Long-Term Anaerobic Mineralization of ... - ACS Publications

EcoTopia Science Institute, Nagoya University, Chikusa, Nagoya 464-8603 Japan. ‡ Department of Civil Engineering, Graduate School of Engineering, Na...
0 downloads 0 Views 873KB Size

Article pubs.acs.org/est

Long-Term Anaerobic Mineralization of Pentachlorophenol in a Continuous-Flow System Using Only Lactate as an External Nutrient Zhiling Li,†,‡ Yasushi Inoue,† Daisuke Suzuki,† Lizhen Ye,† and Arata Katayama*,†,‡ †

EcoTopia Science Institute, Nagoya University, Chikusa, Nagoya 464-8603 Japan Department of Civil Engineering, Graduate School of Engineering, Nagoya University, Chikusa, Nagoya 464-8603 Japan

S Supporting Information *

ABSTRACT: A simple anaerobic upflow column system (15 cm long, 5 cm inner diameter) for complete pentachlorophenol (PCP) mineralization has been established using a microbial consortium requiring only lactate as the external nutrient. With lactate as an electron donor, PCP was dechlorinated to 3chlorophenol (3-CP) and phenol. The degradation of 3-CP and phenol proceeded without an external electron acceptor, indicating fermentative or syntrophic characteristics. A tracer experiment using 14C−U-ring-labeled PCP confirmed the conversion of PCP into CO2 (54.1%), CH4 (48.1%), and biomass (0.6%). The nitrogen required for degradation was supplied by N2-fixation, evidenced from the nitrogen balance and an acetylene reduction assay. A 16S rRNA gene library analysis showed that bottom of the upflow column harbored the potential dechlorinators, Dehalobacter and Desulf itobacterium, and the phenol/3-CP fermentative or syntrophic degraders, Cryptanaerobacter and Syntrophus. The nitrogen-fixing facultative anaerobes, Rhizobiales, were detected in the top of the upflow column, with other possible nitrogen-fixers at both bottom and top of the upflow column. The mineralization rate reached 1.96 μmoles L−1 d−1 for 50 μM of the initial PCP concentration: one of the highest efficiencies reported. This compact anaerobic mineralization system requiring no external supply of an electron acceptor would be useful for the remediation of chlorinated aromatic compounds under anaerobic conditions.

nols.7−9 Regarding in situ application, such systems require a large amount of energy and money (1) to make the anaerobic conditions preferable for reductive dechlorination and (2) then in changing to aerobic conditions to permit the oxidative degradation of aromatic compounds.10 Therefore, these systems have been considered unrealistic for in situ commercial applications. For chlorophenols, we have proposed a microbial remediation technology using an anaerobic dechlorinator and anaerobic oxidizer in combination and achieved the complete mineralization of PCP under anaerobic conditions, in both batch11 and continuous-flow systems.10 This anaerobic− anaerobic microbial combination has provided a more plausible approach for the large-scale applications of anaerobic-permeable reactive biobarriers. However, because the anaerobic−anaerobic microbial combination goes through two distinct bioprocesses, reductive dechlorination and anaerobic oxidation, multiple microbial consortia are involved and various nutrients are required, which increases the cost as well as in situ operational difficulties. Several studies have reported anaerobic PCP mineralization in acclimated sludge or granules12−14 or in an upflow anaerobic

INTRODUCTION Halogenated aromatic compounds are persistent and exist widely in the environment with high toxicity and bioaccumulation.1 During their extensive utilization since the 1920s, there has been serious contamination to underground zones including the soil, aquifers, and groundwater through tank leakages, accidental spills, and illegal dumping. The chlorinated phenols such as pentachlorophenol (PCP) are a typical group of the halogenated aromatic organics and several chlorophenols are listed as priority pollutants by the US Environmental Protection Agency.2 In situ microbial bioremediation is now considered to be a cost-effective and promising technology and many studies on in situ anaerobic remediation have focused on the reductive dechlorination of chloroethanes or chloroethenes.3−5 However, the reductive dechlorination process does not meet the requirements for the remediation of halogenated aromatic compounds, because the less-chlorinated aromatic intermediates formed are still toxic and an environmental burden.6 Therefore, a subsequent treatment is necessary for complete mineralization. In the remediation of chlorinated aromatic compounds such as PCP, studies have shown that further degradation of less chlorinated or nonchlorinated aromatic compounds after reductive dechlorination is conducted more easily under aerobic conditions. Therefore, a combined system composed of anaerobic dechlorination and aerobic degradation has been proposed to achieve the complete mineralization of chlorophe© 2012 American Chemical Society

Received: Revised: Accepted: Published: 1534

September 29, 2012 December 12, 2012 December 19, 2012 December 19, 2012 dx.doi.org/10.1021/es303784f | Environ. Sci. Technol. 2013, 47, 1534−1541

Environmental Science & Technology


field biobarrier system. The top cap of the column had a 4.8-cm diameter opening with a butyl rubber stopper for sampling. The column was packed with 450 g of glass beads (mean diameter 2 mm) and 12 g (dry weight) of the anaerobic paddy field soil slurry using autoclaved distilled water. The porosity of the column was 35% and one pore volume (PV) was 110 mL, according to Li et al.22 The number of PVs, calculated by dividing the cumulative effluent volume with one pore volume (110 mL) of the column, was used as an indicator of the time taken. The influent solution, consisting of different concentrations of PCP− Na and lactate−Na in N2-saturated distilled water, was introduced from a glass bottle (600 mL) through a plunger pump in the upflow mode. The bottle was connected to a gastight 1-L aluminum bag filled with N2 gas to replenish the influent solution. The pH values of the paddy soil slurry and the influent solution were adjusted to 7.0. The column and the influent bottle were installed in a dark incubator with the constant temperature of 30 °C to facilitate the microbial growth.16 At the start of each experiment, 15 mL of the PCPdechlorinating consortium was introduced from the bottom of the column and then the column was immediately supplied with 50 μM of PCP and 20 mM of lactate as the influent without any recirculation. The flow rate was maintained at 0.01 mL min−1, corresponding to a hydraulic retention time (HRT) of 7.6 d for 10.1 PVs (77 d in total). Lactate, as the essential electron donor for dechlorination, gave the highest dechlorination activity at 20 mM.16 After steady rates for dechlorination and phenol degradation had been obtained, the flow rate of the system was decreased to 0.003 mL min−1 (HRT = 25.3 d) to allow for the complete mineralization of PCP by the consortium developed inside the column. The activity of anaerobic PCP mineralization in the system was maintained steadily until the column was disintegrated (at 41.4 PVs) (Phase II). Each column system experiment was carried out three times to confirm the reproducibility of the results. Chemical Analysis of Organics, Anions, and Gases. Aromatic compounds (PCP and the metabolites) were determined with GC-MS as described in the Supporting Information (SI). Concentration of organic acids (lactate, acetate, and propionate), Cl− in solution and effluent gases (H2, CO2, and CH4) in both gas and liquid phases were determined as described previously.22 The total carbon and inorganic carbon of the effluent was determined using a TOC5050A analyzer (TOC-VCPH, Shimadzu Corporation, Kyoto, Japan) using samples obtained with a gastight syringe (SigmaAldrich, St. Louis, MO, USA) and immediately alkalized with 10 M NaOH. For measuring nitrogen content, 15 mL of the influent and effluent samples were digested at 380 °C for 2 h using a DK20 heating digester (VELP Scientifica, Milan, Italy) with a digestion solution (3.28 g L−1 of Na2SeO3 in a mixture of H2SO4 and H3PO4) and H2O2 (30%). The concentration of nitrogen in the digests was determined using the phenol-hypochlorite colorimetric method.23 14 C−PCP Tracer Experiment. The influent solution (400 mL) was spiked with 14C-ring (U)−PCP (99% purity, 80 mCi mmol−1; Sigma-Aldrich) as 100 μL of acetone/toluene solution. The influent solution consisting of 8.48 ± 1.91 × 10 4 disintegrations per minute (dpm) mL−1 of 50 μM of the radiolabeled PCP and cold Na−lactate was introduced continuously into the column for 20.0−23.5 PVs. Effluent samples (2 mL) were taken with a gastight syringe from an opening on the top cap and immediately introduced into sealed N2-filled 15-mL bottles. The sealed bottles were flushed

sludge blanket (UASB) reactor.15 However, these systems either had lower mineralization rates12,13 or required restricted conditions such as the presence of sludge or a mixture of carbon sources to maintain microbial activities.14,15 Our previous study10 documented an anaerobic PCP mineralization system with high mineralization efficiency by combining a reductive PCP-dechlorinating microbial consortium with an anaerobic iron-reducing phenol-degrading consortium under continuous flow conditions. Lactate, supplied as the electron donor for dechlorination, was converted into acetate and propionate.16 This resulted in the unwanted biochemical consumption of FeOOH, supplied as the electron acceptor for phenol degradation, which resulted in the shortened longevity of the system activity.10 However, it was found that phenol fermented in the methanogenic environment utilizing the intrinsic electron acceptor CO217 or was degraded syntrophically, associated with H2-scavenging methanogens, without the need for an additional electron acceptor.18−20 These indicate that the utilization of a fermenting or syntrophic phenol degrader, in combination with a reductive PCP dechlorinator, would make it possible to construct a mineralization system supplied with only an electron donor and would avoid an unwanted biochemical reaction between the electron donor and the electron acceptor introduced externally. In this study, we have succeeded in constructing a complete anaerobic PCP mineralization system under continuous flow conditions by supplying only lactate externally as the electron donor. Biomineralization was demonstrated by a tracer study using 14C−U-ring-labeled PCP and the system was characterized based on the material output balance and microbial community analysis.

EXPERIMENTAL SECTION Soil, Chemicals, and Microorganisms. The soil used was obtained from the subsurface horizon of a paddy field in Yatomi City, Aichi Prefecture, Japan and was immersed in tap water in a sealed plastic bag to maintain the original paddy field environment. The composition of tap water was the same as that of paddy water, except for the addition of hypochlorite and tap water was used after waiting for a while to allow active chlorine to diminish. The chemical composition of the soil on a dry weight basis was 16.4 mg g−1 of C, 8 mg g−1 of H, 2.1 mg g−1 of N, 1.1 mg g−1 of K, 0.2 mg g−1 of Na, 1.3 mg g−1 of Ca, 1.2 mg g−1 of Mg, and 0.31 mg g−1 of P. The C, H, and N contents were analyzed using a CHN CORDER-MT-5 with an MTA-620 autosampler (Yanaco Analyzing Industry, Tokyo, Japan). The metal contents were determined using an ICP-AES optima 3300DV (PerkinElmer Corporation, Santa Clara, CA, USA). The total P content was determined using anti−Mo-Sb spectrophotometry after digestion of the soil with concentrated H2SO4 and HClO4.21 PCP−Na (90% purity) and 3-monochlorophenol (3-CP) (97% purity) were purchased from Wako Pure Chemical Industries (Osaka, Japan). Phenol (99% purity) was purchased from Katayama Chemical Industries (Osaka, Japan). The anaerobic consortium dechlorinating PCP to phenol has been maintained in our laboratory. The consortium has an average microbial density of (1.9 ± 0.3) × 1010 cells mL−1 and consists predominantly of the phyla of Firmicutes and Spirochaetes and an unidentified group of the Archaea.16 Anaerobic Continuous Flow System. The setup of the anaerobic system and the column packing procedures has been described by Li et al.22 Briefly, a glass column (15 cm long, 5 cm inner diameter) with stainless steel caps was used to simulate the 1535

dx.doi.org/10.1021/es303784f | Environ. Sci. Technol. 2013, 47, 1534−1541

Environmental Science & Technology


with pure O2 with a maximum flow of 10 mL min−1 after acidification and the gases in the bottles were separately trapped by a specially made gastight apparatus. The 14CH4 in the bottles was converted into and trapped as 14CO2 by a CuO column treatment at 850 °C. The apparatus and determination methods were according to the method of Yang et al.,11 except for the modifications of the sample volume (2 mL), trapping time of 14 CO2 and 14CH4 for 1 h, and the O2 flow rate at 10 mL min−1. The recoveries of 14CO2 and 14CH4 in this system were determined as 85.6% and 73.1%, using 13C−Na2CO3 and 13C− H4 as the standards, respectively. The 14C aromatic metabolites were examined in an ethylacetate extract from the samples using thin layer chromatography on Silicagel 70 plates (Wako) and a developing solvent of nhexane-ethyl acetate (80:20). The retention factor (Rf) values of PCP, 3-CP, phenol, benzoate and 4-hydroxybenzoic acid (4OHB) were 0.12, 0.20, 0.18, 0.15, and 0.16, respectively. The metabolites were visualized by exposure to iodine vapor, scraped off the plates, and subjected to radioactivity analysis. The radioactive biomass was determined by the protein analysis in the effluent samples (2 mL), obtained by centrifugation at 15 000g after mixing with acetone (−20 °C) for 30 min. The 14C-biomass was calculated from the radioactivity of the protein according to the carbon distribution ratio in the cell (C-protein:C-cell = 0.24:1).24 All the 14C radioactivity values of the samples were determined with a liquid scintillation counter (LSC 5100, Aloka, Tokyo, Japan) using a liquid scintillation cocktail (Ecoscint Ultra, National Diagnostics, Atlanta, GA, USA). Analysis of Microbial Communities in the Column. At 19.3 PVs, the top and bottom column caps were temporarily opened under an atmosphere of gaseous dry ice (CO2) to prevent O2 entry, and samples (a mixture of glass beads, soil, and liquids) were taken by gentle scraping using autoclaved scoops. After the total cell numbers had been counted by epifluorescence microscopy after staining with 4′,6-diamidino-2-phenylindole (Invitrogen Co., Carlsbad, CA, USA), DNA was extracted with a DNA extraction kit, Isoplant II (Nippon Gene Co., Ltd., Toyama, Japan). The microbial community structure was determined by constructing the full-length 16S rRNA gene clone library. The library construction and sequencing procedures are described in the SI. The obtained chimera-free nucleotide sequences were aligned using BioEdit Sequence Alignment Editor (, Ibis Biosciences, Carlsbad, CA, USA) and classified into different operational taxonomic units (OTUs) at a level of sequence similarity of more than 97%. The OTU sequences were analyzed with RDP classifier (http://rdp.cme. msu.edu/classifier/classifier.jsp) and compared with those in the GenBank nucleotide sequence database using the BLAST program (http://blast.ncbi.nlm.nih.gov/Blast.cgi). The nucleotide sequences obtained with the 16S rRNA gene library have been deposited in the DNA Data Bank of Japan nucleotide sequence databases under accession numbers from AB723828 to AB723857. The pylogenetic tree with OTU sequence and the closest related type strain was constructed with the software of MEGA 5.05 (http://www.megasoftware.net) using neighborjoining algorithm incorporating Kimura 2-parameter distance correction. Short-length 16S rRNA genes were also examined to characterize Archaeal and Bacterial community structures using PCR and denaturing gradient gel electrophoresis (PCR− DGGE) following the nucleotide sequencing of the gel bands. The PCR−DGGE procedure is described in the SI.

Degradation Activity of the Consortium and Acetylene Reduction Assay. Degradations of phenol and 3-CP without an additional electron acceptor were confirmed in the batch incubation. Anaerobic medium (40 mL) in batch bottles (68 mL), composed of mineral salts without nitrogenous compounds or trace elements,25 10 mg L−1 of sodium resazurin, and 10 mM Na2S, was prepared by flushing with pure N2 gas for 2 h, sealed, and autoclaved at 120 °C for 30 min. After the bottles had cooled, phenol (20 μM) and 3-CP (5 μM) were added and the headspace gas was changed to N2 (80%) and CO2 (20%). Inocula (2.5 mL) from the top and bottom of the column were introduced into the bottles. After incubation at 30 °C in the dark, the remaining phenol and 3-CP concentrations were determined. An acetylene reduction assay (ARA) was performed on the medium in the bottles incubated with the active degradation inocula. After phenol/3-CP had degraded completely, 5 mL of the culture was transferred into two bottles: one fed with 20 μM of phenol and 5 μM of 3-CP and the other with no feed control. Then, 2 mL of acetylene gas (purity >99.99%, Sogo Kariya Sanso Corporation, Nagoya, Japan) was introduced into the headspace (20 mL) of each bottle and the bottles were vented using a needle to reach equilibrium to atmospheric pressure. The ethylene concentration in the headspace was determined at appropriate intervals using a gas chromatograph as described in the SI. The ethylene concentration in the bottles was calculated as the sum of that in the headspace and in the liquid, using an Ostwald constant of 0.1068 at 30 °C under ambient pressure.26

RESULTS Column Performance at an HRT of 7.6 d (Phase I). In the beginning of phase I of the study (0.6 PVs, 4.6 d), 11 μM of PCP was detected in the effluent, but disappeared completely at 7.0 PVs (53 d; Figure 1A). The concentration of PCP metabolites, phenol, and 3-CP in the effluent stabilized after 8.0 PVs (61 d). During PVs of 8.0−10.1 (61−77 d), the PCP (50 μM) was steadily dechlorinated into phenol (∼15 μM) and 3-CP (∼5 μM) as the final metabolites; no other intermediate chlorinated phenol was detected. The concentration of Cl− produced was maintained at around 248 μM (Figure 1B), which agreed with the stoichiometric Cl− concentration (245 μM) when the chlorine in 50 μM of PCP was completely dechlorinated into Cl−, except for the chlorine in the 3-CP. In the steady state, the total concentration of the aromatic metabolites (phenol and 3-CP) was maintained at about 20 μM (Figure 1A), which was significantly lower than the concentration in the influent (50 μM). This indicated that part of the phenol and 3-CP had been degraded further by the consortium, which would have been enriched in the column during this period. During PVs of 0−6 (46 d), only a part of the lactate was utilized and about 5 mM of lactate flowed out of the column (Figure 1C). After 6.2 PVs (47 d), the lactate had decomposed completely. Acetate (8 mM) and propionate (8 mM) were the main organic acid metabolites and no formate was detected in the effluent. Dissolved H2 and CO2 gases flowed out from the column and a low concentration of CH4 was detected (Figure 1D). Column Performance with an HRT of 25.3 d (Phase II). After 10.1 PVs, the column flow rate was reduced to 0.003 mL min−1, which corresponded to an HRT value of 25.3 d. During phase II, PCP was steadily dechlorinated and the produced Cl− was kept to a stable level (Figure 1A, B). The concentrations of phenol and 3-CP gradually decreased in the effluent in the beginning of phase II and disappeared after 11.2 PVs (105 d) and 1536

dx.doi.org/10.1021/es303784f | Environ. Sci. Technol. 2013, 47, 1534−1541

Environmental Science & Technology


accompanied by a decrease in the concentrations of H2 and CO2, which was probably caused by the activity of H2-scavenging methanogens. After 17.0 PVs, the mineralization system became stably and durably maintained until the column was disintegrated (at 41.4 PVs) (data not shown). 14 C−PCP Mineralization in the Column. After the introduction of 14C−PCP in the influent from 20.0 PVs, the 14 C-radioactivity in the effluent was detected after 21.4 PVs and became stable after 22.0 PVs. Table 1 shows the 14C-metabolites during PVs of 22.1−23.5. The 14C-radioactivity was mainly distributed on 14CO2 (54.1%) and 14CH4 (48.1%), with a trace amount of 14C-biomass (0.6%). No 14C-radioactivity was detected in PCP, 3-CP, phenol, benzoate, 4-OHB, or hydrophobic vapors. This result demonstrated the complete mineralization of PCP in this anaerobic system. Carbon and Nitrogen Balances between the Influent and Effluent. To further evaluate the mineralization system, carbon balance was examined between the influents and effluents at the steady state of phase I (8.0−10.0 PVs) and phase II (17.0− 19.0 PVs) (SI Table S1). During phase I, the organic carbon in the influent was mainly distributed as propionate and acetate, and as CO2 (or carbonate) in the effluent, while CH4, phenol, and 3CP occupied smaller proportions. The following empirical reactions represent the metabolite distributions during phase I (the number denotes the average concentration proportion for each chemical): 21 lactate +0.05 PCP = 0.7 lactate + 7 acetate + 8 propionate + 0.002 (3-CP and phenol) + 18 (CO2 + carbonate) + 0.18 CH4 + 7 biomass-C (104% of recovery). The impurities in the PCP−Na reagent were not included in the mass balance calculation because the concentration of impurities in 50 μM of PCP−Na (10%), calculated as 1.6 mg L−1, was negligible compared with 20 mM of lactate−Na (1800 mg L−1). During phase II when PCP was mineralized, the organic carbon in the influent was mainly converted into CO2 (or carbonate), CH4, and biomass. In the same manner as phase I, the following empirical reaction represents the metabolite distribution of phase II: 21 lactate + 0.05 PCP = 53 (CO2 + carbonate) + 0.4 CH4 + 7 biomass-C, (96.1% of recovery). At 18.5 PVs, the concentration of total carbon in the influent was 60.5 mM and that in the effluent was 59.6 mM, in which the inorganic carbon concentration was 53.1 mM. During the complete PCP mineralization in phase II, the organic nitrogen concentrations in the influent and the effluent were 0.01 mM and 0.99 ± 0.23 mM, respectively, suggesting a significant amount of organic nitrogen production inside the column. Since nitrogen was only supplied as the form of dissolved N2 in the influent, N2 fixation activity appeared to have occurred in the column. Microbial Community at the Bottom of the Column. At 19.3 PVs, the concentration of microbial cells at the bottom was 2.7 × 1.2 × 109 cells mL−1. The full-length 16S rRNA gene libraries revealed that the microbial community consisted of three phyla, Chloroflexi (12.9%), Firmicutes (31.9%), and Proteobacteria (35.3%), at the bottom of the upflow column (SI Table S2). As shown in Figure 2, Dehalobacter sp. and Desulf itobacterium sp. were detected as the potential dechlorinators, with the closest relative of Dehalobacter restrictus strain PER-K23 with dechlorination activity for chloroethenes27 and Desulf itobacterium f rappieri PCP-1 for chlorophenols at the ortho and meta positions,28 respectively. In addition, Cryptanaerobacter sp. was detected, with the closest relative Cryptanaerobacter phenolicus strain LR7.2, an anaerobic degrader of phenol and 4-OHB without utilizing the additional electron

Figure 1. Concentrations of metabolites in the column effluent with pentachlorophenol (PCP) and lactate serving as the influent at a hydraulic retention time of 7.6 d during pore volumes (PVs) of 0−10.1 (Phase I, the left-hand part of the vertical dashed line) and 25.3 d during PVs of 10.1−19.3 (Phase II, the right-hand part of the vertical dashed line). (A) Concentrations of PCP in the influent and of aromatic metabolites in the effluent. (B) Concentration of Cl− in the influent and effluent. (C) Concentrations of lactate in the influent and organic acid metabolites in the effluent. (D) Dissolved concentrations of H2, CH4, and CO2 in the effluent. The horizontal dashed line denotes the stoichiometric Cl− produced by 50 μM of PCP on complete dechlorination. Number of pore volumes (PVs), as an indication of the time taken, was calculated by dividing the cumulative effluent volume with one pore volume (110 mL) in the column. All data are the means of two determinations, and representative results of three column experiments are shown.

13.6 PVs (166 d), respectively (Figure 1A). Because no electron acceptor was supplied in the influent, the anaerobic degradation of phenol and 3-CP was considered to be achieved without external electron acceptor. Complete mineralization of PCP was maintained after 14 PVs (176 d). The concentration of acetate and propionate in the effluent, as the metabolites of lactate, also decreased gradually and reached negligible levels at 11.1 PVs (102 d; Figure 1C). As there was no other organic acid detected, lactate was regarded to have been mineralized completely. Further fermentation of acetate and propionate was suggested by the increase in concentration of H2 and CO2 at PVs of 12.4−14.7 (135−193 d) (Figure 1D). After 14.7 PVs (193 d), the concentration of CH4 increased, 1537

dx.doi.org/10.1021/es303784f | Environ. Sci. Technol. 2013, 47, 1534−1541

Environmental Science & Technology


Table 1. Distribution (%) of 14C on CH4, CO2, and Biomass in Effluent of the Column During Pore Volumes (PVs) of 22.1−23.5 after 14C−PCP was Introduced to the Column Influent at 20.0 PV effluent metabolitesb influent


gas phase


liquid phase 14



concentration ( × 104 dpm mL−1) 14 C distribution percentage (%)d

8.5 ± 1.9



4.6 ± 1.3 54.1



C hydrophobic vapors

4.1 ± 0.7 48.1

N.D.f 0



C aromatic compounds N.D. 0


C biomass

0.05 ± 0.03 0.6

total recovery rate (%)e 102.8


C−PCP radioactivity in the influent was the mean result from three parallel samplings with the standard deviation. Radioactivity in the effluent metabolites was the mean result determined at the four sampling times, PVs of 22.3, 22.6, 23.1, and 23.5, with standard deviation. cRadioactivities of 14 CO2 and 14CH4 were determined by dividing the determined radioactivity with recovery rate of 85.6% and 73.1% for CH4 and CO2, respectively. d Distribution percentage (%) in the effluent metabolites was calculated from dividing the determined radioactivity by that of 14C−PCP in the influent. eTotal recovery rate was calculated by dividing the sum of radioactivity in the effluent metabolites by that of 14C−PCP in the influent. fN.D. denotes “not detected” where the concentration of radioactivity was less than the detection limit, 36 dpm mL−1.

Figure 2. Phylogenetic tree of the 16S rRNA gene sequences of the dechlorinating, the syntrophic/fermentative degrading and nitrogen fixing populations, obtained from the clone libraries on the bottom and the top of the column, as well as their closest related type strains (with a similarity of ≥97%). The branch in bold is the operational taxonomic unit (OTU), with the appearance frequency in brackets on the bottom of the column (% of B) and the top of the column (% of T), respectively. The detailed phylogenetic information is shown in Table S2 in the Supporting Information. The numbers at the nodes indicate the percentages of times that nodes appeared in 1000 bootstrap analyses. The scale bar represents an estimated difference of 10% in nucleotide sequence positions. Dehalococcoides mccartyi strain 195 was used as the out-group.

acceptors, such as SO42−, S2O32−, NO3−, or FeCl3.29 One OTU was assigned to the genus Syntrophus (Figure 2) and the closest relative was Syntrophus gentianae strain DSM 8423, which can degrade benzoate into acetate and CO2 syntrophically in the presence of hydrogen-utilizing partner bacteria or methanogens.30 The PCR−DGGE patterns of the partial 16S rRNA genes of members of the Bacteria and Archaea at the bottom of the column were similar to those of the original PCP-dechlorinating consortium (SI Figure S1). The detected Archaea were classified into the phylum Euryarchaeota, with their closest relatives, Methanothermobacter wolfeii strain DSM and Methanoregula

formicicum strain SMSP, both of which are H2-scavenging methanogens (SI Table S3).31,32 Microbial Community at Top of the Column. At 19.3 PVs, the concentration of microbial cells at the top of the upflow column was 2.3 ± 2.5 × 108 cells mL−1, one-tenth of that at the bottom. The 16S rRNA gene library revealed that the microbial community consisted of four phyla: Bacteroidetes (18.7%), Chloroflexi (7.7%), Firmicutes (18.6%), and Proteobacteria (55%) (SI Table S2). Members of the Bacteroidetes appeared at the top of the column. The number and diversity of clones classified into the Proteobacteria increased. As shown in Figure 2, a Sulf urospirillum sp. of the ε-Proteobacteria was detected as the potential dehalogenator, and has been reported as a 1538

dx.doi.org/10.1021/es303784f | Environ. Sci. Technol. 2013, 47, 1534−1541

Environmental Science & Technology


dehalogenator of polybrominated diphenyl ethers.33 Three nitrogen-fixing Rhizobiales among the class α-Proteobacteria appeared at the top of the column. The closest relatives, Bradyrhizobium betae strain PL7HG1, Azorhizobium caulinodans ORS 571 and Rhizobium endophyticum strain CCGE2052, are all nitrogen-fixing facultative anaerobes.34−36 The genera Desulfovibrio, and Desulf itobacterium also existed in this system (Figure 2) and these are reported to possess the nitrogenase operon and can conduct nitrogen fixation.37,38 Desulfovibrio sp. was detected both at the bottom and top of the column. This high proportion suggested a role as a lactate fermenter as reported,39 although dechlorination ability for chloroaromatic compounds has also been reported.40 The PCR−DGGE pattern of partial 16S rRNA genes of the phylum Bacteria differed from those at the bottom and the original introduced culture. No member of the Archaea was detected in the PCR−DGGE analysis of samples at the top of the column. Microbial Community in the Column Effluents. The mean concentrations of cells in the effluent were (8.5 ± 3.6) × 107 cells mL−1 at PVs of 8.0−10.0 (61−76 d) in phase I and (9.2 ± 2.3) × 107 cells mL−1 at PVs of 12.0−18.0 (125−277 d) in phase II, respectively, which are 2 orders of magnitude lower than the cell concentration at the bottom of the column. The PCR−DGGE analysis of Bacterial 16S rRNA genes showed differences in community structure between phase I (HRT = 7.6 d) and phase II (HRT = 25.3 d) (SI Figure S2; SI Table S4). Major fragments E1 and E2 (SI Figure S2) in the effluent of phase I disappeared in the effluent of phase II. Bands E5, E6, E7, E8, and E9 became the major bands in phase II. No member of the Archaea was detected in the effluent during phases I or II. Performance in the Incubated Batch Culture and ARA Study. In the batch medium without additional electron acceptor, phenol began to decrease after 5 d of inoculation of consortium from the column and disappeared after 22 d (Figure 3A). The 3-CP degradation began after the phenol was exhausted and was completely degraded after 35 d. In the ARA study, the reduction of C2H2 to C2H4 occurred after 10 h in the bottles fed

with phenol and 3-CP, but none in the bottles without them (Figure 3B). This indicated that the nitrogen fixing activity existed in the inoculum from the column, utilizing phenol/3-CP as the carbon source.

DISCUSSION Our study provided an anaerobic PCP mineralization system with the technological merit of only supplying lactate as the external nutrient. Anaerobic PCP mineralization was tracked by 14 C tracer measurements under continuous flow conditions (Table 1). The use of only lactate in the mineralization process simplifies the in situ application difficulties caused by the operational complexity of microbial combination systems with various nutrient supplements. Our system had an anaerobic PCP mineralization rate of 1.96 μmoles L−1 d−1, one of the highest efficiencies reported to date under continuous flow conditions. This was calculated by multiplying the subtractive result of mineralized concentration of PCP between influent and effluent (50 μM) by flow rate (0.003 mL min−1) and dividing by the liquid of system volume (110 mL). This calculation was applied to the continuous-flow system reported previously for the comparison of PCP dechlorination rate: 0.13 μmoles L−1 d−1 in an UASB reactor15 and 3.46 μmoles L−1 d−1 in two combined columns in series.10 It should be noted that the introduced PCP and lactate in our system were transformed into biologically innocuous compounds (Cl−, CO2, CH4 and H2) or microbial biomass in the effluent (SI Table S1). This system would not bring secondary contamination to the areas of application. The flow rate in phase II corresponded to a linear velocity of 0.7 × 10−5 cm s−1, suggesting a potential for this system to be employed in situ with a flow rate in the order of 10−5 cm s−1. Studies using the full-length 16S rRNA gene library suggested the presence of potential PCP dechlorinators: Dehalobacter and Desulf itobacterium detected at the bottom of the upflow column and Sulf urospirillum sp. at the top of the column. The ten times higher population at the bottom of the column suggested that these microorganisms dechlorinated PCP mainly in the bottom of the upflow column. Microbial degraders of phenol and 3-CP were developed by the continuous flow system and by prolongation of the HRT (Figure 1). These did not require an external electron acceptor. The degraders probably utilized metabolites of lactate and/or PCP, such as CO2, as the internal electron acceptor, or were syntrophically associated with H2-scavenging methanogens. The 16S rRNA gene library study suggested the presence of bacteria with the characteristics, Cryptanaerobacter and Syntrophus, at the bottom of the upflow column (Figure 2). The PCR−DGGE analysis also supported the existence of H 2-scavenging methanogens at the bottom of the column (SI Figure S1 and Table S3). It should be noted that neither a potential phenol/3CP degrader nor a methanogen was found at the top of the column (Figure 2; SI Figure S1 and Table S3). The results indicated that the degradation of phenol and 3-CP occurred mainly in the bottom of the upflow column as well as reductive dechlorination. No nitrogenous compound was supplied in the system, whereas a significant increase of organic nitrogen concentration was observed in the effluent (0.98 mM). The ARA study on the batch culture confirmed nitrogen fixation activity in the presence of phenol and 3-CP. Nitrogen-fixing bacteria were detected in the top of the upflow column: Bradyrhizobium, Azorhizobium, and Rhizobium (Figure 2). In addition, the genera Desulfovibrio, and Desulf itobacterium present widely in the column (Figure 2) were

Figure 3. (A) Degradation of phenol and 3-CP without an additional electron acceptor in the batch bottles with NH4+-free medium inoculated from the column. (B) Ethylene produced by the acetylene reduction assay in the batch bottle. 1539

dx.doi.org/10.1021/es303784f | Environ. Sci. Technol. 2013, 47, 1534−1541

Environmental Science & Technology


possible nitrogen fixers.38 Thus, the dissolved N2 in the influent appeared to be fixed into organic nitrogen. These findings signified that there was no need for supplemental nitrogenous compounds during the in situ application. Calculation of the electron balance in the unit volume of the liquid phase (1 L) of the column suggested that there was an excess of H2 in this mineralization system during phase II. Lactate was initially decomposed to H2, acetate, and propionate. Then the acetate and propionate were decomposed further into H2 and CO2 (Figure 1C, D).22 Theoretically, 20 mmol of lactate can release 120 mmol of H2 at most (lactate + 3H2O ⇒ 3CO2 + 6H2).41 While, 0.25 mmol of the released H2 would be utilized for the complete dechlorination of 50 μmoles of PCP,42 because two electrons are required for the removal of each chlorine atom. The synthesis of CH4 at 0.5 mM observed in the effluent suggested the putative consumption of 2 mmol of H2 by hydrogenotrophic methanogens (4H2 + HCO3− + H+ ⇒ 3H2O + CH4).20 In the nitrogen-fixing process (H2 ⇒ 2H+ + 2 e−; N2 + 8H+ + 8e− ⇒ 2NH3 + H2),37 0.98 mM of the nitrogen observed in the effluent indicated the consumption of 2.94 mmol of H2. The degradation of phenol and 3-CP (50 μmoles in total) would release 700 μmoles of H2 (phenol + 5H2O ⇒ 3CH3COOH + 2H2, 3CH3COOH + 6H2O ⇒ 12H2 + 6CO2).20,41 Assuming a chemical formula of C5H9O3N as the anaerobic microbial biomass,43 one unit of microbial biomass carbon generation would require 1.8 units of hydrogen, corresponding to the utilization of 13.0 mmol of hydrogen for 7.2 mmol of biomass carbon in the effluent (SI Table S1). Subtracting the hydrogen for microbial biomass and adding the production from phenol, the system could produce 114 mmol of H2 at most. While, 5.19 mmol of H2 (about 4.6% of the total produced) was estimated to be utilized for the biochemical processes: dechlorination, methane production, and nitrogen fixation. This suggested the presence of excess H2 (electrons) in the system. The remaining H2 (about 95.4%) probably flowed out from the column. The average concentration of H2 observed in the effluent was 0.75 mM (Figure 1D), which is close to the saturated H 2 concentration in water at 30 °C.44 Besides carbon, hydrogen, nitrogen, and oxygen, other trace elements were required for microbial growth and metabolism, such as calcium, potassium, and phosphorus. Sources for these elements would either be the impurities in PCP−Na and lactate− Na or the soil added to the column. The major impurities of PCP−Na are polychlorinated or polybrominated dibenzo-pdioxins and dibenzofurans,45 which are more recalcitrant and toxic than PCP. These impurities would not be utilizable as carbon sources or electron donors in this system. Therefore, they would not affect the PCP mineralization process. The high performance of PCP mineralization and the compact distribution of different microorganisms in the column suggested that a compact single-layer-type biobarrier with a lower engineering cost can be designed by the proper spatial alignment of different microorganisms. Further study is warranted on the spatial distribution of functional microbial populations involved in anaerobic PCP dechlorination, phenol/3-CP degradation, and nitrogen fixation by this system.

PCR−DGGE analysis of partial 16 rRNA genes of the phyla Bacteria and Archaea in the bottom and the top of the column with the original introduced PCP dechlorination culture (Figure S1), the phylogenetic assignments of the DGGE bands (Table S3), the results of the PCR−DGGE analysis in the column effluent (Figure S2), and the phylogenetic assignments of the DGGE bands (Table S4). This material is available free of charge via the Internet at http://pubs.acs.org.


Corresponding Author

*Tel: +81−52−789−5856; fax: +81−52−789−5857; e-mail: [email protected] Notes

The authors declare no competing financial interest.

ACKNOWLEDGMENTS We thank the members of the laboratory for their enthusiastic help. We also thank Dr. Naoko Yoshida of EIIRIS, Toyohashi University of Technology for her guidance of cloning and sequencing of 16S rRNA genes. This study was supported by a University Grant for “Design of cascade utilization system for unused biological resources in the Tokai area” and Grants-in-Aid for Scientific Research (B2:20310041 and B2:23310055) from the Ministry of Education, Culture, Sports, Science and Technology of Japan and the New Energy and Industrial Technology Development Organization.


(1) Shiu, W. Y.; Ma, K. C.; Varhanickova, D.; Mackay, D. Chlorophenols and alkylphenols - A review and correlation of environmentally relevant properties and fate in an evaluative environment. Chemosphere 1994, 29 (6), 1155−1224. (2) Keith, L. H.; Telliard, W. A. Priority pollutants I-a perspective view. Environ. Sci. Technol. 1979, 13 (4), 416−423. (3) Amos, B. K.; Suchomel, E. J.; Pennell, K. D.; Löffler, F. E. Microbial activity and distribution during enhanced contaminant dissolution from a NAPL source zone. Water Res. 2008, 42 (12), 2963−2974. (4) Amos, B. K.; Suchomel, E. J.; Pennell, K. D.; Löffler, F. E. Spatial and temporal distributions of Geobacter lovleyi and Dehalococcoides spp. during bioenhanced PCE-NAPL dissolution. Environ. Sci. Technol. 2009, 43 (6), 1977−1985. (5) Lohner, S. T.; Tiehm, A. Application of electrolysis to stimulate microbial reductive PCE dechlorination and oxidative VC biodegradation. Environ. Sci. Technol. 2009, 43 (18), 7098−7104. (6) Foght, J. Anaerobic biodegradation of aromatic hydrocarbons: Pathways and prospects. J. Mol. Microbiol. Biotechnol. 2008, 15 (2−3), 93−120. (7) Armenante, P. M.; Kafkewitz, D.; Lewandowski, G. A.; Jou, C. J. Anaerobic-aerobic treatment of halogenated phenolic compounds. Water Res. 1999, 33 (3), 681−692. (8) Majumder, P. S.; Gupta, S. K. Removal of chlorophenols in sequential anaerobic-aerobic reactors. Bioresour. Technol. 2007, 98 (1), 118−129. (9) Huang, L. P.; Chai, X. L.; Quan, X.; Logan, B. E.; Chen, G. H. Reductive dechlorination and mineralization of pentachlorophenol in biocathode microbial fuel cells. Bioresour. Technol. 2012, 111, 167−174. (10) Li, Z. L.; Yang, S. Y.; Inoue, Y.; Yoshida, N.; Katayama, A. Complete anaerobic mineralization of pentachlorophenol (PCP) under continuous flow conditions by sequential combination of PCPdechlorinating and phenol-degrading consortia. Biotechnol. Bioeng. 2010, 107 (5), 775−785. (11) Yang, S.; Shibata, A.; Yoshida, N.; Katayama, A. Anaerobic mineralization of pentachlorophenol (PCP) by combining PCPdechlorinating and phenol-degrading cultures. Biotechnol. Bioeng. 2009, 102 (1), 81−90.


* Supporting Information S

Detailed description of the experimental procedures and the results for the carbon balance at the steady state of phase I and phase II (Table S1), the 16 rRNA gene library results on the bottom and the top of the column at 19.3 PV (Table S2), the 1540

dx.doi.org/10.1021/es303784f | Environ. Sci. Technol. 2013, 47, 1534−1541

Environmental Science & Technology


(12) Kamashwaran, S. R.; Crawford, D. L. Anaerobic biodegradation of pentachlorophenol in mixtures containing cadmium by two physiologically distinct microbial enrichment cultures. J. Ind. Microbiol. Biotechnol. 2001, 27 (1), 11−17. (13) Mikesell, M. D.; Boyd, S. A. Complete reductive dechlorination and mineralization of pentachlorophenol by anaerobic microorganisms. Appl. Environ. Microbiol. 1986, 52 (4), 861−865. (14) Kennes, C.; Wu, W. M.; Bhatnagar, L.; Zeikus, J. G. Anaerobic dechlorination and mineralization of pentachlorophenol and 2,4,6trichlorophenol by methanogenic pentachlorophenol-degrading granules. Appl. Microbiol. Biotechnol. 1996, 44 (6), 801−806. (15) Wu, W. M.; Bhatnagar, L.; Zeikus, J. G. Performance of anaerobic granules for degradation of pentachlorophenol. Appl. Environ. Microbiol. 1993, 59 (2), 389−397. (16) Yoshida, N.; Yoshida, Y.; Handa, Y.; Kim, H. K.; Ichihara, S.; Katayama, A. Polyphasic characterization of a PCP-to-phenol dechlorinating microbial community enriched from paddy soil. Sci. Total Environ. 2007, 381 (1−3), 233−242. (17) Karlsson, A.; Ejlertsson, J.; Svensson, B. H. CO2-dependent fermentation of phenol to acetate, butyrate and benzoate by an anaerobic, pasteurised culture. Arch. Microbiol. 2000, 173 (5−6), 398− 402. (18) Knoll, G.; Winter, J. Degradation of phenol via carboxylation to benzoate by a defined, obligate syntrophic consortium of anaerobicbacteria. Appl. Microbiol. Biotechnol. 1989, 30 (3), 318−324. (19) Becker, J. G.; Berardesco, G.; Rittmann, B. E.; Stahl, D. A. The role of syntrophic associations in sustaining anaerobic mineralization of chlorinated organic compounds. Environ. Health Perspect. 2005, 113 (3), 310−316. (20) Qiu, Y. L.; Hanada, S.; Ohashi, A.; Harada, H.; Kamagata, Y.; Sekiguchi, Y. Syntrophorhabdus aromaticivorans gen. nov., sp nov., the first cultured anaerobe capable of degrading phenol to acetate in obligate syntrophic associations with a hydrogenotrophic methanogen. Appl. Environ. Microbiol. 2008, 74 (7), 2051−2058. (21) Harwood, J. E.; Vansteen, R. A; Kuhn, A. L. A comparison of some methods for total phosphate analyses. Water Res. 1969, 3 (6), 425−432. (22) Li, Z. L.; Inoue, Y.; Yang, S. Y.; Yoshida, N.; Katayama, A. Mass balance and kinetic analysis of anaerobic microbial dechlorination of pentachlorophenol in a continuous flow column. J. Biosci. Bioeng. 2010, 110 (3), 326−332. (23) Guebel, D. V.; Nudel, B. C.; Giulietti, A. M. A simple and rapid micro-Kjeldahl method for total nitrogen analysis. Biotechnol. Technol. 1991, 5 (6), 427−430. (24) Lohrenz, S. E.; Taylor, C. D. Primary production of protein 0.1. Comparison of net cellular carbon and protein-synthesis with c-14 derived rate estimates in steady-state cultures of marine-phytoplankton. Mar. Ecol.: Prog. Ser. 1987, 35 (3), 277−292. (25) Widdel, F.; Kohring, G. W.; Mayer, F. Studies on dissimilatory sulfate-reducing bacteria that decompose fatty acids. III. Characterization of the filamentous gliding Desulfonema limicola gen-nov sp-nov, and Desulfonema magnum sp-nov. Arch. Microbiol. 1983, 134 (4), 286− 294. (26) Wilhelm, E.; Battino, R.; Wilcock, R. J. Low-pressure solubility of gases in liquid water. Chem. Rev. 1977, 77 (2), 219−262. (27) Holliger, C.; Hahn, D.; Harmsen, H.; Ludwig, W.; Schumacher, W.; Tindall, B.; Vazquez, F.; Weiss, N.; Zehnder, A. J. B. Dehalobacter restrictus gen. nov. and sp. nov., a strictly anaerobic bacterium that reductively dechlorinates tetra- and trichloroethene in an anaerobic respiration. Arch. Microbiol. 1998, 169 (4), 313−321. (28) Christiansen, N.; Ahring, B. K. Desulf itobacterium hafniense sp nov, an anaerobic, reductively dechlorinating bacterium. Int. J. Syst. Bacteriol. 1996, 46 (2), 442−448. (29) Juteau, P.; Cote, V.; Duckett, M. F.; Beaudet, R.; Lepine, F.; Villemur, R.; Bisaillon, J. G. Cryptanaerobacter phenolicus gen. nov., sp. nov., an anaerobe that transforms phenol into benzoate via 4hydroxybenzoate. Int. J. Syst. Evol. Microbiol. 2005, 55, 245−250. (30) Schocke, L.; Schink, B. Energetics and biochemistry of fermentative benzoate degradation by Syntrophus gentianae. Arch. Microbiol. 1999, 171 (5), 331−337.

(31) Winter, J.; Lerp, C.; Zabel, H. P.; Wildenauer, F. X.; Konig, H.; Schindler, F. Methanobacterium-wolfei, sp-nov, a new tungsten-requiring, thermophilic, autotrophic methanogen. Syst. Appl. Microbiol. 1984, 5 (4), 457−466. (32) Yashiro, Y.; Sakai, S.; Ehara, M.; Miyazaki, M.; Yamaguchi, T.; Imachi, H. Methanoregula formicica sp. nov., a methane-producing archaeon isolated from methanogenic sludge. Int. J. Syst. Evol. Microbiol. 2011, 61, 53−59. (33) He, J. Z.; Robrock, K. R.; Alvarez-Cohen, L. Microbial reductive debromination of polybrominated diphenyl ethers (pbdes). Environ. Sci. Technol. 2006, 40 (14), 4429−4434. (34) Rivas, R.; Willems, A.; Palomo, J. L.; Garcia-Benavides, P.; Mateos, P. F.; Martinez-Molina, E.; Gillis, M.; Velazquez, E. Bradyrhizobium betae sp nov., isolated from roots of beta vulgaris affected by tumour-like deformations. Int. J. Syst. Evol. Microbiol. 2004, 54, 1271−1275. (35) Senthilkumar, M.; Madhaiyan, M.; Sundaram, S. P.; Sangeetha, H.; Kannaiyan, S. Induction of endophytic colonization in rice (Oryza sativa L.) tissue culture plants by Azorhizobium caulinodans. Biotechnol. Lett. 2008, 30 (8), 1477−1487. (36) Hunter, W. J.; Kuykendall, L. D.; Manter, D. K. Rhizobium selenireducens sp nov.: A selenite-reducing alpha-proteobacteria isolated from a bioreactor. Curr. Microbiol. 2007, 55 (5), 455−460. (37) Cheng, Q. Perspectives in biological nitrogen fixation research. J. Integr. Plant. Biol. 2008, 50 (7), 786−798. (38) Ju, X. F.; Zhao, L. P.; Sun, B. L. Nitrogen fixation by reductively dechlorinating bacteria. Environ. Microbiol. 2007, 9 (4), 1078−1083. (39) Bryant, M. P.; Campbell, L. L.; Reddy, C. A.; Crabill, M. R. Growth of Desulfovibrio in lactate or ethanol media low in sulfate in association with H2-utilizing methanogenic bacteria. Appl. Environ. Microbiol. 1977, 33 (5), 1162−1169. (40) Sun, B. L.; Cole, J. R.; Sanford, R. A.; Tiedje, J. M. Isolation and characterization of Desulfovibrio dechloracetivorans sp nov., a marine dechlorinating bacterium growing by coupling the oxidation of acetate to the reductive dechlorination of 2-chlorophenol. Appl. Environ. Microbiol. 2000, 66 (6), 2408−2413. (41) Ohnishi, A.; Hasegawa, Y.; Abe, S.; Bando, Y.; Fujimoto, N.; Suzuki, M. Hydrogen fermentation using lactate as the sole carbon source: Solution for ‘blind spots’ in biofuel production. RSC Adv. 2012, 2 (22), 8332. (42) Mazur, C. S.; Jones, W. J.; Tebes-Stevens, C. H2 consumption during the microbial reductive dehalogenation of chlorinated phenols and tetrachloroethene. Biodegradation 2003, 14 (4), 285−295. (43) Nopharatana, A.; Pullammanappalli, P. C.; Clarke, W. P. Kinetics and dynamic modelling of batch anaerobic digestion of municipal solid waste in a stirred reactor. Waste Manage. (Oxford) 2007, 27 (5), 595− 603. (44) Jauregui-Haza, U. J.; Pardillo-Fontdevila, E. J.; Wilhelm, A. M.; Delmas, H. Solubility of hydrogen and carbon monoxide in water and some organic solvents. Lat. Am. Appl. Res. 2004, 34 (2), 71−74. (45) Xu, M. X.; Yan, J. H.; Lu, S. Y.; Li, X. D.; Chen, T.; Ni, M. J.; Dai, H. F.; Cen, K. F. Source identification of PCDD/FS in agricultural soils near to a Chinese MSWI plant through isomer-specific data analysis. Chemosphere 2008, 71 (6), 1144−1155.


dx.doi.org/10.1021/es303784f | Environ. Sci. Technol. 2013, 47, 1534−1541