International Biodeterioration & Biodegradation 63 (2009) 440–449
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Characterization of the bottom sediments contaminated with polychlorinated biphenyls: Evaluation of ecotoxicity and biodegradability Katarı´na Dercova´ a, *, Jana Sˇeligova´ a, Hana Duda´sˇova´ a, Ma´ria Mikula´sˇova´ b, Katarı´na Sˇilha´rova´ c, Lı´via To´thova´ c, Pavel Hucko c a
Slovak University of Technology, Faculty of Chemical and Food Technology, Institute of Biotechnology and Food Science, Department of Biochemical Technology, Radlinske´ho 9, 812 37 Bratislava, Slovakia ´ dolina, Comenius University, 842 15 Bratislava, Slovakia Institute of Cell Biology, Faculty of Natural Sciences, Mlynska c ´bre Water Research Institute, Na zie arm. gen. L. Svobodu 5, 812 49 Bratislava, Slovakia b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 7 November 2008 Received in revised form 4 December 2008 Accepted 8 December 2008 Available online 31 January 2009
ske in East Slovakia, a large amount of PCBs At the locality of the former producer of PCBs Chemko Stra´z (the commercial mixture DELOR 103, an equivalent of AROCLOR 1242) is still persisting in sediments and negatively influences health of the population. The objective of this work was to provide a study of ecotoxicity and genotoxicity of PCBs in contaminated sediments. Toxicity of the PCB-contaminated sediments sampled from Zemplı´nska sˇı´rava and Stra´ zsky canal (surroundings of the former producer of PCBs) was determined applying a standard aquatic plant toxicity test using Lemna minor. The endpoints for the test were frond numbers and frond areas. The sediment sampled from Zemplı´nska sˇı´rava was more toxic to L. minor than the one sampled from Stra´ zsky canal. The results on genotoxicity showed that both sediments were not mutagenic toward the standard strains of the Ames test, Salmonella typhimurium TA98 and TA100. This work deals also with biodegradation of PCBs in two samples of the above mentioned contaminated sediments: a) in the natural sediments by autochthonous microbial consortium and b) in the bioaugmented sediments inoculated by allochthonous bacterial strains, two bacterial isolates from long-term PCB-contaminated soil Pseudomonas stutzeri and Alcaligenes xylosoxidans. Both approaches were applied under the biostimulation conditions, with addition of glucose or biphenyl as co-substrates, as well. The highest PCB degradation was observed in the bioaugmented sediment inoculated with bacterial strain P. stutzeri. Addition of biphenyl, as the co-substrate and the inducer, positively affected degradation of PCBs. The bphA1 gene, encoding enzyme biphenyldioxygenase, responsible for the start of PCB degradation, was identified in genome of P. stutzeri, a potential PCB-degrader isolated from long-term PCB-contaminated soil, but not in genome of A. xylosoxidans. Ó 2009 Elsevier Ltd. All rights reserved.
Keywords: Biodegradation bphA1 gene Ecotoxicity Pseudomonas stutzeri Polychlorinated biphenyls Sediments
1. Introduction Sediment is an essential, integral and dynamic part of the hydrological system. However, because sediments are the ultimate reservoir for the numerous potential chemical and biological contaminants that may be contained in effluents originating from agricultural and industrial lands, contaminated sediments in rivers, lakes, coastal harbors, and estuaries have the potential to pose ecological and human health risks (Apitz et al., 2006). In the recent decades, several hundred tons of PCBs have been released into the environment. Due to their hydrophobic properties, PCBs tend to be adsorbed by natural organic matter in soil, sediment, and sludge.
* Corresponding author. Tel.: þ421 2 59325 710; fax: þ421 2 52967 085. E-mail address:
[email protected] (K. Dercova´). 0964-8305/$ – see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.ibiod.2008.12.005
Major part of the PCBs released into the aquatic environment was thus cumulated in the aqueous sediments (Wiegel and Wu, 2000; Bedard, 2003; Field and Alvarez, 2008). Polychlorinated biphenyls (PCBs) are widespread pollutants that have been produced on a large scale in the past decades transfer between all environmental media. PCBs are classified as the most persistent and toxic industry-produced compounds that have been detected as contaminants in almost every component of the global ecosystem. They represent serious ecological problem due to low degradability, high toxicity, and strong bioaccumulation. These features are associated with their hydrophobicity and low chemical reactivity. The main obstacles preventing mineralization of PCBs in natural environment may be defined as: a) recalcitrance, b) toxicity of degradation intermediates, and c) limited physical availability for microorganisms or chemical oxidants (Vrana et al., 1996a,b; Dercova´ et al., 1996).
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The production of PCBs was stopped after they had been proved to be toxic to humans. Once in the environment, they tend to accumulate in different stages of the food chain and besides their toxicity, they are also potential carcinogenic compounds (McLachlan, 1996). Environmental and economic reasons have urged the development of bioremediation technologies for the PCB removal from contaminated sites. A crucial step is the isolation or genetic construction of microbial strains with biodegradation potential (Menn et al., 2000). Several aerobic microorganisms are able to biodegrade certain, usually less chlorinated, PCB congeners (Bedard et al., 1987; Kohler et al., 1988). Finding or development of PCBdegrading microorganisms, preferably bacteria, with a high degradation potential and colonization ability is the essential element of a successful bioremediation technology of contaminated areas (Pritchard, 1992; Havel and Reineke, 1993). Although the complete PCB degradation is the ultimate goal, often only partial solutions of this complex problem can be reached resulting in elimination of the most harmful properties. The top priority for dealing with PCBs is to lower their bioaccumulation potential or, equivalently, lipophilicity. In the aerobic biodegradation process, the enzymes dioxygenases introduce two hydroxyl groups into the PCB molecule, reduce its lipophilicity and make a subsequent cleavage easier. The initial oxidation and the following ring cleavage seem to be the rate-limiting steps of the process (Sondossi et al., 1991). In our previous papers, a simple apparatus for effective monitoring of the PCB evaporation kinetics in batch biodegradation experiments was described (Vrana et al., 1995; Vrana et al., 1996a,b) together with a simple mathematical model taking evaporation and biodegradation into account (Dercova´ et al., 1999; Vrana et al., 2008). In the locality of the former producer of PCBs in East Slovakia, Chemko Stra´ zske, large amount of these substances is still persisting in sediments of Stra´ zsky canal and Zemplı´nska sˇı´rava water reservoir and negatively influences human health (Langer et al., 2006; Zhiwei et al., 2007). It is recognized that the bottom sediments are substantially contaminated by the commercial mixtures of PCBs DELOR 103 and DELOR 106 (equivalent to AROCLOR 1242 and 1260). The Zemplı´nska sˇı´rava reservoir is filled with the waters of Laborec River at southern piedmont of the Vihorlat Mountain. Its function is to secure sufficiency of water for the Vojany Thermal Power Station and for irrigation purposes. Another function is also a flood run-off regulation at the Laborec River and partly in the whole area of the Eastern Slovakia Lowland. It is also extensively used as a resort for holiday recreation. The objectives of this work were to provide characterization of PCB-contaminated bottom sediments sampled from the above mentioned Zemplı´nska sˇı´rava water reservoir and industrial effluent Stra´ zsky canal in order to study their ecotoxicity and biodegradation of PCBs in these sediments by indigenous (autochthonous) and inoculated (allochthonous) bacterial strains. To determine the genetic potential for biodegradability, the gene bphA1 encoding the enzyme biphenyldioxygenase, responsible for the first step of PCB aerobic degradation, was identified in the bacterial isolates using a PCR technique. 2. Materials and methods 2.1. Chemicals The following reagents were used: agarose (BioRad, USA), bacteriological agar (Oxoid, UK), biphenyl (Sigma–Aldrich, USA), zske, DELOR 103 (commercial mixture of PCBs, Chemko Stra´ Slovakia), diethyl pyrocarbonate (Sigma–Aldrich, USA), DNA markers (standards for the horizontal electrophoresis, Promega, USA), 100 bp ladder, 1 kb ladder DNA polymerase (Finzymes,
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Finland), ethidium bromide (Fluka, Germany), n-hexane (Chromservis, Czech Republic), IRS solution (Mo-Bio Laboratories, USA), Luria-Bertani (LB) medium (Oxoid, UK), 3-(5-nitro-2-furyl)acrylic acid – NAFA (Zentiva Hlohovec, Slovakia), Plate Count Agar (PCA) (Oxoid, UK), primers F350 and R674 (Biotech, Czech Republic), sodium pyrophosphate (Sigma–Aldrich, USA), reaction buffer solution for DNA polymerase (Finzymes, Finland), Trisma base (Lachema, Czech Republic), and mixture of nucleotides (Promega, USA). 2.2. Bacterial strains and sampling of sediments The bacterial strains Burkholderia xenovorans LB400 (bph operon) and the two isolates from long-term contaminated soil (surroundings of the former producer of PCBs) Alcaligenes xylosoxidans and Pseudomonas stutzeri (Dercova´ et al., 1996), purified, identified and maintained as safe keeping in the Czech Collection of Microorganisms (Masaryk University, Brno, Czech Republic) were used. The standard strains for determination of mutagenicity of Salmonella typhimurium TA98 (CCM 3811) and TA100 (CCM 3812), received also from the Czech Collection of Microorganisms (Masaryk University, Brno, Czech Republic), were used as well. Sediment sampling protocol was in agreement with the technical norm ISO 5667-12, 2001. Overview of all sampling sites is presented in Fig. 1a. The sediment corer sampler (UWITEC, Corp. Austria) was used as a sampling device. The sampler used is a transparent plastic tube that allows for visual examination and sample partition into several layers (Fig. 1b). 2.3. Cultivation media Luria-Bertani (LB) medium: It was prepared according to the instruction for use. To congelate the medium, 15 g l1 agar was added (Nobel Agar, Difco, UK). Minimal liquid medium (MM): 1 g (NH4)2SO4; 2.7 g KH2PO4; 10.95 g Na2HPO4$12H2O filled with distilled water to 1 l. After sterilization of the solution, salts containing particular trace elements sterilized by filtration were added: 250 ml FeSO4$7H2O (2 g l1); 250 ml, Ca (NO3)2$4H2O (6 g l1); and 250 ml MgSO4$7H2O (40 g l1), each in 50 ml of the prepared medium. Minimal solid medium: 5.37 g Na2HPO4$12H2O; 1.30 g KH2PO4; 0.50 g NH4Cl; and 0.20 g MgSO4$7H2O filled with distilled water to 1 l. To congelate the medium, 15 g l1 agar was added (Nobel Agar, Difco, UK). Plate Count Agar (PCA): 23.5 g PCA was dissolved in 1 l distilled water in accordance with the guide of Oxoid Company, UK. Phosphate buffer (pH 7.4): Solution A: 13.8 g NaH2PO4$H2O filled with distilled water to 500 ml; solution B: 14.2 g Na2HPO4 filled with distilled water to 500 ml. A total of 60 ml of solution A and 440 ml of solution B were mixed. TBE buffer (pH 8): 54 g Tris, 27.5 g boric acid; 3.72 g EDTA, filled with distilled water to 1 l. TE buffer (pH 8): 0.12 g Tris and 0.037 g EDTA filled with distilled water to 100 ml. Thornton agar: 5 g KNO3, 1 g K2HPO4, 0.2 g MgSO4, 0.1 g CaCl2, 0.1 g NaCl, 0.002 g FeCl3, 0.5 g asparagine, and 1 g D-mannitol were dissolved in 1 l distilled water. The value of pH was adjusted to 7.5. Then, 15 g of agar was added. The agar solution was sterilized (20 min, 120 kPa) and after cooling to 40 C applied to sterile Petri dishes. Minimal glucose medium: Solution I: 10.5 g K2HPO4, 4.5 g KH2PO4, 1 g (NH4)2SO4, 0.55 g C6H5Na3O7$2H2O and sodium citrate were dissolved in 500 ml distilled water under moderate heating, pH was adjusted to 7.2 and solution was sterilized for 20 min at 121 C. Solution II: 15 g Noble agar (Difco) was dissolved in 500 ml
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´ et al. / International Biodeterioration & Biodegradation 63 (2009) 440–449 K. Dercova
Fig. 1. a) Location of the most frequent sampling sites in the Zemplı´nska sˇı´rava water reservoir and vicinity of Stra´zske: 1 – The Stra´zsky canal (our sampling site No. 1 – SK); 2/1, 2/2, 2/3 – west part of the Zemplı´nska sˇı´rava (2/1 – our sampling site No. 2– ZS); 3/1, 3/2 – The ZSR – stone-pit locality; 4/1, 4/2 – middle part of the ZS, Kaluza; 5 – east part of the ZS, Kusı´n; 6 – Laborec River – Krivosˇtˇany; 7 – Laborec River – Petrovce. b) The sediment corer sampler (UWITEC, Corp. Austria).
distilled water. The aqueous agar was sterilized for 20 min at 121 C. Solution III: 20 g MgSO4$7H2O was dissolved under moderate heating in 100 ml distilled water and sterilized for 20 min at 121 C. Solution IV: 20 g glucose was dissolved in 100 ml distilled water under moderate heating and sterilized for 20 min at 121 C. Solution I (500 ml) and solution II (500 ml) were cooled to 50 C and mixed together. 1 ml of solution III and 10 ml of solution IV were added to this mixture. The complete solution was carefully mixed and applied to sterile Petri dishes.
Top agar with minimal contents of histidine: 6 g Noble agar (Difco) and 5 g NaCl were filled with distilled water to 1000 ml. The solution was sterilized for 20 min at 121 C and congelated at the laboratory temperature in the transfusion bottles. They were maintained in the dark at low temperature. Solution HIS/BIO: 9.6 mg L-histidine and 12.36 mg D-biotin were dissolved in 100 ml redistilled water. The solution was sterilized for 20 min at 121 C. After dissolving and heating agar to 42 C on a water bath, 5 ml solution of HIS/BIO was added to 50 ml liquid
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agar with histidine (0.5 mmol l1). Finally, the complete mixture was stirred. 2.4. Toxicity of PCBs (DELOR 103) toward the used bacterial strains Ten ml of MM medium was added into sterile L-tubes. Glucose was used as a carbon source at final concentration 5 g l1. The concentration of the stock solution of DELOR 103 in dimethylsulfoxide (DMSO) was 10 g l1. The final concentrations of DELOR 103 in the individual L-tubes were 5, 10, 30, 50, 75, 100, 150, 200, and 250 mg l1. The final concentration of DMSO was lower than 2% (v/ v). Inoculum was prepared by cultivation of the bacterial strains A. xylosoxidans, P. stutzeri, and Burkholderia sp. LB400 in the nutritional broth during 24 h on a rotary shaker (180 rpm) at 28 C. Biomass was centrifuged and re-suspended in physiologic solution with a final concentration 0.5 g l1. The test tubes were incubated on reciprocal shaker (180 rpm) at 28 C in the dark. The growth of biomass was determined spectrophotometrically at 620 nm.
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Estimation of growth rate: The growth rate was calculated from Eq. (1):
r ¼
ln xt2 ln xt1 t2 t1
(1)
where r is the growth rate per day, xt1 is the value of observation parameter at t1 days; xt2 is the value of observation parameter at t2 days; t2 t1 is the time period between xt1 and xt2, in days. The percent inhibition of growth rate for each test concentration was calculated according to Eq. (2):
Ir ¼
rc rt 100 rc
(2)
where Ir is the inhibition of the average specific growth rates, in percent, %; rc is the average specific growth rate of the control, in d1; rt is the average specific growth rate of the treatment groups, in d1.
2.5. Ecotoxicity of PCB-contaminated bottom sediments
2.6. Cell counting
Specific inhibitory effects of PCB-contaminated bottom sediments on the growth of the bioindicator Lemna minor were measured applying a standard aquatic plant toxicity test. Plant of the species L. minor are allowed to grow as monocultures in different concentrations of the test sample over a period of seven days. The objective of the test is to quantify substance-related effects on vegetative growth over this period based on assessment on biomass (total frond area). To quantify substance-related effects, the growth rate in the test solutions is compared with that of the controls and the concentration resulting in a specified inhibition of growth rate is determined and expressed as the I(r)x. Toxicity test was performed within 7 days in the presence of the toxicant. In the beginning of the test and after 2, 4, and 7 days, the images of the beakers were taken for analysis. To evaluate ecotoxicity of tested sediments, inhibition of number of fronds (individual leaf-like structure on a duckweed colony, the smallest unit capable of reproducing) and frond area were used. The test was performed according to the ISO standard 20079 (2005) using the Steinberg medium. The biotest was carried out in 150 ml beakers filled with 100 ml of a slurry. Preparation of sediment slurries: 50 mg of dry sediment was mixed with 50 ml of Steinberg medium. Ten to 12 fronds (green leaves) were used as the inoculum for each beaker, using only the plants with two or three fronds. Three control replicates and two slurry replicates were used. The data presented are the arithmetical average of data obtained with the replicates. The tests were carried out in a climatic exposure test cabinet, calibrated at 25 2 C with light intensity adjusted to 6500 Lux. The evaluation of L. minor toxicity test was performed using the digital image analysis system Scanalyzer (LemnaTec, GmbH, Germany). The inhibition of frond growth rate (Ir) and frond biomass (Ia) were determined in accordance with ISO 20079 (2005). The parameter Ir is more relevant concerning the toxic effects of the presence of PCBs (DELOR 103) than Ia. Stock solutions for preparation of Steinberg medium (ISO 20079, 2005): Solution I: KNO3, 17.5 mg l1, KH2PO4, 4.5 mg l1, K2HPO4 0.63 mg l1; Solution II: MgSO4$7H2O, 5.0 mg l1; Solution III: Ca(NO3)2$4H2O, 14.75 mg l1; Solution IV: H3BO3, 120 mg l1; Solution V: ZnSO4$7H2O, 180 mg l1; Solution VI: Na2MoO4$2H2O, 44 mg l1; Solution VII: MnCl2$4H2O, 180 mg l1; Solution VIII: FeCl3$6H2O, 760 mg l1, Na2EDTA$2H2O, 1500 mg l1. Twenty millilitres of the stock solutions I, II, III, and 1 ml of stock solutions IV, V, VI, VII, and VIII were supplemented with de-ionized water to 1 l. The value of pH was adjusted to 5.5 0.2 using minimal volume of NaOH or HCl.
Hundred grams of dry sediment was flooded with 300 ml of sterile distilled water and 24 h leach away under intermittent mixing. A volume of 125 ml of a diluted (100 or 1000 times in three replicates) sediment extract was applied to Petri dishes with Thornton agar and incubated for 5 days at 28 C. Viable cells were counted as colony forming units per ml (CFU ml1) using a cell counter and recalculated on 1 g of dry sediment. 2.7. Determination of the mutagenicity The Salmonella/microsome reversion assay was conducted using the plate incorporation procedure described by Maron and Ames (1983). The Ames test was performed utilizing two bacterial strains of S. typhimurium: strain TA98 (CCM 3811) for frame-shift mutations and strain TA100 (CCM 3812) for base-pair substitutions. The tester strains were maintained and stored according to the standard procedures (Mortelmans and Zeiger, 2000). The samples were tested without metabolic activation to detect direct mutagenic compounds. The concentration of stock solution of PCBs dissolved in DMSO was 5 g l1. A mixture of 2 ml of melted top agar containing 50 mM of Lhistidine–biotin, 0.1 ml of bacterial culture (cultivated for 16 h at 37 C, approximate cell density 2–5 108 cells ml1), and 0.1 ml of a solution of the tested compound was poured onto a minimal glucose agar plate and incubated at 37 C for 48 h in the dark. The number of hisþ revertants was estimated with the counter. Data points represent three separate experiments, each performed in triplicate. Positive (3-(5-nitro-2-furyl)-acrylic acid – NAFA) and negative (DMSO) controls were included in each experiment. 2.8. Isolation of the bacterial DNA Bacterial DNA of the used strains was isolated by thermolysis. For isolation, a culture from solid media was used and incubated in 50 ml sterile distilled water for 10 min at 95 C in a thermoblock. After thermolysis, the mixture was stirred and centrifuged (1300 rpm). QIAamp DNA Mini Kit (Qiagen, USA), a set for isolation of the total DNA from pure culture was used. 2.9. Amplification of the bphA1 gene For detection and amplification of the gene of bacterial DNA, a PCR technique with specific primers was used. Composition of the mixture was 34.5 ml sterile de-ionized water, 5 ml of polymerase
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buffer (Finzymes, Finland), 1 ml dNTP (10 mM), 2 1 ml of primers for detection of gene bphA1 (primers F350 and R674 at a concentration of 10 pmol ml1), 2 ml BSA, 0.5 ml of polymerase (Finzymes, Finland), and 5 ml of DNA sample (Rysˇlava´ et al., 2003). PCR amplifications were performed in an automated thermal cycler Techne (Progene, USA) with an initial denaturing (94 C for 5 min), followed by 35 cycles of denaturing (94 C, 30 s), annealing (55 C, 30 s), extension (72 C, 1 min), and concluded by a single final extension (72 C, 10 min). 2.10. Detection of the PCR products by gel electrophoresis A suitable concentration of the agarose gel was used: from 1% (for fragments with number of bp higher than 100) up to 1.5% (for fragments with number of bp below 100). To determine fragment size, a marker of 100 bp DNA standard was used (Sambrook et al., 1989). 2.11. Biodegradation assay in the bottom sediments Biodegradation was carried out in Erlenmeyer flasks. Each flask was equipped with sorbent Silipor (0.5 g) column that was closed with a cotton wool stopper to maintain sterile environment and allow for gas diffusion. This apparatus for simultaneous monitoring of evaporation and biodegradation of PCBs was described previously (Dercova´ et al., 1999). The apparatus allows to determine mass balance of the tested compound. The evaporation rate constants of the individual PCB congeners presented in DELOR 103 were published in our previous papers (Vrana et al., 1995; Vrana et al., 1996a,b). To assess aerobic biodegradation of PCBs present in natural sediments from Stra´ zsky canal and Zemplı´nska sˇı´rava without additional bioaugmentation, 15 g dried sediment and 100 ml liquid minimal medium were added. As the primary carbon source, biphenyl or glucose (5 g l1) was used. Biphenyl was added in a small amount in the form of crystals. Biodegradation was carried out on a rotary shaker (180 rpm) at 28 C in the dark. The control for the abiotic decrease of PCBs was represented by sterilized sediment. The samples were analyzed in the beginning of the experiment and on the 7th and 14th days. Total flask contents were taken for PCB analysis. The amounts of PCBs in the liquid medium and on the sorbent were analyzed. To evaluate aerobic biodegradation of PCBs in the bioaugmented sediments, 10 g natural sediment and inoculum of selected bacterial isolate with final concentration 2 g l1 were used in the 100 ml liquid minimal medium. Other conditions were the same as in non-bioaugmented sediment described previously. 2.12. PCB analysis The sediments were extracted 4 h in the Soxhlet apparatus with n-hexane, filtered and purified through a 5 cm layer of florisil until the sample became transparent. One microlitre of the hexane extract was injected into the Hewlett–Packard gas chromatograph HP5890 series II. Separation of the PCBs was performed on a capillary non-polar column HP-5 (25 m 0.32 mm inner diameter) coated with 0.52 mm (5%-phenyl)-methylpolysiloxane phase. Electron-capture detector was applied and nitrogen was used as carrier gas (flow rate 1.6 ml min1). The temperature program was 160 C for 0.5 min, followed by an increase at a rate of 4 C min1 until the temperature reached 260 C and then kept at this temperature for 5 min. The indicator PCB congeners recommended by EPA (US EPA Methods 8089/8081) were analyzed (IUPAC numbers 8, 28, 52, 101, 118, 138, 153, 180, and 203). Quantitative evaluation of the results was performed based on comparison of the peak areas
corresponding to the indicator congeners in the chromatograms of the sample and of the standard (ISO 6468, 1996). 3. Results and discussion This work concerns determination of ecotoxicity and genotoxicity of two sediments sampled from Zemplı´nska sˇı´rava and Stra´ zsky canal. The attention was also focused on biodegradation of DELOR 103, the commercial mixture of PCBs in natural and bioaugmented sediments. The concentrations of the indicator PCB congeners determined in both sediments are illustrated in Fig. 2 and presented in Table 1 in natural sediments by native autochthonous microbial consortium and in the bioaugmented sediment inoculated by allochthonous microorganisms. Biodegradation was studied in the presence of co-substrates, with addition of glucose and with addition of biphenyl as a structural analog and inducer of biphenyl/PCB degradation pathway. Bioaugmentation of sediments was performed with A. xylosoxidans and P. stutzeri (bacterial strains isolated from long-term PCB-contaminated soil sampled from surroundings of the former PCB producer), as well as using Burkholderia sp. LB400, the control strain possessing a bphA1 gene. To determine the genetic potential of the microbial isolates for biodegradability, the gene bphA1 encoding the enzyme biphenyldioxygenase, responsible for the first step of PCB aerobic degradation, was identified using a PCR technique. The cell counts of the indigenous microorganisms in sampled sediments and toxicity of PCBs toward the bacterial strains used in bioaugmentation were studied, as well. 3.1. Determination of the bottom sediment ecotoxicity It has been recognized that PCBs elicit in humans chronic rather than acute effects. In this work we performed the ecotoxicity assessment of the sediments applying aquatic plant toxicity test using the standard bioindicator L. minor. Steinberg medium was added to the sediments sampled from Zemplı´nska sˇı´rava and Stra´ zsky canal in the ratio 1:1 (v/v). Sterilized empty flasks were filled with 100 ml of the prepared mixtures and growth of inoculated L. minor was determined using Scanalyzer at the beginning of the experiment and on the 2nd, 4th, and 7th days. Frond (means green leaf) numbers and frond areas were evaluated and growth rates and biomass amount were calculated. The measurements were performed in four parallel samples for sediments and in three control samples. The comparison of frond numbers established in the control and sediment samples is illustrated in Fig. 3. Inhibition of growth was evident in both sediments in comparison to the control samples. Marked differences were observed on the 7th day. The tendency observed with the frond numbers was similar to that of the frond areas (data not shown). After comparison with the control, the percentage of the growth rate inhibition (% Ir) and the percentage of biomass inhibition (% Ia) were calculated (Fig. 4). The parameter of growth rate inhibition Ir is more relevant to toxicity than Ia. Study of ecotoxicity has confirmed that both used sediments sampled from Stra´ zsky canal and Zemplı´nska sˇı´rava water reservoir, located in the vicinity of the former producer of PCBs in East Slovakia, were toxic for the tested standard organisms. It has been established that PCB-contaminated sediments represent a source of adverse effects on life functions of the used bioindicator L. minor. According to our results, the sediment from Zemplı´nska sˇı´rava was more toxic (20% higher inhibition of the growth rate) than that from Stra´ zsky canal despite the fact that PCB concentration in sediment sampled from Zemplı´nska sˇı´rava was lower than in sediment from Stra´ zsky canal. This finding indicated a need to perform further analyses of the sediments to explain the obtained
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445
200000
- PCB 101 A20.476 re a: 1.2 51 66 e+ 00 6
250000
- PCB 52 A16.175 re a: 72 66 00
- PCB 8 A10.114 re a: 1. 03 47 1e +0 06
300000
150000
- PCB 180 A29.057 re a: 1. 80 17 5e +0 06 31.104 PCB 203 Ar ea :8 1
14.454 - PCB 28 Ar ea :3 .4 51 06 e+ 00 6
350000
- PCB 118 A23.414 re a: 1. 8 - 3PCB 91 153 A24.575 re 7e a: +0 1. 06 77 48 - PCB 138 25.875 9e Ar ea +0 :1 06 .5 99 61 e+ 00 6
ECD1 B, (181005AM\008F1201.D) counts 400000
100000 50000 0
10
5
15
20
25
30
min
Ar
ea
14000 13000
- PCB 203 A31.103 re a: 53 86 .5 6
- PCB 180 A28.757 re a: 98 3. 05 8
15000
- PCB 118 A23.419 re a: 1-7 PCB 153 24.296 Ar 21 ea 4. 3 :1 18 60 .8 25.877 PCB 138 Ar ea :1 24 46 .3
16000
- PCB 101 A20.192 re a: 80 29 .2 8
17000
A16.390 : 1 Are rea 00 a: : 1 52 17 83 .7 07 33 3. .1 2
- PCB 8 A10.122 re a: 27 66 8
18000
14.461 Ar - PCB 2 ea :5 28 57 .4
ECD1 B, (191005AM\001F0101.D) counts
12000 11000 10
15
20
25
30
min
Fig. 2. Chromatogram records of sediments sampled from a) Stra´zsky canal, and b) Zemplı´nska sˇı´rava.
results in more detail. Toxicity of sediments can be caused by many other possible factors such as the presence of ammonium or sulfides, heavy metals, or pesticides. Besides, Zemplı´nska sˇı´rava being a lake represents stagnant water with higher probability of heavy and toxic metals, PAHs, and pesticides accumulation, in contrast with Stra´ zsky canal, which has flowing water. Our analyses really confirmed the presence of compounds mentioned above, fortunately below quantification limit (data not shown). Our previous results (Dercova´ et al., 2008) obtained using Microtox test (inhibition of bioluminiscence of standard bacterial strain Vibrio fischeri in the presence of toxic compound) demonstrated higher toxicity of the sediment sampled from Stra´ zsky canal than of that sampled from Zemplı´nska sˇı´rava. Therefore, our
Table 1 Detected amount of the indicator PCB congeners in the sediments sampled from Stra´ zsky canal and Zemplı´nska sˇı´rava water reservoir. Indicator PCB congener
Sk (mg kg1)
Zsˇ (mg kg1)
8 28 52 101 118 138 153 180 203 P
77.90 92.45 36.58 33.80 74.14 36.55 32.64 41.64 14.13 439.83
0.2083 0.1878 0.0263 0.0291 0.0443 0.0192 0.0225 0.0025 0.0069 0.5469
results confirmed that a battery of tests, especially at higher trophic levels is required to determine real toxicity of contaminated sediments. 3.2. Determination of mutagenicity of the commercial mixture of PCBs DELOR 103 Mutagenicity of contaminants is the property that may complicate the use of biological systems for environment decontamination due to their potential danger for biota. In this work, the commercial mixture of PCBs, DELOR 103, present in the tested sediments, was assessed directly for the mutagenic effect applying the Ames test using the standard strains S. typhimurium TA98 and TA100. Statistical evaluation was performed using Student’s t-test. The number of spontaneous revertant colonies in our experiment was comparable with that of the induced revertant colonies (Table 2). As the positive and the negative controls, nitrofurylacrylic acid (NAFA) or 1% solution of DMSO, respectively, were used. Addition of DELOR 103 did not affect the frequency of spontaneous mutations and no increase in the number of revertants was observed in the range of the concentrations used without metabolic activation. These results are in agreement with those published in the literature (EFSA, 2005). The genotoxicity test has therefore not proved the mutagenic effect of DELOR 103 toward the standard microorganisms S. typhimurium TA98 and TA100. The mutagenicity of both tested sediments was determined in our previous work (Dercova´ et al., 2008). The results did not confirm the genotoxic effect toward the standard microorganisms S. typhimurium TA97 and TA100 as well.
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Table 2 Number of the induced revertants at different concentrations of (the commercial mixture DELOR 103).
120
fronds number
100 80
40
0 0
2
4
7 (days)
120
fronds number
Number of revertants
Standard deviation
TA100 Number of revertants
Standard deviation
5 106 5 105 5 104 5 103 5 102 PC NC SR
27 32 28 23 30 456* 29 31
6.51 4.04 2.52 4.73 1.53 27.68 8.96 3.06
79 80 87 73 69 1693* 73 84
7.94 7.77 2.65 8.19 9.29 51.55 5.86 6.51
PC – positive control (NAFA: 3-(5-nitro-2-furyl)acrylic acid); NC – negative control (1% DMSO); SR – spontaneous revertants; *P < 0.001; data points represent three separate experiments, each run in triplicate.
100 80 60 40 20 0 0
2
4
7 (days)
120
fronds number
TA98
60
20
100 80 60 40 20 0 0
2
4
7 (days)
Fig. 3. Number of fronds at the beginning and on the 2nd, 4th, and 7th days for a) K2, , K3), b) sediment Zemplı´nska sˇı´rava ( ZS1, , ZS2, - ZS3, control (- K1, ZS4), and c) sediment Stra´zsky canal ( SK1, , SK2, - SK3, SK4).
3.3. Determination of the number of indigenous microorganisms in sediments Knowledge of the colonization density, i.e. cell counts is very important at the evaluation of viability of relevant sediment microorganisms. The observed number of colony forming units 50
40
effect (%)
DELOR 103 (mg ml1)
30
20
10
(CFUs) using the agarized nutritional broth was 6.8 104 per gram of dry sediment sampled from Stra´ zsky canal. This medium is more suitable for fast growing bacteria, whereas Thornton agar is more suitable for cultivation of soil and sediment bacteria since it ensures balanced growth of indigenous microorganisms. Although cultivation using Thornton agar takes longer time, it provides more precise values of the amount of microbial consortium present in the studied environmental sample. The number of native bacteria observed upon the growth on Thornton agar was higher reaching 5.2 105 per gram of dry sediment. At applying this indirect method we have probably obtained lower bacteria counts than by a direct method using water suspension of sediment and counting bacteria in a natural suspension using a microscope. We are aware of the fact that CFUs concern only cultivable microorganisms. It is known that only as little as 1% of soil microorganisms are cultivated and characterized. Torsvik and Ovreas (2002) established that the total bacteria counts determined by plating was 4 106 per gram of soil (dry weight), and 4 1010 established by using fluorescent microscopy. The difference between cultivable and non-cultivable bacteria represents 104 counts per gram of soil. It is possible to study identification, physiology, and genetics of non-cultivable organisms from sediments using a metagenomic approach by isolation of the total DNA from the contaminated soil or sediment (Dercova´ et al., 2008). 3.4. Determination of the growth inhibition of DELOR 103 toward the bacterial strains Prior to commencing biodegradation, it is very important to know inhibition concentration of PCBs toward the bacterial strains that will be inoculated into sediments for the purpose of biodegradation. Growth inhibition of A. xylosoxidans, P. stutzeri, and Burkholderia sp. LB400 in the presence of different concentrations of DELOR 103 (5, 10, 30, 50, 75, 100, 150, 200, and 250 mg l1; control sample was 2% DMSO without addition of DELOR 103) was determined. Cultivation was performed in the mineral medium containing glucose (2 g l1) on a rotary shaker (180 rpm) at 28 C in the dark. Toxicity of PCBs toward bacterial strains A. xylosoxidans, P. stutzeri, and Burkholderia sp. LB400 was evaluated. Determination of toxicity (parameter ID50) was based on inhibition of biomass growth in the presence of different concentrations of PCBs considered as growth inhibitors. The resistance of the used bacterial strains to DELOR 103 decreased in the following order: P. stutzeri > Burkholderia sp. LB400 > A. xylosoxidans (Fig. 5).
0 Ir
Ir
Ia
Ia (%)
Fig. 4. Average percentages of inhibition of growth rate Ir and inhibition of the amount ZS, of biomass Ia expressed as number of fronds or frond area for both sediments: , SK.
3.5. PCR amplification of bphA1 gene Our experiments were based on the fact that biphenyldioxygenase determines the specificity of PCB cleavage and that it is the
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447
0.2
cotton wool stopper
0.18 0.16
ID50 (g.l-1)
0.14
glass column
0.12 0.1 0.08
sorbent SILIPOR C18 glass frit
0.06 0.04 0.02
Erlenmeyer flask 0 Alcaligenes.x
Pseudomonas s.
Burkholderia sp.
Fig. 5. Inhibition concentration ID50 of DELOR 103 for the studied bacterial strains.
sediment mineral medium PCBs microorganisms
first enzyme of the biphenyl/PCB pathway. At amplification, the primers R350 and R674 that have been previously defined and verified were used (Rysˇlava´ et al., 2003). The results of amplification and identification of the bphA1 gene are presented in Fig. 6 illustrating the gel electrophoretic data. The strain Burkholderia sp. LB400 was used in this experiment as the positive control – as the strain possessing bphA1 gene. As can be seen, bphA1 gene, encoding
Fig. 7. Apparatus for monitoring of evaporation and biodegradation of PCBs (Dercova´ et al., 1996).
enzyme biphenyldioxygenase, responsible for the start of PCB degradation, is present in the isolate P. stutzeri. In another used bacterial isolate, A. xylosoxidans, bphA1 gene was not detected. ´ 3.6. Biodegradation of PCBs in the sediments sampled from Stra zsky canal This part of the work concerns biodegradation of PCBs in the samples of contaminated sediments under the laboratory conditions – in a natural sediment by an autochthonous microbial consortium and in a bioaugmented sediment inoculated with allochthonous bacterial strain. Since the sediment sampled from Stra´ zsky canal contained higher concentration of PCBs needed as the carbon and energy source, this sediment was used for biodegradation experiments. Biodegradation of PCBs was performed with the addition of biphenyl or glucose as the co-substrates. Number of CFUs of the indigenous soil microorganisms present in sediment sampled from Stra´ zsky canal was 5.2 105 in 1 g of the dried sediment. Concentration of inoculum of the bacterial isolates was 2 g l1. Biodegradation was performed in an apparatus (Fig. 7) that allows simultaneous monitoring of biodegradation and evaporation of PCBs as described previously (Dercova´ et al., 1996). The experiment was
Fig. 6. Gel electrophoretic proof (1.5% gel) of the presence of bphA1 gene in the DNA isolated from strains Pseudomonas stutzeri, Alcaligenes xylosoxidans, and Burkholderia sp. LB400.
Table 3 Extent of biodegradation after 14 days in sediment Stra´ zsky canal (in the presence of indigenous microorganisms) with addition of glucose (Glc) or biphenyl (Bph) inoculated with Alcaligenes xylosoxidans (A.x.), Pseudomonas stutzeri (P.s.), and Burkholderia sp. LB 400 (B.s.).
Gel description
Co-substrate
Microorganism
Biodegradation (%)
Glucose
Sediment þ A.x. Sediment þ B.s. Sediment þ P.s.
35 42 46
Biphenyl
Sediment þ A.x. Sediment þ B.s. Sediment þ P.s.
40 53 46
Lane Lane Lane Lane Lane Lane
A B C D E F
bphA1 gene Marker 100 bp P. stutzeri P. stutzeri A. xylosoxidans B. sp. LB400 Negative control
Yes Yes No Yes
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Table 4 The residual amount of the PCB congeners during biodegradation in sediment sampled from Stra´ zsky canal and inoculated under the laboratory conditions with Pseudomonas stutzeri (concentration of inoculum 2 g l1 added to 15 g of dried sediment). P Days Sediment þ PCB congener (mg/kg) PCBs (mg/kg) co-substrate þ MO 8 28 52 101 118 138 153 180 203 0 7 7 7 7 14 14 14 14
SED SED þ Glc SED þ Bph SED þ Glc þ P.s. SED þ Bph þ P.s. SED þ Glc SED þ Bph SED þ Glc þ P.s. SED þ Bph þ P.s.
77.90 9.33 10.80 18.89 9.97 10.52 8.44 12.45 11.23
92.45 46.19 48.52 63.14 49.85 49.72 48.37 59.58 30.88
36.58 17.21 18.29 24.82 20.56 10.80 18.69 11.88 13.09
33.80 17.40 18.25 23.44 21.89 20.85 23.42 25.60 26.90
74.14 41.30 43.60 51.10 45.47 51.74 53.04 62.34 37.34
36.55 19.81 21.14 25.45 23.37 25.54 25.69 27.26 32.47
32.64 21.30 18.46 23.63 26.51 30.30 23.36 26.37 29.03
41.64 22.66 20.34 30.25 26.42 29.65 28.67 29.97 35.19
14.13 5.81 7.30 10.23 8.63 10.41 10.44 10.70 12.79
439.83 201.01 206.70 270.95 232.67 239.23 240.33 265.98 228.92
SED – sediment with the presence of native microorganisms, P.s. – Pseudomonas stutzeri, PCB congeners: 28 (tri-CB), 52 (tetra-CB), 101 (penta-CB), 138 (hexa-CB), 153 (hexa-CB), 180 (hepta-CB), and 203 (octa-CB).
carried out on a rotary shaker (180 rpm) at 28 C in the dark. The initial concentration of PCBs (the total sum of the indicator PCB congeners) established in this sediment was 439.82 mg kg1. Bacterial strains A. xylosoxidans, P. stutzeri, and sp. LB400 (the control strain possessing bphA1 gene) were used separately as inoculum for bioaugmentation of the sediment. Biphenyl was added in small amount in the form of crystals. Water solubility of biphenyl is very low (7 mg l1). It can be assumed that in the process of bacterial degradation, biphenyl amount was continually replenished from the crystals into the liquid medium. Our results indicated that for all inoculated strains and also for native sediment biphenyl is a more suitable co-substrate for PCB biodegradation than glucose. About 5% of the initial amount of PCBs was estimated as that evaporated and sorbed on sorbent Silipor in bioaugmented and non-bioaugmented sediments as well. Biodegradation of PCBs in the bioaugmented sediments with addition of biphenyl increased in the following order: A. xylosoxidans < P. stutzeri < Burkholderia sp. LB400 (Table 3). The amounts of the individual PCB congeners estimated during biodegradation of PCBs in the presence of P. stutzeri are presented in Table 4. As can be seen, the most abundant PCB congeners at the beginning of experiments were PCB congener 8 (mono-CB), 28 (triCB), and 118 (penta-CB). The lowest abundance was determined for PCB congener 203 (octa-CB). The most significant decrease throughout biodegradation was observed predominantly with the lower chlorinated PCB congeners 8, 28, and 52. With other PCB congeners, the decrease was not so pronounced. The amount of PCB congener 203 slightly decreased after the first seven days and then remained at the constant level. The effect of bioaugmentation was more pronounced after 14 days of degradation probably due to adaptation of bacterial inoculum in the contaminated sediment. It was recognized that certain organic compounds, including plant terpenoids, stimulate PCB degradation by microorganisms in some environments. However, the usefulness of these amendments for improving PCB removal by microorganisms, described also in our previous paper (Tandlich et al., 2001) has not been sufficiently explored (Hernandez et al., 1997; Gilbert and Crowley, 1997; Gilbert and Crowley, 1998; Rhodes et al., 2007). Effects of selected organic compounds on aerobic removal of PCBs and on bacterial community composition in sediments were studied also by Luo et al. (2007). Their results obtained with the amendment of soil and sediments with biphenyl, glucose, and salicylic acid indicated that addition of organic matter had site-specific effects on bacterial populations and PCB removal. In our work, the highest PCB degradation under the bioaugmentation conditions was observed in the sediment Stra´ zsky canal inoculated with the bacterial strain P. stutzeri. Addition of biphenyl as the co-substrate and the inducer resulted in higher
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