since it can overcome the limitations that are imposed by cultivation. It is often estimated that less than 1e10% of indigenous microorganisms are estimated to be typically cultivated by standard techniques. There are still extreme types and numbers of microorganisms that are viable but unculturable (Amann et al., 1995). The first step in the biodegradation of neutral aromatic rings in the presence of dioxygenase enzyme is the addition of one or two oxygen molecules to the aromatic rings, making the compounds more hydrophilic and hence easily biodegradable. Dioxygenase is a multicomponent enzyme comprised of a short electron transport chain and terminal oxygenase (Cerniglia, 1992; Mason and Commack, 1992). The terminal oxygenase consists of two dissimilar subunits (a2b2): large subunits (a) as the catalytic core and functionally unknown small subunits (b). Every large subunit contains a Reiske-type iron-sulfur center (Batie et al., 1987; Geary et al., 1984), which has two characteristic amino acid sequence motifs surrounding a region of amino acids whose sequence varies from one enzyme to another. The two conservative motifs were used by Cigolini (2000) to design a degenerate primer set for the amplification of dioxygenase gene fragments and hence for the detection of the enzymes. This principle has been used to assess the biodegradation potential of neutral aromatics in soil without cultivation of indigenous microorganisms. In addition to dioxygenase, glutathione S-transferase (GST) has also been used for designing specific primers (Lloyd-Jones et al., 1999). This type of genetic assessment of the biodegradation potential of aromatic hydrocarbons has been widely demonstrated in recent years (Hamann et al., 1999; Lloyd-Jones et al., 1999; Sei et al., 1999; Joshi and Walia, 1996; Wikstro¨m et al., 1996). To develop a remediation technology for an abandoned manufactured gas plant (MGP) site, we determined the biodegradation potential of polycyclic aromatic hydrocarbons (PAHs) in MGP soil. The determinations were done using two culture-independent methods: polymerase chain reaction (PCR) coupled with DNA hybridization and radiorespirometric analysis. For the genetic approach, a degenerate primer set that is specific for initial dioxygenase was used. For the respirometric determination, the evolution of 14CO2 from radiolabeled PAHs by indigenous microorganisms was measured. The results from both approaches were compared, and the implications of the data was discussed in terms of the current limitations of the genetic approach.
2. Materials and methods 2.1. Soil samples and determination of PAHs The area investigated was an MGP site where contaminated by coal tar residues for over 100 years. Soil
samples were collected from four depths (0e2, 3e5, 6e 8, and 10e12 m below the surface) of the contaminated site and from an adjacent uncontaminated site at three depths (1e2, 4e5, and 6e8 m below the surface). The texture of the soil at different depths and the concentrations of six selected PAHs (i.e., naphthalene, fluorene, phenanthrene, pyrene, chrysene, and benzo[a]pyrene) of each soil sample were analyzed, and the results are presented in Table 1. To determine the concentrations, PAHs were extracted from the soil samples by a solvent extraction method as described below. Five grams of a soil sample were mixed with 10 ml each of dichloromethane and acetone in a 50-ml Teflon centrifuge tube. After mixing the soilesolvent suspension on a horizontal shaker (200 rpm) for 48 h at room temperature, the tube was centrifuged at 18,600 ! g for 15 min. The extract was then passed through a 0.45-mm PTFE syringe filter to remove any particulates present. When needed, the extract was concentrated to 1e2 ml by using a rotary evaporator. The extract was analyzed by a gas chromatograph (GC) equipped with a flame ionization detector (Varian Star 3500; Varian Chromatograph Systems, Walnut, CA). The GC was installed with a Rtx-5 silica column crossbonded with 5% diphenyl and 95% dimethylpolysiloxane (30 m ! 0.53 mm inner diameter; Restek Corporation, Bellefonte, PA). The oven temperature was set to 40 C for 6 min, followed by a linear increase of 10 C per minute to 300 C, and then the temperature was held for 15 min. Injector and detector temperatures were maintained at 300 C. Two microliters of the extract were injected, and nitrogen was used as a carrier gas. 2.2. Isolation of DNA from soil samples A total of seven soil samples were used for DNA extraction and further experiments. For the isolation of total DNA from soil, a freeze-and-thaw method (Kerkhof and Ward; 1993) was used with some modifications. Approximately 200 mg of each sample were suspended in 75 ml of 500 mM EDTA solution (pH 9.4) in a 1-ml microcentrifuge vial and the suspension was mixed gently by tapping. The vial was immersed in liquid nitrogen Table 1 Textures and PAHs concentrations of MGP soils collected from various depths PAHs
Amount extracted (mg/g soil) 0e2 m
NAP FLU PHE PYR CHR BaP Texture a
1685 223 539 171 162 91 Loamy sand
Not detected.
3e5 m a
nd 2.7 9.3 3.0 4.1 0.4 Silty clay loam
6e8 m
10e12 m
nd nd nd nd nd nd Sand
nd 2.9 6.2 16.6 18.2 9.5 Sandy loam
538
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
and thawed quickly by placing it in a hot water bath. The process was repeated four times. The soil sample was resuspended with 225 ml of a miniprep solution containing 50 mM glucose, 10 mM EDTA, and 25 mM TriseHCl (pH 8.0) and 100 ml of a lysozyme solution (10 mg/ml). To the vial, 50 ml of 10% sodium dodecyl sulfate (SDS) was added immediately, followed by the addition of 800 ml of phenolechloroform (50:50). The suspension was mixed by a vortex mixer for 1 min to form an emulsion and then centrifuged at 14,000 ! g for 3 min. The top aqueous phase was transferred to a new tube and subjected to a second phenolechloroform extraction. The genetic material in the aqueous phase was precipitated with 50 ml of 3.0 M sodium acetate and 1.0 ml of absolute ethanol. The precipitated DNA was resuspended in 300 ml of sterile deionized water. To remove soil impurities such as humic materials, 0.5 g of CsCl and 10 ml of ethidium bromide (10 mg/ml) were added to the resuspended solution and the mixture was centrifuged for 16 h at 80,000 rpm using a Beckman tabletop ultracentrifuge (Beckman Optima TL, Fullerton, CA). The portion containing DNA was collected by pipetting and placed on a 0.025-mm pore-sized nitrocellulose filter and floated on sterile deionized water in a petri dish for 1e2 h, to remove ethidium bromide and CsCl. 2.3. PCR amplification and identification of dioxygenase genes The biodegradation potential for seven soil samples was determined by using PCR amplification of dioxygenase gene fragments and subsequent DNA hybridization. To detect and amplify dioxygenase genes from soil DNA extracts, a set of degenerate oligonucleotide primers was designed for the conserved Reiske-type iron-sulfur motif from initial dioxygenases found in 23 bacterial species capable of degrading aromatic hydrocarbons (Cigolini, 2000). These dioxygenases mediate the initial step in the conversion of neutral aromatic hydrocarbons to form cis-dihydrodiol intermediates. The following degenerate oligonucleotide primers were used; 5#-AGG GAT CCC CAN CCR TGR TAN SWR CA-3# and 5#-GGA ATT CTG YMG NCA YMG NGG-3#. PCR amplifications were performed in a total volume of 50 ml using Perkin Elmer reagents (Branchburg, NJ). Each reaction contained 5 ml of 10! PCR buffer consisting of 10 mM TriseHCl (pH 8.3) and 50 mM KCl, 100 mM of each dNTP, 2.5 U Taq polymerase, 2.5 mM MgCl2, 40 nmol of each primer, and 100 ng target DNA. Following a hold of 1 min at 95 C, PCR was performed as follows: 1-min denaturation at 95 C, 1-min annealing at 55 C, and 1-min extension at 72 C for 30 cycles followed by a 10-min hold at 72 C. The PCR products were analyzed by electrophoresis on a 1.5% (w/v) agarose gel. The PCR products, about 80 bp long, were further identified by cloning and subsequent DNA sequencing
to determine whether they were really dioxygenase fragments. For this purpose, the putative dioxygenase gene fragments were recovered from the agarose gel using a QIAquick PCR purification kit (QIAGEN Inc., Valencia, CA) and the recovered DNA was cloned using the pGEM-T vector system (Promega, Madison, WI), and over 100 clones were randomly selected. The clones were sequenced by automated fluorescent-Taq cycle sequencing with the ABI 373 sequencer (Applied Biosystems, Foster City, CA). The clones identified as dioxygenase based on GenBank database homology searches using the BLAST algorithm were chosen and a phylogenetic tree was generated with the 100 partial dioxygenase gene sequences. Nucleotide sequence alignments were performed using the Lasergene program MEGALIGN (DNA STAR, Inc., Madison, WI), and the phylogenetic tree was drawn based on the sequence alignments of deduced amino acids by the Clustal method. 2.4. DNA-DNA hybridization To analyze the cloned DNA by Southern hybridization, DNA samples isolated from the seven soils were digested with ClaI, EcoRV, HindIII, NotI, XhoI, and XhoIBamHI and the digested DNA fragments were separated on 1% agarose gel at 120 V for 1 h. After washing the gel twice with deionized water, the gel was shaken in a denaturing solution (0.5 N NaOH, 1.5 M NaCl) for 1 h, neutralized in a neutralizing solution (1 M TriseHCl, pH 8.0; 1.5 M NaCl) for 30 min and then blotted onto a Hybond-N membrane (Amersham-Pharmacia Biotech Inc., Piscataway, NJ) for 10 h. Experimental procedures not specifically outlined herein were as described by Sambrook et al. (1989). The blotted membrane was washed, dried in the air, and then cross-linked by SpectrolinkerÔ (Spectronics Corporation, Rochester, NY). The membrane was prehybridized in a hybridization solution (30% formamide, 5! Denhardt’s solution, 5! SSPE, 100 mg/ml salmon sperm DNA) for 2 h at 42 C, and then hybridized with the denatured probes in the same solution for 20 h at 42 C. The 2-kb XhoI-HindIII fragment of the denatured probes was randomly labeled with [a-32P]dCTP using T4 polynucleotide kinase according to the instructions from Promega. The hybridized membrane was washed three times in the 1st washing solution (2! SSC; 0.05% SDS) for 10 min each time at 42 C and twice in the 2nd washing solution (0.2! SSC; 0.1% SDS) followed by washing twice for 10 min each at 68 C. The washed membrane was air-dried and exposed to X-ray film in a ÿ70 C freezer. 2.5. Respirometric determination of biodegradation potential Radiolabeled phenanthrene (PHE; [9-14C]; specific activity, 14.0 mCi/mmol), anthracene (ANT; [1,2,3,4,
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
4A,9A-14C]; specific activity, 20.6 mCi/mmol), pyrene (PYR; [4,5,9,10-14C]; specific activity, 58.7 mCi/mmol), and benzo[a]pyrene (BaP; [7-14C]; specific activity, 26.6 mCi/mmol) were used to determine biodegradation potential by respirometry. In one instance to measure microbial activity in a pristine soil collected from a farm in NJ (i.e., Lansdale loam), acetate ([UL-14C]; specific activity, 21.2 mCi/mmol) was used. All chemicals were purchased from Sigma Chemical Company (St. Louis, MO). On receipt from the supplier, each radiolabeled PAH was dissolved and diluted in acetone, except for [14C]acetate which was dissolved in sterile deionized water. All radiolabeled stock solutions were then stored at ÿ20 C until needed. The biodegradation activity of the indigenous microbial population was determined by measuring the amount of 14CO2 evolved from the soil samples. Fivegram portions of four contaminated soil samples were placed in 125-ml flasks, and 50 ml of inorganic salts solution (0.10 g CaCl2$2H2O, 0.01 g FeCl3, 0.10 g MgSO4$7H2O, 0.10 g NH4NO3, 0.20 g KH2PO4, and 0.80 g K2HPO4 lÿ1 sterile deionized water; pH 7.0) was added to each flask. To individual soil slurries, each radiolabeled PAH was spiked at the level of 105 dpm (disintegration per minute). For three uncontaminated soil samples, unlabeled PAHs (a mixture of 500 mg for PHE and 50 mg each for ANT, PYR, BaP all dissolved in acetone) were also spiked together with the radiolabeled PAHs to ensure that the amounts of PAHs provided were enough to initiate and maintain microbial degradation activity. Each flask was then sealed with a Teflon-wrapped silicon stopper, through which was placed an 18-gauge needle and a 16-gauge steel cannula. A small vial containing 1.5 ml of 0.5 N NaOH was suspended from the cannula to trap the 14CO2 released from mineralization. The NaOH solution was periodically removed and replaced with a fresh solution, and the amount of evolved 14CO2 was determined with a liquid scintillation counter (Model LS 5000 TD; Beckman Instruments, Inc., Irvine, CA). The flasks were incubated at room temperature for 35 days on a rotary shaker (120 rpm).
539
size of PCR products (Fig. 1). The putative dioxygenase gene fragments were recovered from the lane loaded with DNA extracted from the 0e2 m sample (i.e., lane 5 in Fig. 1) and the recovered gene fragments were introduced into the pGEM-T vectors to obtain the clonal library. After harvesting over 100 clones from the agar plates selective to transformants, genetic materials were isolated from the clones, and they were then sequenced to verify whether they were really a part of the initial dioxygenase genes. About 100 identified dioxygenase clones were obtained by this procedure and numbered in series. They were further categorized into six groups based on a phylogenetic relationship (Fig. 2). Over 70% of the sequences showed very high similarity with one- or two-base difference, and the clone named Diox 20 was chosen as a representative of the majority sequence. In addition, the clones named Diox 11, Diox 35, and Diox 98 were selected as genetically divergent dioxygenase genes. The majority sequence and the sequences selected for their divergence are compared and shown in Fig. 3. 3.2. DNA-DNA hybridization The selected four dioxygenase clones, which were chosen to represent the full range of sequence diversity found in the clonal library, were used as probes in a subsequent round of Southern hybridization. Total soil DNA isolated from each of the seven soil samples was used for hybridization with the four probes to determine the PAH-degrading potential of the soils. Very strong hybridization signals were obtained with DNA samples extracted from the surface soil (0e2 m) and the coal tarcontaminated site (3e5 m depth) (Fig. 4). Since the two
3. Results 3.1. PCR amplification and identification of dioxygenases DNA extracts obtained from four contaminated and three uncontaminated soil samples were used for PCR amplification with the universal primers. The expected size of PCR products was about 80 bp, which was putative dioxygenase gene fragments. Out of the seven DNA samples, only the DNA extracts of contaminated soils collected from 0e2 and 3e5 m produced the expected
Fig. 1. Putative dioxygenase gene fragments amplified with the universal primer set. The expected size of the PCR products was about 80 bp, and two bands (lanes 5 and 6) were observed at the corresponding location on the 1.5% agarose gel. Lane 1, size marker (l-Hind III); Lane 2, uncontaminated soil from 1e2 m; Lane 3, uncontaminated soil from 3e5 m; Lane 4, uncontaminated soil from 6e8 m; Lane 5, contaminated soil from 0e2 m; Lane 6, contaminated soil from 3e5 m; Lane 7, contaminated soil from 6e8 m; Lane 8, contaminated soil from 10e12 m.
540
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
Fig. 2. Phylogenetic tree of dioxygenase gene fragments obtained from the contaminated topsoil showing a great variety of dioxygenases in the soil.
signal intensities were not easily distinguished by visual inspection, they were measured by a PhosphorImager SI (Molecular Dynamics, Sunnyvale, CA) (data not shown). The strongest intensity was found in the surface soil DNA extract and the second in the DNA isolated from 3e5 m depth. Very weak or no hybridization was obtained from DNA samples isolated from soils at greater depths and from uncontaminated soils of all three depths. These results correlated positively with the amounts of PAHs extracted from the soil samples (Table 1). However, in fact, all the seven soil samples showed significant amounts of biodegradation activity when determined by the respirometric method as described below. The results indicate a great variety of
dioxygenase genes that are present in uncontaminated and contaminated soils, and the dioxygenase genes responsible for the degradation seem to be very diverse among the seven soil samples. The gene probes that had been made from the DNA extract of contaminated surface soil did not hybridize significantly with the dioxygenase genes present in the DNA samples isolated from three uncontaminated soils. 3.3. Respirometric determination of biodegradation potential Biodegradation of PAHs in the same soil samples by indigenous microorganisms was also tested using
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
541
Fig. 3. Deduced amino acid sequences of four gene probes used for DNA-DNA hybridization. Diox 20 was selected as a representative of the majority sequence of dioxygenases, and the other three were to represent the full range of sequence diversity found in the soil DNA extracts.
radioisotope analyses. 14C-labeled PAHs including phenanthrene (PHE), anthracene (ANT), pyrene (PYR), and benzo[a]pyrene (BaP) were individually spiked to soil samples and the degradation of each compound in individual samples was monitored by determining the evolved 14 CO2 using liquid scintillation counting. In uncontaminated soils, PHE was readily mineralized to about 40e 60% depending on the depths of the soil samples collected (Fig. 5). The general order of mineralization extents was PHE O ANT O PYR O BaP. Surface soil and the second depth soil had similar biodegradation potential for each of the four PAHs tested while the soil sample from 6e8 m showed much lower degradation potential. Although the uncontaminated soil samples did not show recognizable lag periods for the onset of
biodegradation of for instance PHE and ANT (Fig. 5), a pristine farm soil (i.e., Lansdale loam) exhibited substantial lag periods before the initiation of biodegradation. As shown in Fig. 6, the soil showed appreciable lag periods with respect to molecular weight dependence before initiation of the biodegradation of PHE, ANT, and PYR. In the case of BaP, essentially no 14CO2 was evolved. In contrast, acetate tested as a universal substrate was readily utilized, indicating that the active microflora in the soil were not yet adapted for the degradation of xenobiotics such as PAHs when the test began. The biodegradation data of the four PAHs in the contaminated soils were slightly complex. The biodegradation potential for PHE was the highest and the
Fig. 4. DNA-DNA hybridization of DNA extracts isolated from the seven soil samples with the four probes selected. Lane descriptions are the same as Fig. 1.
542
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
Fig. 5. Biodegradation of PAHs in uncontaminated soils collected from three different depths by indigenous microorganisms. The extent of mineralization was determined by measuring the evolved 14CO2. Points in the figures are the means of triplicate determinations (filled triangle, soil sample from 0e2 m from the surface; open circle, soil sample from 3e5 m from the surface; filled square, soil sample from 6e8 m from the surface).
degradation for ANT was the second highest (Fig. 7). However, BaP was more easily mineralized than PYR, an observation that was not consistent with the results from uncontaminated soils. Furthermore, in all cases except for PHE, the highest biodegradation potential was observed with the soil sample collected from 10e12 m.
Fig. 6. Biodegradation of acetate as a universal substrate and PAHs in a pristine farm soil by indigenous microorganisms. The extent of mineralization was determined by measuring the evolved 14CO2. Points in the figures are the means of triplicate determinations.
The extent of mineralization of ANT and BaP, in particular between the soil samples from 10e12 m differed from that of the other depths by about 20%.
4. Discussion Degenerate PCR primers were designed to amplify dioxygenase gene fragments from enzymes capable of dihydroxylating neutral aromatic hydrocarbons. The primers were created by aligning consensus regions in the amino acid sequences of known dioxygenase genes. A short stretch of 26 amino acids located in the large subunit of the aligned dihydroxylating dioxygenases was determined to be the most appropriate region for primer design. The specificity of the primers used in this study was extensively confirmed with DNA extracts of pure strains and environmental samples by a previous study (Cigolini, 2000). In the present study, the primer set successfully amplified dioxygenases with extremely high specificity from MGP soil samples. Over 99% of the clones possessing the amplified PCR products was confirmed to be dioxygenase gene fragments by DNA sequencing. Besides specificity, the primer set was also able to detect several different groups of dioxygenase genes from MGP soil samples. The observed diversity
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
543
Fig. 7. Biodegradation of PAHs in the contaminated soils collected from four different depths by indigenous microorganisms. The extent of mineralization was determined by measuring the evolved 14CO2. Points in the figures are the means of triplicate determinations (filled diamond, soil sample from 0e2 m from the surface; open square, soil sample from 3e5 m from the surface; filled triangle, soil sample from 6e8 m from the surface; open circle, soil sample from 10e12 m from the surface).
is, in a sense, natural since the soil has been contaminated with a variety of hydrocarbons for over 100 years. DNA hybridization data demonstrate that signal intensity increased as the concentration of PAHs in the soil increased, a result that is consistent with a previous report by Wikstro¨m et al. (1996). It has been shown that introduction of xenobiotics changed the structure of the indigenous microbial community in the area where contamination occurred (Stapleton et al., 2000), and such a change is related to the contaminant concentrations (Hubert et al., 1999). When contaminated by the same pollutants, the finally established communities were found to be very similar in different soils, regardless of the starting community structures (Massol-Deya et al., 1997). However, considering that most bacterial species degrading PAHs have been isolated from contaminated soils so far and the gene probes used in this study also originated from MGP soil, it is likely that the probes used are much more homologous to dioxygenase genes found in contaminated soils than dioxygenases present in uncontaminated soils. This might have caused the absence or weakness of hybridization signals in uncontaminated soils despite the fact that the soils had appreciable amounts of actual biodegradation activity as confirmed through the respirometric study. The data from the respirometric study showed that biodegradation of aromatic hydrocarbons was the lowest in the pristine farm soil, higher in the uncontaminated
soils, and highest in the contaminated soils. The appreciable amounts of biodegradation with short lag periods in the uncontaminated soils indicate that the soil microflora have been somewhat influenced, probably due to their close proximity to the MGP site, although PAHs were not detected in the uncontaminated soils. Our data show that soils from higher depths at the MGP site exhibited unexpectedly active biodegradation. Although the reason for this unexpected result is not clear at the moment, it is plausible that indigenous microflora at these depths have been suppressed by anoxic and other prevailing conditions to some extent in the most heavily contaminated surface soil. Indeed, it is known that heavy contamination can devastate the microbial communities of a site so that an active biodegradation of the contaminated compounds is sometimes observed in a less contaminated site (Long et al., 1995). In addition, the observation indicates that a great variety of microorganisms that can aerobically utilize PAHs are present in subsurface soils. After being sampled, the soils were stored aerobically so that aerobic microflora responsible for PAH degradation might have been enriched. This provides a possibility that deep subsurface contaminants can be aerobically biodegraded when oxygen is delivered to the site of interest. It was shown that vinyl chloride in groundwater could be biodegraded by the introduction of a phosphate-intercalated peroxygen (Davis and Carpenter, 1990; Hartsman and DeBont, 1992). A natural attenuation study also
544
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
demonstrated the successful development of an aerobic BTEX-degrading community in sediments following petroleum spills (Stapleton et al., 2000). Comparison of the data from DNA hybridization and respirometry suggests that this kind of genetic approach may incorrectly predict the actual biodegradation potential of a site. At a glance, the results from the two culture-independent methods match well for the MGP soil samples. Contaminated soil samples from 0e2 and 3e5 m showed the greatest DNA hybridization signal, and higher mineralization activities than the samples from deeper levels. However, a more prudent examination of the two data revealed that the amount of 14CO2 evolved was the greatest in the contaminated surface soil only when phenanthrene was used as a substrate, which was consistent with DNA hybridization data. For all the other aromatic hydrocarbons tested, biodegradation data were not quantitatively consistent with the gene probe analyses. In terms of mineralization extents, the genetic approach failed to predict the biodegradation potential of soils from higher depths of the MGP site and all soil samples from the uncontaminated site. Since the primer set used for this study was designed based on the dioxygenase sequences from cultivated bacteria it is highly possible that the biodegradation activity of PAHs observed in soils from higher depths of the MGP site might have originated from the bacterial species that have not been identified yet. A similar result was also previously reported. In a study using three known sequences (i.e., nahAc from P. putida G7, phnAc from Burkholderia RP007, and the GSTencoding gene from Sphingomonas WP01) as probes, only 47% of 79 isolated PAH-degrading bacteria hybridized to any of the three PAH catabolic probes (Lloyd-Jones et al., 1999). Our data suggest the potential limitations of the current primer design and of the use of such primers for PCR amplification and subsequent DNA hybridization to characterize the biodegradation potential of contaminants at a site. Such limitation may be due to the limited database available for dioxygenase sequences. The primer sequences were designed based on known dioxygenase sequences of cultivated bacteria from the NCBI database (National Center for Biotechnology Information, http:// www.ncbi.nlm.nih.gov). The size of the database is expanding by adding the sequences from uncultivated bacterial species. In the field of bacterial phylogeny, for example, a variety of rRNA sequences have been uploaded by cloning directly from environmental DNA, or after amplification by the PCR. The sequences are from uncultivated bacterial species, and so far 13 of 36 phylogenetic divisions are characterized only by environmental sequences. They currently have no cultivated representatives and are termed ‘candidate divisions’ (Dojka et al., 2000; Hugenholtz et al., 1998). Such expansion in the database is limited to the structural diversity of
bacterial species. In terms of bacterial functions such as for instance ring oxidation, however, the database is mainly based on cultivated bacterial sequences and thus the available genetic data is still not enough. The present study suggests the abundant presence of unknown microorganisms capable of degrading PAHs in MGP soils. Some of the microorganisms may have little genetic homology to the dioxygenases found in the current database. Some studies have proposed that dioxygenases of different genera and substrates specificities are more divergent or of independent origin (Goyal and Zylstra, 1996; Zylstra et al., 1996). Our data also demonstrated that the true biodegradation potential of a site may be considerably underestimated if the assessment relies on some limited known dioxygenase genes only (Hamann et al., 1999; Berardesco et al., 1998; Herrick et al., 1997; Whyte et al., 1996). Therefore, genetic approaches should be performed with deliberate considerations of their limitations and completeness of the databases on which the amplification and sequencing are based. Wherever possible, it is more reliable to conduct a supporting experiment such as respirometric determination until the current limitations are resolved. Acknowledgements This research was supported by grants from the New Jersey Hazardous Substance Management Research Center and the National Institute of Environmental Health Sciences. Additional support was provided by the KOSEF through the AEBRC at POSTECH. The authors also thank the Research Institute of Engineering Science at Seoul National University for technical assistance. References Amann, R.I., Ludwig, W., Schleifer, K.H., 1995. Phylogenetic identification and in situ detection of individual microbial cells without cultivation. Microbiological Reviews 59, 143e169. Batie, C.J., LaHarie, E., Ballou, D.P., 1987. Purification and characterization of phthalateoxygenase and phthalate oxygenase reductase from Pseudomonas cepacia. Journal of Biological Chemistry 262, 1510e1518. Berardesco, G., Dyhrman, S., Gallagher, E., Shiaris, M.P., 1998. Spatial and temporal variation of phenanthrene-degrading bacteria in intertidal sediments. Applied and Environmental Microbiology 64, 2560e2565. Cerniglia, C.E., 1992. Biodegradation of polycyclic aromatic hydrocarbons. Biodegradation 3, 351e368. Cigolini, J.F., 2000. Molecular analysis of polycyclic aromatic hydrocarbon degradation by Mycobacterium sp. strain PY01. PhD dissertation, Rutgers University, NJ, USA. Davis, J., Carpenter, C.L., 1990. Aerobic biodegradation of vinyl chloride in groundwater samples. Applied and Environmental Microbiology 56, 3878e3880. Dojka, M.A., Harris, K., Pace, N.R., 2000. Expanding the known diversity and environmental distribution of an uncultured
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545 phylogenetic division of bacteria. Applied and Environmental Microbiology 66, 1617e1621. Geary, P.J., Saboowalla, F., Patil, D.S., Commack, R., 1984. An investigation of the iron-sulphur proteins of benzene dioxygenase from Pseudomonas putida by electron spin resonance spectroscopy. Biochemistry Journal 217, 667e673. Goyal, A.K., Zylstra, G.J., 1996. Molecular cloning of novel genes for polycyclic aromatic hydrocarbon degradation from Comamonas testosteroni GZ 39. Applied and Environmental Microbiology 62, 230e236. Hamann, C., Hegemann, J., Hildebrandt, A., 1999. Detection of polycyclic aromatic hydrocarbon degradation genes in different soil bacteria by polymerase chain reaction and DNA hybridization. FEMS Microbiology Letters 173, 255e263. Hartsman, S., DeBont, J.A.M., 1992. Aerobic vinyl chloride metabolism in Mycobacterium Aurum L1. Applied and Environmental Microbiology 58, 1220e1226. Herrick, J.B., Stuart-Keil, K.G., Ghiorse, W.C., Madsen, E.L., 1997. Natural horizontal transfer of a naphthalene dioxygenase gene between bacteria native to a coal tar-contaminated field site. Applied and Environmental Microbiology 63, 2330e2337. Hubert, C., Shen, Y., Voordouw, G., 1999. Composition of toluenedegrading microbial communities from soils at different concentrations of toluene. Applied and Environmental Microbiology 65, 3064e3070. Hugenholtz, P., Goebel, B.M., Pace, N.R., 1998. Impact of cultureindependent studies on the emerging phylogenetic view of bacterial diversity. Journal of Bacteriology 180, 4765e4774. Joshi, B., Walia, S., 1996. PCR amplification of catechol 2,3-dioxygenase gene sequences from naturally occurring hydrocarbon degrading bacteria isolated from petroleum hydrocarbon contaminated groundwater. FEMS Microbiology Ecology 19, 5e15. Kerkhof, L., Ward, B., 1993. Comparison of nucleic acid hybridization and fluorometry for measurement of RNA/DNA relationship with growth rate in a marine bacterium. Applied and Environmental Microbiology 59, 1303e1307. Long, S.C., Aelion, C.M., Dobbins, D.C., Pfaender, F.K., 1995. A comparison of microbial community characteristics among
545
petroleum-contaminated and uncontaminated subsurface soil samples. Microbiology Ecology 30, 297e307. Lloyd-Jones, G., Laurie, A.D., Hunter, D.W.F., Fraser, R., 1999. Analysis of catabolic genes for naphthalene and phenanthrene degradation in contaminated New Zealand soils. FEMS Microbiology Ecology 29, 69e79. Mason, J.R., Commack, R., 1992. The electron-transport proteins of hydroxylating bacterial dioxygenases. Annual Reviews in Microbiology 46, 277e305. Massol-Deya, A., Weller, R., Rios-Hernandez, L., Zhou, J.-Z., Hickey, R.F., Tiedje, J.M., 1997. Succession and convergence of biofilm communities in fixed-film reactors treating aromatic hydrocarbons in groundwater. Applied and Environmental Microbiology 63, 270e276. Sambrook, J., Fritch, E.F., Maniatis, T., 1989. Molecular Cloning: A Laboratory Manual. Cold Spring Harbor Laboratory Press, Cold Spring Harbor, New York. Sei, K., Asano, K.-I., Tateishi, N., Mori, K., Ike, M., Fujita, M., 1999. Design of PCR primers and gene probes for the general detection of bacterial populations capable of degrading aromatic compounds via catechol cleavage pathways. Journal Bioscience and Bioengineering 88, 542e550. Stapleton, R.D., Sayler, G.S., Boggs, J.M., Libelo, E.L., Stauffer, T., Macintyre, W.G., 2000. Changes in subsurface catabolic gene frequencies during natural attenuation of petroleum hydrocarbons. Environmental Science and Technology 34, 1991e1999. Whyte, L.G., Greer, C.W., Inniss, W.E., 1996. Assessment of the biodegradation potential of psychrotrophic microorganisms. Canadian Journal of Microbiology 42, 99e106. Wikstro¨m, P., Wiklund, A., Anderson, A.-C., Forsman, M., 1996. DNA recovery and PCR quantification of catechol 2,3-dioxygenase genes from different soil types. Journal of Bacteriology 52, 107e120. Zylstra, G.J., Wang, X.P., Kim, E., Didolkar, V.A., 1996. Cloning and analysis of the genes for polycyclic aromatic hydrocarbon degradation. Annals of The New York Academy of Sciences 721, 386e398.
Determination of biodegradation potential by two culture-independent methods in PAH-contaminated soils H.S. Moon a, H.-Y. Kahng b, J.Y. Kim a, J.J. Kukor c, K. Nam a,* a
School of Civil, Urban & Geosystem Engineering, Seoul National University, Seoul 151-742, South Korea Department of Environmental Education, Sunchon National University, Sunchon 540-742, South Korea c Biotechnology Center for Agriculture and the Environments, Rutgers University, New Jersey, NJ 08901, USA b
Received 16 August 2004; accepted 30 June 2005
Biodegradation potential of PAHs in contaminated soils is determined by two culture-independent methods and recommendations made on best approaches. Abstract Biodegradation potentials of polycyclic aromatic hydrocarbons (PAHs) were determined with soil samples collected from various depths of a PAH-contaminated site and of a site nearby where PAHs were not found. Putative dioxygenase genes were amplified by a primer set specific for initial dioxygenases and identified by web-based database homology search. They were further categorized into several groups of which four dioxygenases were selected as probes for DNA hybridization. The hybridization signals according to the presence of putative dioxygenases were positively related to the extent of PAH contamination. However, the signal intensities varied depending on the probes hybridized and moreover were not consistent with PAH biodegradation activities determined by CO2 evolution. Despite widely accepted advantages of molecular biodegradation assessment, our data clearly present the variations of assessment results depending on the genetic information used and suggest that the methodology may tend to underestimate the real biodegradation capacity of a site probably due to the limited dioxygenase database available at the moment. Therefore, the molecular assessment of biodegradation potential should involve a very careful primer and probe design and an extensive microbiological examination of a site of interest to accurately delineate the biodegradation potential of the site. Ó 2005 Elsevier Ltd. All rights reserved. Keywords: Biodegradation potential; Polymerase chain reaction; Dioxygenase; DNA hybridization; Respirometric method; Polycyclic aromatic hydrocarbons; Former manufactured gas plant site
1. Introduction When organic pollutants are released into the environment, especially into soils and sediments, the function and structure of the indigenous microflora change. It is reasonable to assume that microbial species that can degrade contaminants will flourish in a site where contamination occurs, while those that cannot compete with the enriched species or withstand such pollution will diminish with * Corresponding author: Tel.: C82 2 880 1448; fax: C82 2 889 0032. E-mail address: [email protected] (K. Nam). 0269-7491/$ - see front matter Ó 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2005.06.028
time. To carry out a successful bioremediation of a site, it is helpful to understand the microbiological characteristics of the area such as the degradability of the contaminant and the composition of the indigenous microbial population. Recently, the methodology consisting mainly of direct extraction of DNA from soil and hybridization of metabolic genes of interests with specific probes has been proposed to assess the biodegradation potential of a site. In addition to its specificity, simplicity, and speediness (Hamann et al., 1999; Lloyd-Jones et al., 1999; Sei et al., 1999) this type of genetic approach has a great advantage over the conventional cultivation
537
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
method since it can overcome the limitations that are imposed by cultivation. It is often estimated that less than 1e10% of indigenous microorganisms are estimated to be typically cultivated by standard techniques. There are still extreme types and numbers of microorganisms that are viable but unculturable (Amann et al., 1995). The first step in the biodegradation of neutral aromatic rings in the presence of dioxygenase enzyme is the addition of one or two oxygen molecules to the aromatic rings, making the compounds more hydrophilic and hence easily biodegradable. Dioxygenase is a multicomponent enzyme comprised of a short electron transport chain and terminal oxygenase (Cerniglia, 1992; Mason and Commack, 1992). The terminal oxygenase consists of two dissimilar subunits (a2b2): large subunits (a) as the catalytic core and functionally unknown small subunits (b). Every large subunit contains a Reiske-type iron-sulfur center (Batie et al., 1987; Geary et al., 1984), which has two characteristic amino acid sequence motifs surrounding a region of amino acids whose sequence varies from one enzyme to another. The two conservative motifs were used by Cigolini (2000) to design a degenerate primer set for the amplification of dioxygenase gene fragments and hence for the detection of the enzymes. This principle has been used to assess the biodegradation potential of neutral aromatics in soil without cultivation of indigenous microorganisms. In addition to dioxygenase, glutathione S-transferase (GST) has also been used for designing specific primers (Lloyd-Jones et al., 1999). This type of genetic assessment of the biodegradation potential of aromatic hydrocarbons has been widely demonstrated in recent years (Hamann et al., 1999; Lloyd-Jones et al., 1999; Sei et al., 1999; Joshi and Walia, 1996; Wikstro¨m et al., 1996). To develop a remediation technology for an abandoned manufactured gas plant (MGP) site, we determined the biodegradation potential of polycyclic aromatic hydrocarbons (PAHs) in MGP soil. The determinations were done using two culture-independent methods: polymerase chain reaction (PCR) coupled with DNA hybridization and radiorespirometric analysis. For the genetic approach, a degenerate primer set that is specific for initial dioxygenase was used. For the respirometric determination, the evolution of 14CO2 from radiolabeled PAHs by indigenous microorganisms was measured. The results from both approaches were compared, and the implications of the data was discussed in terms of the current limitations of the genetic approach.
2. Materials and methods 2.1. Soil samples and determination of PAHs The area investigated was an MGP site where contaminated by coal tar residues for over 100 years. Soil
samples were collected from four depths (0e2, 3e5, 6e 8, and 10e12 m below the surface) of the contaminated site and from an adjacent uncontaminated site at three depths (1e2, 4e5, and 6e8 m below the surface). The texture of the soil at different depths and the concentrations of six selected PAHs (i.e., naphthalene, fluorene, phenanthrene, pyrene, chrysene, and benzo[a]pyrene) of each soil sample were analyzed, and the results are presented in Table 1. To determine the concentrations, PAHs were extracted from the soil samples by a solvent extraction method as described below. Five grams of a soil sample were mixed with 10 ml each of dichloromethane and acetone in a 50-ml Teflon centrifuge tube. After mixing the soilesolvent suspension on a horizontal shaker (200 rpm) for 48 h at room temperature, the tube was centrifuged at 18,600 ! g for 15 min. The extract was then passed through a 0.45-mm PTFE syringe filter to remove any particulates present. When needed, the extract was concentrated to 1e2 ml by using a rotary evaporator. The extract was analyzed by a gas chromatograph (GC) equipped with a flame ionization detector (Varian Star 3500; Varian Chromatograph Systems, Walnut, CA). The GC was installed with a Rtx-5 silica column crossbonded with 5% diphenyl and 95% dimethylpolysiloxane (30 m ! 0.53 mm inner diameter; Restek Corporation, Bellefonte, PA). The oven temperature was set to 40 C for 6 min, followed by a linear increase of 10 C per minute to 300 C, and then the temperature was held for 15 min. Injector and detector temperatures were maintained at 300 C. Two microliters of the extract were injected, and nitrogen was used as a carrier gas. 2.2. Isolation of DNA from soil samples A total of seven soil samples were used for DNA extraction and further experiments. For the isolation of total DNA from soil, a freeze-and-thaw method (Kerkhof and Ward; 1993) was used with some modifications. Approximately 200 mg of each sample were suspended in 75 ml of 500 mM EDTA solution (pH 9.4) in a 1-ml microcentrifuge vial and the suspension was mixed gently by tapping. The vial was immersed in liquid nitrogen Table 1 Textures and PAHs concentrations of MGP soils collected from various depths PAHs
Amount extracted (mg/g soil) 0e2 m
NAP FLU PHE PYR CHR BaP Texture a
1685 223 539 171 162 91 Loamy sand
Not detected.
3e5 m a
nd 2.7 9.3 3.0 4.1 0.4 Silty clay loam
6e8 m
10e12 m
nd nd nd nd nd nd Sand
nd 2.9 6.2 16.6 18.2 9.5 Sandy loam
538
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
and thawed quickly by placing it in a hot water bath. The process was repeated four times. The soil sample was resuspended with 225 ml of a miniprep solution containing 50 mM glucose, 10 mM EDTA, and 25 mM TriseHCl (pH 8.0) and 100 ml of a lysozyme solution (10 mg/ml). To the vial, 50 ml of 10% sodium dodecyl sulfate (SDS) was added immediately, followed by the addition of 800 ml of phenolechloroform (50:50). The suspension was mixed by a vortex mixer for 1 min to form an emulsion and then centrifuged at 14,000 ! g for 3 min. The top aqueous phase was transferred to a new tube and subjected to a second phenolechloroform extraction. The genetic material in the aqueous phase was precipitated with 50 ml of 3.0 M sodium acetate and 1.0 ml of absolute ethanol. The precipitated DNA was resuspended in 300 ml of sterile deionized water. To remove soil impurities such as humic materials, 0.5 g of CsCl and 10 ml of ethidium bromide (10 mg/ml) were added to the resuspended solution and the mixture was centrifuged for 16 h at 80,000 rpm using a Beckman tabletop ultracentrifuge (Beckman Optima TL, Fullerton, CA). The portion containing DNA was collected by pipetting and placed on a 0.025-mm pore-sized nitrocellulose filter and floated on sterile deionized water in a petri dish for 1e2 h, to remove ethidium bromide and CsCl. 2.3. PCR amplification and identification of dioxygenase genes The biodegradation potential for seven soil samples was determined by using PCR amplification of dioxygenase gene fragments and subsequent DNA hybridization. To detect and amplify dioxygenase genes from soil DNA extracts, a set of degenerate oligonucleotide primers was designed for the conserved Reiske-type iron-sulfur motif from initial dioxygenases found in 23 bacterial species capable of degrading aromatic hydrocarbons (Cigolini, 2000). These dioxygenases mediate the initial step in the conversion of neutral aromatic hydrocarbons to form cis-dihydrodiol intermediates. The following degenerate oligonucleotide primers were used; 5#-AGG GAT CCC CAN CCR TGR TAN SWR CA-3# and 5#-GGA ATT CTG YMG NCA YMG NGG-3#. PCR amplifications were performed in a total volume of 50 ml using Perkin Elmer reagents (Branchburg, NJ). Each reaction contained 5 ml of 10! PCR buffer consisting of 10 mM TriseHCl (pH 8.3) and 50 mM KCl, 100 mM of each dNTP, 2.5 U Taq polymerase, 2.5 mM MgCl2, 40 nmol of each primer, and 100 ng target DNA. Following a hold of 1 min at 95 C, PCR was performed as follows: 1-min denaturation at 95 C, 1-min annealing at 55 C, and 1-min extension at 72 C for 30 cycles followed by a 10-min hold at 72 C. The PCR products were analyzed by electrophoresis on a 1.5% (w/v) agarose gel. The PCR products, about 80 bp long, were further identified by cloning and subsequent DNA sequencing
to determine whether they were really dioxygenase fragments. For this purpose, the putative dioxygenase gene fragments were recovered from the agarose gel using a QIAquick PCR purification kit (QIAGEN Inc., Valencia, CA) and the recovered DNA was cloned using the pGEM-T vector system (Promega, Madison, WI), and over 100 clones were randomly selected. The clones were sequenced by automated fluorescent-Taq cycle sequencing with the ABI 373 sequencer (Applied Biosystems, Foster City, CA). The clones identified as dioxygenase based on GenBank database homology searches using the BLAST algorithm were chosen and a phylogenetic tree was generated with the 100 partial dioxygenase gene sequences. Nucleotide sequence alignments were performed using the Lasergene program MEGALIGN (DNA STAR, Inc., Madison, WI), and the phylogenetic tree was drawn based on the sequence alignments of deduced amino acids by the Clustal method. 2.4. DNA-DNA hybridization To analyze the cloned DNA by Southern hybridization, DNA samples isolated from the seven soils were digested with ClaI, EcoRV, HindIII, NotI, XhoI, and XhoIBamHI and the digested DNA fragments were separated on 1% agarose gel at 120 V for 1 h. After washing the gel twice with deionized water, the gel was shaken in a denaturing solution (0.5 N NaOH, 1.5 M NaCl) for 1 h, neutralized in a neutralizing solution (1 M TriseHCl, pH 8.0; 1.5 M NaCl) for 30 min and then blotted onto a Hybond-N membrane (Amersham-Pharmacia Biotech Inc., Piscataway, NJ) for 10 h. Experimental procedures not specifically outlined herein were as described by Sambrook et al. (1989). The blotted membrane was washed, dried in the air, and then cross-linked by SpectrolinkerÔ (Spectronics Corporation, Rochester, NY). The membrane was prehybridized in a hybridization solution (30% formamide, 5! Denhardt’s solution, 5! SSPE, 100 mg/ml salmon sperm DNA) for 2 h at 42 C, and then hybridized with the denatured probes in the same solution for 20 h at 42 C. The 2-kb XhoI-HindIII fragment of the denatured probes was randomly labeled with [a-32P]dCTP using T4 polynucleotide kinase according to the instructions from Promega. The hybridized membrane was washed three times in the 1st washing solution (2! SSC; 0.05% SDS) for 10 min each time at 42 C and twice in the 2nd washing solution (0.2! SSC; 0.1% SDS) followed by washing twice for 10 min each at 68 C. The washed membrane was air-dried and exposed to X-ray film in a ÿ70 C freezer. 2.5. Respirometric determination of biodegradation potential Radiolabeled phenanthrene (PHE; [9-14C]; specific activity, 14.0 mCi/mmol), anthracene (ANT; [1,2,3,4,
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
4A,9A-14C]; specific activity, 20.6 mCi/mmol), pyrene (PYR; [4,5,9,10-14C]; specific activity, 58.7 mCi/mmol), and benzo[a]pyrene (BaP; [7-14C]; specific activity, 26.6 mCi/mmol) were used to determine biodegradation potential by respirometry. In one instance to measure microbial activity in a pristine soil collected from a farm in NJ (i.e., Lansdale loam), acetate ([UL-14C]; specific activity, 21.2 mCi/mmol) was used. All chemicals were purchased from Sigma Chemical Company (St. Louis, MO). On receipt from the supplier, each radiolabeled PAH was dissolved and diluted in acetone, except for [14C]acetate which was dissolved in sterile deionized water. All radiolabeled stock solutions were then stored at ÿ20 C until needed. The biodegradation activity of the indigenous microbial population was determined by measuring the amount of 14CO2 evolved from the soil samples. Fivegram portions of four contaminated soil samples were placed in 125-ml flasks, and 50 ml of inorganic salts solution (0.10 g CaCl2$2H2O, 0.01 g FeCl3, 0.10 g MgSO4$7H2O, 0.10 g NH4NO3, 0.20 g KH2PO4, and 0.80 g K2HPO4 lÿ1 sterile deionized water; pH 7.0) was added to each flask. To individual soil slurries, each radiolabeled PAH was spiked at the level of 105 dpm (disintegration per minute). For three uncontaminated soil samples, unlabeled PAHs (a mixture of 500 mg for PHE and 50 mg each for ANT, PYR, BaP all dissolved in acetone) were also spiked together with the radiolabeled PAHs to ensure that the amounts of PAHs provided were enough to initiate and maintain microbial degradation activity. Each flask was then sealed with a Teflon-wrapped silicon stopper, through which was placed an 18-gauge needle and a 16-gauge steel cannula. A small vial containing 1.5 ml of 0.5 N NaOH was suspended from the cannula to trap the 14CO2 released from mineralization. The NaOH solution was periodically removed and replaced with a fresh solution, and the amount of evolved 14CO2 was determined with a liquid scintillation counter (Model LS 5000 TD; Beckman Instruments, Inc., Irvine, CA). The flasks were incubated at room temperature for 35 days on a rotary shaker (120 rpm).
539
size of PCR products (Fig. 1). The putative dioxygenase gene fragments were recovered from the lane loaded with DNA extracted from the 0e2 m sample (i.e., lane 5 in Fig. 1) and the recovered gene fragments were introduced into the pGEM-T vectors to obtain the clonal library. After harvesting over 100 clones from the agar plates selective to transformants, genetic materials were isolated from the clones, and they were then sequenced to verify whether they were really a part of the initial dioxygenase genes. About 100 identified dioxygenase clones were obtained by this procedure and numbered in series. They were further categorized into six groups based on a phylogenetic relationship (Fig. 2). Over 70% of the sequences showed very high similarity with one- or two-base difference, and the clone named Diox 20 was chosen as a representative of the majority sequence. In addition, the clones named Diox 11, Diox 35, and Diox 98 were selected as genetically divergent dioxygenase genes. The majority sequence and the sequences selected for their divergence are compared and shown in Fig. 3. 3.2. DNA-DNA hybridization The selected four dioxygenase clones, which were chosen to represent the full range of sequence diversity found in the clonal library, were used as probes in a subsequent round of Southern hybridization. Total soil DNA isolated from each of the seven soil samples was used for hybridization with the four probes to determine the PAH-degrading potential of the soils. Very strong hybridization signals were obtained with DNA samples extracted from the surface soil (0e2 m) and the coal tarcontaminated site (3e5 m depth) (Fig. 4). Since the two
3. Results 3.1. PCR amplification and identification of dioxygenases DNA extracts obtained from four contaminated and three uncontaminated soil samples were used for PCR amplification with the universal primers. The expected size of PCR products was about 80 bp, which was putative dioxygenase gene fragments. Out of the seven DNA samples, only the DNA extracts of contaminated soils collected from 0e2 and 3e5 m produced the expected
Fig. 1. Putative dioxygenase gene fragments amplified with the universal primer set. The expected size of the PCR products was about 80 bp, and two bands (lanes 5 and 6) were observed at the corresponding location on the 1.5% agarose gel. Lane 1, size marker (l-Hind III); Lane 2, uncontaminated soil from 1e2 m; Lane 3, uncontaminated soil from 3e5 m; Lane 4, uncontaminated soil from 6e8 m; Lane 5, contaminated soil from 0e2 m; Lane 6, contaminated soil from 3e5 m; Lane 7, contaminated soil from 6e8 m; Lane 8, contaminated soil from 10e12 m.
540
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
Fig. 2. Phylogenetic tree of dioxygenase gene fragments obtained from the contaminated topsoil showing a great variety of dioxygenases in the soil.
signal intensities were not easily distinguished by visual inspection, they were measured by a PhosphorImager SI (Molecular Dynamics, Sunnyvale, CA) (data not shown). The strongest intensity was found in the surface soil DNA extract and the second in the DNA isolated from 3e5 m depth. Very weak or no hybridization was obtained from DNA samples isolated from soils at greater depths and from uncontaminated soils of all three depths. These results correlated positively with the amounts of PAHs extracted from the soil samples (Table 1). However, in fact, all the seven soil samples showed significant amounts of biodegradation activity when determined by the respirometric method as described below. The results indicate a great variety of
dioxygenase genes that are present in uncontaminated and contaminated soils, and the dioxygenase genes responsible for the degradation seem to be very diverse among the seven soil samples. The gene probes that had been made from the DNA extract of contaminated surface soil did not hybridize significantly with the dioxygenase genes present in the DNA samples isolated from three uncontaminated soils. 3.3. Respirometric determination of biodegradation potential Biodegradation of PAHs in the same soil samples by indigenous microorganisms was also tested using
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
541
Fig. 3. Deduced amino acid sequences of four gene probes used for DNA-DNA hybridization. Diox 20 was selected as a representative of the majority sequence of dioxygenases, and the other three were to represent the full range of sequence diversity found in the soil DNA extracts.
radioisotope analyses. 14C-labeled PAHs including phenanthrene (PHE), anthracene (ANT), pyrene (PYR), and benzo[a]pyrene (BaP) were individually spiked to soil samples and the degradation of each compound in individual samples was monitored by determining the evolved 14 CO2 using liquid scintillation counting. In uncontaminated soils, PHE was readily mineralized to about 40e 60% depending on the depths of the soil samples collected (Fig. 5). The general order of mineralization extents was PHE O ANT O PYR O BaP. Surface soil and the second depth soil had similar biodegradation potential for each of the four PAHs tested while the soil sample from 6e8 m showed much lower degradation potential. Although the uncontaminated soil samples did not show recognizable lag periods for the onset of
biodegradation of for instance PHE and ANT (Fig. 5), a pristine farm soil (i.e., Lansdale loam) exhibited substantial lag periods before the initiation of biodegradation. As shown in Fig. 6, the soil showed appreciable lag periods with respect to molecular weight dependence before initiation of the biodegradation of PHE, ANT, and PYR. In the case of BaP, essentially no 14CO2 was evolved. In contrast, acetate tested as a universal substrate was readily utilized, indicating that the active microflora in the soil were not yet adapted for the degradation of xenobiotics such as PAHs when the test began. The biodegradation data of the four PAHs in the contaminated soils were slightly complex. The biodegradation potential for PHE was the highest and the
Fig. 4. DNA-DNA hybridization of DNA extracts isolated from the seven soil samples with the four probes selected. Lane descriptions are the same as Fig. 1.
542
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
Fig. 5. Biodegradation of PAHs in uncontaminated soils collected from three different depths by indigenous microorganisms. The extent of mineralization was determined by measuring the evolved 14CO2. Points in the figures are the means of triplicate determinations (filled triangle, soil sample from 0e2 m from the surface; open circle, soil sample from 3e5 m from the surface; filled square, soil sample from 6e8 m from the surface).
degradation for ANT was the second highest (Fig. 7). However, BaP was more easily mineralized than PYR, an observation that was not consistent with the results from uncontaminated soils. Furthermore, in all cases except for PHE, the highest biodegradation potential was observed with the soil sample collected from 10e12 m.
Fig. 6. Biodegradation of acetate as a universal substrate and PAHs in a pristine farm soil by indigenous microorganisms. The extent of mineralization was determined by measuring the evolved 14CO2. Points in the figures are the means of triplicate determinations.
The extent of mineralization of ANT and BaP, in particular between the soil samples from 10e12 m differed from that of the other depths by about 20%.
4. Discussion Degenerate PCR primers were designed to amplify dioxygenase gene fragments from enzymes capable of dihydroxylating neutral aromatic hydrocarbons. The primers were created by aligning consensus regions in the amino acid sequences of known dioxygenase genes. A short stretch of 26 amino acids located in the large subunit of the aligned dihydroxylating dioxygenases was determined to be the most appropriate region for primer design. The specificity of the primers used in this study was extensively confirmed with DNA extracts of pure strains and environmental samples by a previous study (Cigolini, 2000). In the present study, the primer set successfully amplified dioxygenases with extremely high specificity from MGP soil samples. Over 99% of the clones possessing the amplified PCR products was confirmed to be dioxygenase gene fragments by DNA sequencing. Besides specificity, the primer set was also able to detect several different groups of dioxygenase genes from MGP soil samples. The observed diversity
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
543
Fig. 7. Biodegradation of PAHs in the contaminated soils collected from four different depths by indigenous microorganisms. The extent of mineralization was determined by measuring the evolved 14CO2. Points in the figures are the means of triplicate determinations (filled diamond, soil sample from 0e2 m from the surface; open square, soil sample from 3e5 m from the surface; filled triangle, soil sample from 6e8 m from the surface; open circle, soil sample from 10e12 m from the surface).
is, in a sense, natural since the soil has been contaminated with a variety of hydrocarbons for over 100 years. DNA hybridization data demonstrate that signal intensity increased as the concentration of PAHs in the soil increased, a result that is consistent with a previous report by Wikstro¨m et al. (1996). It has been shown that introduction of xenobiotics changed the structure of the indigenous microbial community in the area where contamination occurred (Stapleton et al., 2000), and such a change is related to the contaminant concentrations (Hubert et al., 1999). When contaminated by the same pollutants, the finally established communities were found to be very similar in different soils, regardless of the starting community structures (Massol-Deya et al., 1997). However, considering that most bacterial species degrading PAHs have been isolated from contaminated soils so far and the gene probes used in this study also originated from MGP soil, it is likely that the probes used are much more homologous to dioxygenase genes found in contaminated soils than dioxygenases present in uncontaminated soils. This might have caused the absence or weakness of hybridization signals in uncontaminated soils despite the fact that the soils had appreciable amounts of actual biodegradation activity as confirmed through the respirometric study. The data from the respirometric study showed that biodegradation of aromatic hydrocarbons was the lowest in the pristine farm soil, higher in the uncontaminated
soils, and highest in the contaminated soils. The appreciable amounts of biodegradation with short lag periods in the uncontaminated soils indicate that the soil microflora have been somewhat influenced, probably due to their close proximity to the MGP site, although PAHs were not detected in the uncontaminated soils. Our data show that soils from higher depths at the MGP site exhibited unexpectedly active biodegradation. Although the reason for this unexpected result is not clear at the moment, it is plausible that indigenous microflora at these depths have been suppressed by anoxic and other prevailing conditions to some extent in the most heavily contaminated surface soil. Indeed, it is known that heavy contamination can devastate the microbial communities of a site so that an active biodegradation of the contaminated compounds is sometimes observed in a less contaminated site (Long et al., 1995). In addition, the observation indicates that a great variety of microorganisms that can aerobically utilize PAHs are present in subsurface soils. After being sampled, the soils were stored aerobically so that aerobic microflora responsible for PAH degradation might have been enriched. This provides a possibility that deep subsurface contaminants can be aerobically biodegraded when oxygen is delivered to the site of interest. It was shown that vinyl chloride in groundwater could be biodegraded by the introduction of a phosphate-intercalated peroxygen (Davis and Carpenter, 1990; Hartsman and DeBont, 1992). A natural attenuation study also
544
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545
demonstrated the successful development of an aerobic BTEX-degrading community in sediments following petroleum spills (Stapleton et al., 2000). Comparison of the data from DNA hybridization and respirometry suggests that this kind of genetic approach may incorrectly predict the actual biodegradation potential of a site. At a glance, the results from the two culture-independent methods match well for the MGP soil samples. Contaminated soil samples from 0e2 and 3e5 m showed the greatest DNA hybridization signal, and higher mineralization activities than the samples from deeper levels. However, a more prudent examination of the two data revealed that the amount of 14CO2 evolved was the greatest in the contaminated surface soil only when phenanthrene was used as a substrate, which was consistent with DNA hybridization data. For all the other aromatic hydrocarbons tested, biodegradation data were not quantitatively consistent with the gene probe analyses. In terms of mineralization extents, the genetic approach failed to predict the biodegradation potential of soils from higher depths of the MGP site and all soil samples from the uncontaminated site. Since the primer set used for this study was designed based on the dioxygenase sequences from cultivated bacteria it is highly possible that the biodegradation activity of PAHs observed in soils from higher depths of the MGP site might have originated from the bacterial species that have not been identified yet. A similar result was also previously reported. In a study using three known sequences (i.e., nahAc from P. putida G7, phnAc from Burkholderia RP007, and the GSTencoding gene from Sphingomonas WP01) as probes, only 47% of 79 isolated PAH-degrading bacteria hybridized to any of the three PAH catabolic probes (Lloyd-Jones et al., 1999). Our data suggest the potential limitations of the current primer design and of the use of such primers for PCR amplification and subsequent DNA hybridization to characterize the biodegradation potential of contaminants at a site. Such limitation may be due to the limited database available for dioxygenase sequences. The primer sequences were designed based on known dioxygenase sequences of cultivated bacteria from the NCBI database (National Center for Biotechnology Information, http:// www.ncbi.nlm.nih.gov). The size of the database is expanding by adding the sequences from uncultivated bacterial species. In the field of bacterial phylogeny, for example, a variety of rRNA sequences have been uploaded by cloning directly from environmental DNA, or after amplification by the PCR. The sequences are from uncultivated bacterial species, and so far 13 of 36 phylogenetic divisions are characterized only by environmental sequences. They currently have no cultivated representatives and are termed ‘candidate divisions’ (Dojka et al., 2000; Hugenholtz et al., 1998). Such expansion in the database is limited to the structural diversity of
bacterial species. In terms of bacterial functions such as for instance ring oxidation, however, the database is mainly based on cultivated bacterial sequences and thus the available genetic data is still not enough. The present study suggests the abundant presence of unknown microorganisms capable of degrading PAHs in MGP soils. Some of the microorganisms may have little genetic homology to the dioxygenases found in the current database. Some studies have proposed that dioxygenases of different genera and substrates specificities are more divergent or of independent origin (Goyal and Zylstra, 1996; Zylstra et al., 1996). Our data also demonstrated that the true biodegradation potential of a site may be considerably underestimated if the assessment relies on some limited known dioxygenase genes only (Hamann et al., 1999; Berardesco et al., 1998; Herrick et al., 1997; Whyte et al., 1996). Therefore, genetic approaches should be performed with deliberate considerations of their limitations and completeness of the databases on which the amplification and sequencing are based. Wherever possible, it is more reliable to conduct a supporting experiment such as respirometric determination until the current limitations are resolved. Acknowledgements This research was supported by grants from the New Jersey Hazardous Substance Management Research Center and the National Institute of Environmental Health Sciences. Additional support was provided by the KOSEF through the AEBRC at POSTECH. The authors also thank the Research Institute of Engineering Science at Seoul National University for technical assistance. References Amann, R.I., Ludwig, W., Schleifer, K.H., 1995. Phylogenetic identification and in situ detection of individual microbial cells without cultivation. Microbiological Reviews 59, 143e169. Batie, C.J., LaHarie, E., Ballou, D.P., 1987. Purification and characterization of phthalateoxygenase and phthalate oxygenase reductase from Pseudomonas cepacia. Journal of Biological Chemistry 262, 1510e1518. Berardesco, G., Dyhrman, S., Gallagher, E., Shiaris, M.P., 1998. Spatial and temporal variation of phenanthrene-degrading bacteria in intertidal sediments. Applied and Environmental Microbiology 64, 2560e2565. Cerniglia, C.E., 1992. Biodegradation of polycyclic aromatic hydrocarbons. Biodegradation 3, 351e368. Cigolini, J.F., 2000. Molecular analysis of polycyclic aromatic hydrocarbon degradation by Mycobacterium sp. strain PY01. PhD dissertation, Rutgers University, NJ, USA. Davis, J., Carpenter, C.L., 1990. Aerobic biodegradation of vinyl chloride in groundwater samples. Applied and Environmental Microbiology 56, 3878e3880. Dojka, M.A., Harris, K., Pace, N.R., 2000. Expanding the known diversity and environmental distribution of an uncultured
H.S. Moon et al. / Environmental Pollution 140 (2006) 536e545 phylogenetic division of bacteria. Applied and Environmental Microbiology 66, 1617e1621. Geary, P.J., Saboowalla, F., Patil, D.S., Commack, R., 1984. An investigation of the iron-sulphur proteins of benzene dioxygenase from Pseudomonas putida by electron spin resonance spectroscopy. Biochemistry Journal 217, 667e673. Goyal, A.K., Zylstra, G.J., 1996. Molecular cloning of novel genes for polycyclic aromatic hydrocarbon degradation from Comamonas testosteroni GZ 39. Applied and Environmental Microbiology 62, 230e236. Hamann, C., Hegemann, J., Hildebrandt, A., 1999. Detection of polycyclic aromatic hydrocarbon degradation genes in different soil bacteria by polymerase chain reaction and DNA hybridization. FEMS Microbiology Letters 173, 255e263. Hartsman, S., DeBont, J.A.M., 1992. Aerobic vinyl chloride metabolism in Mycobacterium Aurum L1. Applied and Environmental Microbiology 58, 1220e1226. Herrick, J.B., Stuart-Keil, K.G., Ghiorse, W.C., Madsen, E.L., 1997. Natural horizontal transfer of a naphthalene dioxygenase gene between bacteria native to a coal tar-contaminated field site. Applied and Environmental Microbiology 63, 2330e2337. Hubert, C., Shen, Y., Voordouw, G., 1999. Composition of toluenedegrading microbial communities from soils at different concentrations of toluene. Applied and Environmental Microbiology 65, 3064e3070. Hugenholtz, P., Goebel, B.M., Pace, N.R., 1998. Impact of cultureindependent studies on the emerging phylogenetic view of bacterial diversity. Journal of Bacteriology 180, 4765e4774. Joshi, B., Walia, S., 1996. PCR amplification of catechol 2,3-dioxygenase gene sequences from naturally occurring hydrocarbon degrading bacteria isolated from petroleum hydrocarbon contaminated groundwater. FEMS Microbiology Ecology 19, 5e15. Kerkhof, L., Ward, B., 1993. Comparison of nucleic acid hybridization and fluorometry for measurement of RNA/DNA relationship with growth rate in a marine bacterium. Applied and Environmental Microbiology 59, 1303e1307. Long, S.C., Aelion, C.M., Dobbins, D.C., Pfaender, F.K., 1995. A comparison of microbial community characteristics among
545
petroleum-contaminated and uncontaminated subsurface soil samples. Microbiology Ecology 30, 297e307. Lloyd-Jones, G., Laurie, A.D., Hunter, D.W.F., Fraser, R., 1999. Analysis of catabolic genes for naphthalene and phenanthrene degradation in contaminated New Zealand soils. FEMS Microbiology Ecology 29, 69e79. Mason, J.R., Commack, R., 1992. The electron-transport proteins of hydroxylating bacterial dioxygenases. Annual Reviews in Microbiology 46, 277e305. Massol-Deya, A., Weller, R., Rios-Hernandez, L., Zhou, J.-Z., Hickey, R.F., Tiedje, J.M., 1997. Succession and convergence of biofilm communities in fixed-film reactors treating aromatic hydrocarbons in groundwater. Applied and Environmental Microbiology 63, 270e276. Sambrook, J., Fritch, E.F., Maniatis, T., 1989. Molecular Cloning: A Laboratory Manual. Cold Spring Harbor Laboratory Press, Cold Spring Harbor, New York. Sei, K., Asano, K.-I., Tateishi, N., Mori, K., Ike, M., Fujita, M., 1999. Design of PCR primers and gene probes for the general detection of bacterial populations capable of degrading aromatic compounds via catechol cleavage pathways. Journal Bioscience and Bioengineering 88, 542e550. Stapleton, R.D., Sayler, G.S., Boggs, J.M., Libelo, E.L., Stauffer, T., Macintyre, W.G., 2000. Changes in subsurface catabolic gene frequencies during natural attenuation of petroleum hydrocarbons. Environmental Science and Technology 34, 1991e1999. Whyte, L.G., Greer, C.W., Inniss, W.E., 1996. Assessment of the biodegradation potential of psychrotrophic microorganisms. Canadian Journal of Microbiology 42, 99e106. Wikstro¨m, P., Wiklund, A., Anderson, A.-C., Forsman, M., 1996. DNA recovery and PCR quantification of catechol 2,3-dioxygenase genes from different soil types. Journal of Bacteriology 52, 107e120. Zylstra, G.J., Wang, X.P., Kim, E., Didolkar, V.A., 1996. Cloning and analysis of the genes for polycyclic aromatic hydrocarbon degradation. Annals of The New York Academy of Sciences 721, 386e398.