Genetic bioaugmentation of activated sludge with dioxin-catabolic

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Remediation and Control Technologies

Genetic bioaugmentation of activated sludge with dioxincatabolic plasmids harbored by Rhodococcus sp. strain p52 Chongyang Ren, Yiying Wang, Lili Tian, Meng Chen, Jiao Sun, and Li Li Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b04633 • Publication Date (Web): 02 Apr 2018 Downloaded from http://pubs.acs.org on April 3, 2018

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Environmental Science & Technology

Genetic bioaugmentation of activated sludge with dioxin-catabolic plasmids harbored by Rhodococcus sp. strain p52

Chongyang Ren ⋅ Yiying Wang ⋅ Lili Tian ⋅ Meng Chen ⋅ Jiao Sun ⋅ Li Li*

Shandong Provincial Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Shandong University, Jinan 250100, China

* Corresponding author Phone: +86-531-88364250; fax: +86-531-88364513; e-mail address: [email protected]

ABSTRACT Horizontal transfer of catabolic plasmids is used in genetic bioaugmentation for environmental pollutant remediation. In this study, we examined the effectiveness of genetic bioaugmentation with dioxin-catabolic plasmids harbored by Rhodococcus sp. strain p52 in laboratory-scale sequencing batch reactors (SBRs). During 100 days of operation, bioaugmentation decreased the dibenzofuran content (120 mg L−1) in the synthetic wastewater by 32.6%–100% of that in the non-bioaugmented SBR. Additionally, dibenzofuran was removed to undetectable levels in the bioaugmented SBR, in contrast, 46.8±4.1% of that in the influent remained in the non-bioaugmented SBR after 96 d. Moreover, transconjugants harboring pDF01 and pDF02 were

ACS Paragon Plus Environment


isolated from the bioaugmented SBR after 2 d, and their abilities to degrade dibenzofuran were confirmed. After 80 d, the copy numbers of strain p52 decreased by three orders of magnitude and accounted for 0.05±0.01% of the total bacteria, while transconjugants were present at around 106 copies mL˗1 sludge and accounted for 8.2±0.3% of the total bacteria. Evaluation of the bacterial community profile of sludge by high-throughput 16S rRNA gene sequencing revealed that genetic bioaugmentation led to a bacterial community with an even distribution of genera in the SBR. This study demonstrates the promise of genetic bioaugmentation with catabolic plasmids for dioxins remediation.

TOC/ABSTRACT

1. INTRODUCTION Polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs), generally termed as dioxins, have attracted considerable attention over recent decades because of their persistence, bioaccumulation, toxicity and global dispersion.1 During the last three decades, global release of PCDDs/PCDFs has decreased greatly;2 however, unregulated waste production or improper disposal of


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historical contaminates still poses a risk to human health.3–5 Accordingly, efficient methods of decontamination are needed to reduce the contaminant level of PCDDs/PCDFs deposits.3,5 Removal of these contaminants by microorganisms is an important method of remediation. Some aerobic bacteria can effectively mineralize the carbon backbones of PCDDs/PCDFs and degrade low-chlorinated PCDDs/PCDFs.6 Aerobic bacteria utilize a series of enzymes to catalyze the breakdown of aromatic rings of dioxins. The key step in this process is initial dihydroxylation, which is mainly catalyzed by the angular dioxygenase that attacks the ring adjacent to the ether oxygen.7–9 Interestingly, some Gram-positive (G+) bacteria possess two conserved gene clusters, dbfA and dfdA, encoding two angular dioxygenases involved in dioxin dihydroxylation.9,10 The coexistence of dbfA and dfdA clusters in these G+ bacteria is considered to be beneficial to their capacity for degradation

of

a

broad

range

of

compounds,

including

dibenzofuran,

dibenzo-p-dioxin, chlorodibenzofurans, carbazole, dibenzothiophene, anthracene, phenanthrene and biphenyl.9,10 Because dioxins are recalcitrant, when the potential for in situ degradation is absent or low, bioaugmentation is a preferred strategy. Aerobic bacteria capable of degrading dioxins are attractive candidates for bioaugmentation.11 Moreover, genes involved in dioxin catabolism are often located on plasmids, especially broad-host conjugative plasmids, suggesting their potential for application in genetic bioaugmentation for dioxins bioremediation.12,13 Genetic bioaugmentation is a bioremediation method in which contaminant degradation potential is enhanced through plasmid conjugative transfer to indigenous microorganisms by introducing



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donor bacteria harboring conjugative (or mobilizable) catabolic plasmids that encode for the degradation of target contaminants.14,15 Genetic bioaugmentation aims to transfer relevant genes into indigenous microorganisms with greater fitness for survival in the contaminated environment; therefore, it has an advantage over conventional bioaugmentation, which relies on the survival of exogenous microorganisms.14,16 During the past two decades, the feasibility and factors affecting genetic bioaugmentation have been examined.15–25 In these studies, the transfer of catabolic plasmids resulted in enhanced contaminant degradation, indicating the potential for application of genetic bioaugmentation.17,23,24,26 Factors that influence plasmid transfer such as donors, nutrients and the genetic characteristics of recipients, have been investigated.15,16,20,22,27 Attention has been paid to the contributors to degradation enhancement (i.e., donor or transconjugants)18 and indigenous bacterial communities influenced by donors.28,29 In a wastewater treatment facility, high density and rapid growth of bacteria in a complex microbial community are in favor of plasmid conjugative transfer. Therefore such facilities constitute good model systems for investigations into genetic bioaugmentation.30 Poor bioreactor performance during wastewater treatment is ascribed to the lack of a sufficient number of specific microorganisms harboring a key metabolic route to transform the target contaminants,31 which may be ameliorated by successful genetic bioaugmentation. Although

there

are

some

bacteria

harboring

conjugative

dioxin-catabolic

plasmids,12,13 only a few studies have addressed strategies of genetic bioaugmentation to date,13,33 and there is no report on genetic bioaugmentation of bioreactors for dioxin


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removal to our knowledge. We previously isolated the dioxin-degrader, Rhodococcus sp. strain p52, which has two distinct gene clusters (dfdA and dbfA) involved in initial dihydroxylation of dioxins located on two mega plasmids, pDF01 and pDF02, respectively.32 In a previous study, conjugative transfer of the dioxin-catabolic plasmids pDF01 and pDF02 harbored by strain p52 to tested recipients was confirmed. Moreover, their transfer to activated sludge bacteria was detected during mixing strain p52 with the sludge bacteria, revealing the potential for application of pDF01 and pDF02 for genetic bioaugmentation.13 In this study, we conducted bioaugmentation of activated sludge in a sequencing batch reactor (SBR) with strain p52 harboring pDF01 and pDF02 for dioxin removal. Dibenzofuran, the non-halogenated analogue of PCDFs, was used as a model compound to study dioxin biodegradation. The purpose of this study was to learn: (1) the effects of genetic bioaugmentation on dibenzofuran removal, (2) the fate of donor strain p52 along with the occurrence of transconjugants in the reactor, and (3) the effects of bioaugmentation on the sludge indigenous microbial community. To accomplish this, we compared the removal of dibenzofuran in a bioaugmented reactor to that in a non-bioaugmented reactor. In addition, transconjugants were isolated from sludge in the bioaugmented reactor and their abilities to degrade dibenzofuran were confirmed. We also monitored the quantity of bacteria in the SBRs including strain p52, the transconjugants and the total bacteria. Furthermore, the bacterial communities in the bioaugmented and non-bioaugmented reactors were characterized and compared.



2. MATERIALS AND METHODS 2.1. Laboratory reactor setup. Two parallel laboratory-scale reactors (60 mm internal diameter and 450 mm height) with a working volume of 1 L were operated in sequencing batch mode. An air diffuser was installed at the bottom of each reactor to aerate at a rate of 0.8 L min-1. The operational temperature was maintained around 17 °C–25 °C and the pH was controlled at 6.0–8.0 in both reactors. The hydraulic retention time and solids retention time was set at 24 h and 20 days, or adjusted to be 12 h and 15 days, respectively. Each cycle consisted of 5 min filling, 15 or 30 min settling (which was increased if the sludge settling properties deteriorated), 5 min drawing, and aeration in the remaining time. The volume exchange ratio was set at 50%. Both reactors were seeded with sludge obtained from the outlet of an aeration tank in a municipal wastewater treatment plant of the Guangda Waterworks (Jinan, China). Prior to inoculation, the seed sludge was aerated for one week to exhaust the storage of the carbon source. The initial content of mixed liquor volatile suspended solids (MLVSS) in each reactor was 1.713 g L−1. Synthetic wastewater was fed into two SBRs which was prepared according to Van Loosdrecht et al.34 with a modification. The synthetic wastewater contained (per L): glucose 0.454 g, sodium acetate 0.621 g, NH4Cl 0.573 g, KH2PO4 0.088 g, EDTA 0.040 g, MgSO4·7H2O 0.050 g, FeCl3·6H2O 0.003 g, FeSO4·7H2O 0.040 g, NaHCO3 0.125 g and 0.5 mL of trace element solution (its composition is detailed in the Supporting Information). The pH of the synthetic wastewater was about 7.0. The SBRs were pre-operated for 10 days


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when both reached a similar performance for chemical oxygen demand (COD) removal. Then dibenzofuran, dissolved in ethanol as a stock solution (120 g L−1), was supplemented to the synthetic wastewater, feeding the SBRs in each cycle. The resulted concentration of COD in the feeding was approximately 3000 mg L−1, and the ratio of COD, N (NH4+-N), and P (PO43–-P) was 100: 5: 1. Two batches of experiments were conducted independently, which differed in the concentration of supplemented dibenzofuran and the bioaugmentation frequency (Table S1). 2.2. Bioaugmentation. In the two SBRs, one reactor was bioaugmented with Rhodococcus sp. strain p52 harboring catabolic plasmids pDF01 and pDF02, while the other reactor was not bioaugmented as a control. Prior to inoculation, strain p52 was cultured in carbon-free mineral medium32 supplemented with 500 mg L−1 dibenzofuran at 30 °C, with shaking at 180 rpm for 48 h. Cells of strain p52 were then harvested, washed twice and suspended in mineral medium for inoculation. Strain p52 was inoculated into the SBR at a final cell density of 106–107 colony-forming units mL−1. Bioaugmentation was conducted differently in the two independent experiments (see Supporting Information). To determine the fate of the induced strain in the SBR, this study focused on the first experiment, in which strain p52 was inoculated only once, along with the supplemental feeding of dibenzofuran. In the second experiment, strain p52 was inoculated into the bioaugmented reactor three times instead of a single inoculation (Table S1). Duplicate mixed-liquor samples taken from each SBR at the end of each aeration period were analyzed for residual dibenzofuran. Dibenzofuran extraction and analysis



by gas chromatography (GC) were conducted according to Peng et al.32 Briefly, the mixed-liquor samples were extracted with an equal volume of ethyl acetate. The extract was measured by GC using an Agilent 6890 equipped with an HP-5 column (length 30 m, inner diameter 0.32 mm, and film thickness 0.25 µm) with a flame ionization detector. The following oven temperature program was implemented: increased from 60 °C to 180 °C at a rate of 10 °C min−1, followed by an increase at 15 °C min−1 to 280 °C. The injector and detector temperatures were 300 °C. Data are expressed as the means of analyses for duplicate samples. 2.3. DNA extraction and quantitative PCR analysis (qPCR). Ten milliliter mixed-liquor samples were collected (at t = 0, 10, 30, 51, 75 and 80 days of treatment) from each SBR and subjected to DNA extraction. The total DNA of sludge samples was extracted using a soil DNA isolation kit (Omega Bio-tek, Inc., Norcross, GA, USA) according to the manufacturer’s instructions. DNA quality and quantity were determined on 1% agarose gels and using a NanoDrop 2000 spectrophotometer. A 10-fold or 100-fold dilution of DNA extract of the sludge was used as template to remove PCR inhibition. Real-time qPCR was conducted to quantify dioxygenase genes (dfdA1 and dbfA1), the specific DNA sequence of strain p52 (the intergenic spacer region between the 16S and 23S rRNA genes, ISR), and the 16S rRNA genes of the total bacteria in the sludge samples. On the basis of the genome sequence of strain p52 (GenBank accession numbers CP016819, CP016820, and CP016821), primers were designed using the Primer-BLAST tool on the National Center for Biotechnology Information (NCBI). Primers and amplicons for qPCR analysis are


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described in Table 1. The reaction mixture contained 2–10 ng of template DNA, 10 µL SYBR Green Mastermix (BioRad, Hercules, CA, USA), and 500 nM of each primer in a final reaction volume of 20 µL. qPCR was performed on a BioRad iCycler (BioRad) with an initial denaturation step of 2 min at 98 °C, followed by 40 cycles of denaturation at 98 °C for 5 s and annealing/extension for 10 s at 56 °C. Melting curves were constructed after each qPCR run to check the specificity of the PCR for the target. Genomic DNA of strain p52, which has been completely sequenced and has a total size of 5,431,587 bp, was used as a standard. A five-point serial decimal dilution of the standard was run in triplicate with each set of reactions to generate a standard curve of threshold cycle (Ct) values versus the number of gene copies. The mass of the genomic DNA was spectrophotometrically determined, and the gene copy number was calculated as described by Yun et al.36 Standard curves were linear (R2 > 0.97) for all targets, with an efficiency of 98%–102%. Absolute copy numbers of target DNA fragments were calculated using the Ct values based on the standard curves. All qPCR assays were run in triplicate for each sludge sample along with non-template negative controls. Data were expressed as average gene copy numbers per milliliter of sludge ± standard deviations based on triplicate analyses.

Table 1. Primers and amplicon information for PCR analysis Primer usage

Target

Oligonucleotide sequence

Amplicon

Source or

(5’ to 3’)

size (bp)

reference



qPCR

dbfA1

F: TACAGCAACAACGGCGATCT

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206

This study

153

This study

210

This study

155

This study

193

This study

352

Reference

R: TCGAAGGCCAGGTCCAGATA Conventional

dbfA1

F: AGGCGATCGAGTACGTCAAC

PCR qPCR

R: ATTGGAAGAGCCCGAAGACG dfdA1

F: CCGGTATAACCTGTCTCGGC R: CGGTAGTCTCCTACACCGGA

Conventional

dfdA1

F: GTGCGTACGACCGAGGTAAT

PCR qPCR

R: GAGCCTGTCGTCAAGGAGTC Strain

F: GCCAGAGACCGATTGTCCC

p52 ISR*

R:AAGAAAACTGATTTCCTTGTTTC GC

qPCR

Bacterial 16S

F: ATGGCTGTCGTCAGCT

rRNA R: ACGGGCGGTGTGTAC

35

gene

*ISR refers to the intergenic spacer region between the 16S and 23S rRNA genes

2.4. Transconjugants isolation, degradation test, and plasmid stability test. Mixed-liquor samples collected (at t = 0, 1, 2, 3, 6, 10, 20 and 30 days) from the bioaugmented reactor were diluted and spread onto selective plates prepared from dibenzofuran-supplemented carbon-free mineral medium as described by Sun et al.13 Because no seed sludge grew on the selective plates within 72 h, the colonies that did growing on these plates (at 30 °C) with different morphologies from that of strain p52


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(orange-red, round, and less than 1 mm in diameter) were selected as potential transconjugants. To confirm the transconjugants, polymerase chain reaction (PCR) was conducted to amplify a 155 bp dfdA1 fragment of pDF01 and a 153 bp dbfA1 fragment of pDF02 using the primer sets shown in Table 1. The amplicons were confirmed by sequencing analysis. Simultaneously, the identity of transconjugant was determined by 16S rRNA gene amplification from the same colony and sequencing analysis. Based on the 16S rRNA gene sequences of the isolated transconjugants, phylogenetic analysis was conducted (see Supporting Information). The confirmed transconjugants were also subjected to a degradation test. Briefly, transconjugants were pre-cultured in LB medium overnight, then harvested by centrifuging at 6000×g for 10 min, washed twice and suspended in mineral medium for inoculation. The degradation experiment was conducted in 250 mL Erlenmeyer flasks containing 50 mL of carbon-free mineral medium supplemented with 500 mg L−1 dibenzofuran as the sole carbon source. The transconjugants were inoculated into the flasks to an initial OD600 of approximately 0.2. All flasks were incubated aerobically at 30 °C with vigorous shaking at 180 rpm. After 96 h, the residual dibenzofuran was extracted and measured by gas chromatography as described above. Data are reported as the means with average deviations of independent triplicate experiments. Plasmid stability in the isolated transconjugants was tested by a continuous passage method. Briefly, each transconjugant tested was inoculated onto LB plates and cultivated for 24 h, then colonies of each transconjugant were randomly picked and streaked onto fresh LB plates, as well as selective plates. After 24 h, the colonies



on LB plates were streaked onto fresh LB plates and the selective plates again. The colonies grown on selective plates for 72 h after each transfer were subjected to PCR amplification of dfdA1 and dbfA1 and the presence of the plasmids was confirmed by sequencing analysis as described above. This process was repeated until the colonies no longer grew on the selective plates. 2.5. Analysis of the bacterial community composition by Illumina MiSeq sequencing. Total DNA was extracted from seed sludge as well as mixed-liquor samples collected from two SBRs after 75 d of operation using a soil DNA isolation kit (Omega Bio-tek, Inc., Norcross, GA, USA) according to the manufacturer’s instructions. The bacterial composition was analyzed by sequencing the amplified V3-V4 region of the 16S rRNA gene using the bacterial universal primers 341F (5ʹ– CCTACGGGNGGCWGCAG–3ʹ) and 805R (5ʹ–GACTACHVGGGTATCTAATCC– 3ʹ).37 The amplification was performed under the following conditions: initial denaturation at 94 °C for 3 min followed by 5 cycles of 30 s at 94 °C, 20 s at 45 °C and 30 s at 65 °C, 20 cycles of 20 s at 94 °C, 20 s at 55 °C and 30 s at 72 °C, and then a final extension at 72 °C for 5 min. PCR products were purified with Agencourt AMPure XP beads (Beckman Coulter, San Diego, CA, USA) and quantified using the NanoVue Plus Spectrophotometer (GE Healthcare, Piscataway, NJ, USA). Purified amplicons were pooled in equimolar amounts, and samples were multiplexed using a dual-indexing approach. The final library was paired-end sequenced at 2×300-bp on the Illumina MiSeq platform. High-quality sequences were assigned to samples according to barcodes. Chimeric


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sequences were identified and removed using UCHIME. OTUs were clustered with a 97% similarity cutoff using Usearch (http://www.drive5.com/usearch/). OTUs that reached 97% similarity were employed for alpha diversity estimations, which included diversity (Shannon, Simpson), richness (Chao I and ACE) and Good’s coverage.38 Taxonomic classifications were assigned to OTUs by RDP Classifier (http://rdp.cme.msu.edu/) with a confidence threshold of 80%, as well as the BLASTN program of NCBI with an output of > 90% sequence identity over 90% coverage. The raw sequencing data obtained from this study was deposited in the NCBI Sequence Read Archive under accession number SRP115514. 2.6. Statistical analysis. Statistical analysis was performed using SPSS version 22.0. The unpaired, two-tailed Student's t-test was used to identify statistical differences between samples from the control and bioaugmented SBRs. Statistical significance was accepted at the p0.05) between the MLVSS of the bioaugmented reactor (1.702 g L−1±0.012 (n=3)) and the non-bioaugmented control



(1.685 g L−1±0.032 (n=3)), however a slight decline occurred in the latter when dibenzofuran strongly accumulated. The effects of bioaugmentation with strain p52 on removal of dibenzofuran in the SBR were examined. As shown in Figure 1, significant enhancement of dibenzofuran removal occurred in the bioaugmented reactor compared with the non-bioaugmented control throughout the experimental period. Because of the insolubility of dibenzofuran, insufficient degradation resulted in its accumulation in the non-bioaugmented reactor (days 1–13 and 51–60), but this occurred to a lesser degree in the bioaugmented reactor. During days 1–50, the dibenzofuran concentration in the bioaugmented SBR was lower than that in the non-bioaugmented SBR. More specifically, the former accounted for 3.8%–34% of the latter. During days 51–100, as the cycle time was reduced from 24 h to 12 h, the dibenzofuran in the bioaugmented SBR was less than 67.4% of that in the non-bioaugmented SBR. After 96 d, dibenzofuran was degraded to below the detection limit in the bioaugmented reactor, in comparison to the non-bioaugmented reactor where 46.8±4.1% of that in the influent remained. When bioaugmentation was conducted by repeated inoculation of strain p52 at low dosage, significant enhancement of dibenzofuran removal was also observed in the bioaugmented reactor (Figure S1). After 53 d, the dibenzofuran content was around 21.87±8.75 mg L−1 in the bioaugmented reactor, while more than 68.54±17.79 mg L−1 remained in the non-bioaugmented reactor (p 0.5%) level is shown in Figure 3A, and the heat map of the bacterial genera (relative abundance > 1.5%) is shown in Figure 3B.



Figure 3. Relative abundance of bacterial 16S rRNA genes at the phylum level (A), and heat map of the bacterial genera (B) in the sludge samples from the bioaugmented reactor (R1), the non-bioaugmented control reactor (R2) and seed sludge (SS). Phyla with a relative abundance > 0.5% and genera with a relative abundance > 1.5% in at least one sample are included. Taxa not included are summarized in the artificial group “Others”.


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As shown in Figure 3A, bacterial communities were all dominantly colonized by Proteobacteria and Bacteroidetes, but at different densities in the sludge samples from those in the bioaugmented reactor (69.5% and 21.5% of reads, respectively), the non-bioaugmented reactor (87.4% and 9.0%, respectively) and seed sludge (74.2% and 14.0%, respectively). Among the Proteobacteria, the most abundant OTUs fell into the Methylobacteriaceae family (64.0% of reads) in the non-bioaugmented reactor sludge, but were scattered among the Methylobacteriaceae (18.0%), Sphingomonadaceae (13.4%) and Rhodobacteraceae (11.4%) in the bioaugmented reactor sludge. Microorganisms in these families are characterized by their flexible or diverse nutritional modes and adaption to aquatic environments. By contrast, Proteobacteria

were

mainly

represented

by

Rhodocyclaceae

(30.7%)

and

Alteromonadaceae (19.9%) in seed sludge. As shown in Figure 3B, the density of genera differed among sludge bacterial communities. In particular, the bacterial community in the bioaugmented reactor displayed a more even distribution of genera compared with the non-bioaugmented control. A Venn diagram of the number of genera with a relative abundance > 0.1% revealed that genera shared by the three sludge samples accounted for 7.9% of all genera observed (Figure 4), including Phaeodactylibacter, Brevundimonas and Paracoccus (Figure 3B), while about 11.9% of genera were common to the sludge sampled from the SBRs (Figure 4), including Meganema and Sphingobium (Figure 3B). There were a higher proportion of genera detected specifically in each sludge sample, e.g., Pseudomonas (6.1% of reads), Enterobacter (5.0%) and Hartmannibacter (2.1%) were detected only in the



bioaugmented reactor. Interestingly, genera affiliated with most of the isolated transconjugants were only detected in the bioaugmented reactor, but Arthrobacter, Corynebacterium and Klebsiella accounted for less than 0.5% of all reads. These findings indicated that both selective pressure (dibenzofuran) and genetic bioaugmentation play a role in development of the bacterial community composition.

Figure 4. Venn diagram of genera in different sludge samples in the bioaugmented reactor (R1), the non-bioaugmented control reactor (R2) and seed sludge (SS). Only genera with a relative abundance > 0.1% were included. Bacterial genera detected in sludge samples (R1, R2 and SS) are displayed in blue, yellow and green colored circles, respectively. Common bacterial genera are displayed in the overlapping regions of the circles. The number of genera included and their proportion of the total genera are also presented.


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4. DISCUSSION This study has demonstrated the possible use of genetic bioaugmentation for remediation of dioxins contamination. The study was conducted in laboratory-scale SBRs which served as a model engineering system. The plasmid-mediated bioaugmentation was facilitated in the SBR by the complex bacterial community, close contact between donor and recipient, relatively stable selection pressure, and controllable environmental conditions. A success of genetic bioaugmentation is judged by the enhancement of contaminant removal accomplished by the efficient function

of

transconjugants.15,39

Further,

the

long-term

effect

of

genetic

bioaugmentation might be inferred by the variation of the indigenous bacterial community. The success of genetic bioaugmentation can be influenced by many factors. One vital factor is that the acquisition of plasmids by recipient cells can affect the fitness of the host.39 Despite the adaptive benefits conferred by plasmids, they also produce a burden (fitness cost) in the host.40 The costs that are associated with the replication, repair and expression of plasmid genes, which occupy the energy and regulation machinery of the host cells.41 However, catabolic plasmids provide fitness benefit to the host cells under selective pressure, which counterbalances the fitness costs. Supplementation with dibenzofuran by feeding in each cycle of SBR imposed a periodic selection pressure on the sludge bacteria. If dibenzofuran could not be degraded efficiently, this selection pressure became continuous. In the present study, there were several types of carbon sources available for bacterial growth in the



synthetic waste, including glucose, sodium acetate and ethanol except dibenzofuran. The total carbon source in the influent corresponded to a COD of 3000 mg L−1, which was sufficient to support the growth of the sludge bacteria. Catabolic plasmid acquisition could confer on the recipient bacteria the ability to use an additional carbon source and proliferate in the vacant ecological niches, thus gaining a competitive advantage over other sludge bacteria. Contaminants as selection pressure play an important role in genetic bioaugmentation for environment bioremediation.14,39 Although transconjugants were detected in the SBR from day 2, their proliferation was accomplished after 50 d along with the decline of the donor strain p52. Top et al. proposed that the transconjugants need to have a selective advantage over the other indigenous bacteria if they are to grow to high densities.14 As most dibenzofuran was used by the donor strain p52 within 50 d, this was insufficient to confer a selective advantage for the transconjugants. From day 51, the cycle time was reduced from 24 h to 12 h, which created a stronger selective advantage for the transconjugants over the other sludge bacteria, and the transconjugants consequently increased rapidly in number. It was noted that the enhanced selection pressure also led to an adaptive selection of catabolic genes within the seed sludge bacteria, which served as a catabolic gene pool.42 In a previous study, Ikuma and Gunsch suggested that environmentally relevant concentrations of toluene might not exert sufficient selection pressure for transfer of the TOL plasmid.15 Although the occurrence of transconjugants in the SBR might not depend on the degree of selection pressure, their proliferation appeared to


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be influenced by the level of selection pressure. It was therefore proposed that the effectiveness of contaminants in exerting selection pressure for the conjugation of catabolic plasmids may differ significantly depending on the compound and the plasmid.15,39 Phylogenetic relationship between donor and recipient strains, were considered to influence the success of genetic bioaugmentation.39 In the present study, it was indicated that some transconjugants affiliated to Actinobacteria, distributing across two families, were closely related to the actinomyces donor. A phylogenetic analysis of microbial genomes using a networks approach suggested that horizontal gene transfer occurs frequently between phylogenetically related species, and a crucial factor in this is that the donors and recipients possess similar genomic G+C contents.43 The G+C content of the strain p52 genome was 67.84%, compared with 65.7% and 66.2% for pDF01 and pDF02, respectively. According to the genome sequence data in the GenBank database, the average genomic G+C content for the transconjugants,

including

Enterobacter

(55.4%),

Klebsiella

(56.5%),

Corynebacterium (60.5%), and Arthrobacter (65.9%). Maintenance of pDF01 and pDF02 and the expression of catabolic genes (revealed by degradation tests) in different transconjugants showed an ambiguous relationship with G+C contents. Similarly, Ikuma and Gunsch proposed that differences in the G+C content and phylogenetic relationships between donor and transconjugant strains did not affect the stability of the TOL plasmid in transconjugants.16 However, they also demonstrated that the differences negatively affected the expression of the catabolic gene.16 This



implied that the impact of the biotic factors on genetic bioaugmentation was connected to the plasmids and the donor strain. It should be noted that some of the transconjugants were affiliated with Gammaproteobacteria. Bacteria that belong to the Gammaproteobacteria have been identified as the most frequent recipients for IncP plasmid transfer among the indigenous bacteria.44 Garbisu et al. proposed that the influence of abiotic and biotic factors has been studied in only a limited number of bacterial strains and under controlled simplified environmental conditions.39 Thus, much remain to be done to clarify the effect of biotic factors in the natural environment. Proliferation of transconjugants leading to a change in the bacterial community has been reported.17,25 During SBR operation, the less active bacteria tended to disperse and detach from flocs,45 which may then be discharged in the effluent after each cycle. Since only the relatively abundant members of the sludge bacterial communities could be detected by high-throughput 16S rRNA gene sequencing, the richness and evenness of OTUs among sludge bacterial communities differed for the bioaugmented reactor, the non-bioaugmented control, and the seed sludge. In wastewater treatment facilities, rapid evolution of bacteria of large population sizes and short generation times is facilitated by the introduction of conjugative plasmids under continuously imposed selection pressure.30,43 Because the fitness costs associated with plasmid acquisition can be offset by the benefits to bacterial populations of stability against potential environmental changes, the diversity of bacteria with the capacity to metabolize the target contaminant resulted from the vast number of potential


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recipients in the sludge. Hence, the bacterial community in the bioaugmented reactor had an even distribution of OTUs thereby displayed relatively high diversity. The initial evenness of bacterial community is a key factor in preserving the functional stability of an ecosystem resistant to environmental stress.46 Therefore, in the present study, genetic bioaugmentation of the SBR not only enhanced removal of the target contaminant but also might confer an ecological benefit to the bacterial community regarding its long-term functional stability.

ASSOCIATED CONTENT Supporting Information. This material is available free of charge via the internet at http://pubs.acs.org

ACKNOWLEDGMENTS Funding was provided by the Natural Science Foundation of China (grant no. 21377069).

REFERENCES 1.

Nadal, M.; Marquès, M.; Mari, M.; Domingo, J.L. Climate change and

environmental concentrations of POPs: A review. Environ. Res. 2015, 143 (Pt A), 177–185; DOI 10.1016/j.envres.2015.10.012. 2.

Urban, J. D.; Wikoff, D. S.; Bunch, A. T.; Harris, M. A.; Haws, L. C. A review of

background dioxin concentrations in urban/suburban and rural soils across the United



States: implications for site assessments and the establishment of soil cleanup levels. Sci. Total Environ. 2014, 466–467, 586–597; DOI 10.1016/j.scitotenv. 2013.07.065. 3.

Amirova, Z.; Weber, R. Massive PCDD/F contamination at the Khimprom

organochlorine plant in Ufa—a review and recommendations for future management. Environ. Sci. Pollut. Res. 2015, 22(19), 14416–14430; DOI 10.1007/s11356-0155048-8. 4.

Awasthi, A. K.; Zeng, X.; Li, J. Environmental pollution of electronic waste

recycling in India: A critical review. Environ. Pollut. 2016, 211, 259–270; DOI 10.1016/j.envpol.2015.11.027. 5.

Iyer, R.; Aggarwal, J.; Iken, B. A review of the Texas, USA San Jacinto

Superfund site and the deposition of polychlorinated dibenzo-p-dioxins and dibenzofurans in the San Jacinto River and Houston Ship Channel. Environ. Sci. Pollut. Res. 2016, 23 (23), 23321–23338; DOI 10.1007/s11356-016-7501-8. 6.

Field, J. A.; Sierra-Alvarez, R. Microbial degradation of chlorinated dioxins.

Chemosphere 2008, 71 (6), 1005–1018; DOI 10.1016/j.chemosphere.2007.10.039. 7.

Habe, H.; Chung, J. S.; Lee, J. H.; Kasuga, K.; Yoshida, T.; Nojiri, H.; Omori, T.

Degradation of chlorinated dibenzofurans and dibenzo-p-dioxins by two types of bacteria having angular dioxygenases with different features. Appl. Environ. Microbiol. 2001, 67 (8), 3610–3617; DOI 10.1128/AEM.67.8.3610-3617.2001. 8.

Nam, I. H.; Kim, Y. M.; Schmidt, S.; Chang, Y. S. Biotransformation of 1,2,3-Tri-

and 1,2,3,4,7,8-hexachlorodibenzo-p-dioxin by Sphingomonas wittichii strain RW1. Appl. Environ. Microbiol. 2006, 72 (1), 112–116; DOI 10.1128/aem.72.1.112-


Page 30 of 38

Page 31 of 38


116.2006. 9.

Aly, H. A.; Huu, N. B.; Wray, V.; Junca, H.; Pieper, D. H. Two angular

dioxygenases contribute to the metabolic versatility of dibenzofuran-degrading Rhodococcus sp. strain HA01. Appl. Environ. Microbiol. 2008, 74 (12), 3812–3822; DOI 10.1128/AEM.00226-08. 10.

Kasuga, K.; Nitta, A.; Kobayashi, M.; Habe, H.; Nojiri, H.; Yamane, H.; Omori,

T.; Kojima, I. Cloning of dfdA genes from Terrabacter sp. strain DBF63 encoding dibenzofuran 4,4a-dioxygenase and heterologous expression in Streptomyces lividans. Appl. Microbiol. Biotechnol. 2013, 97 (10), 4485–4498; DOI 10.1007/s00253-0124565-3. 11. Jeon, J. R.; Murugesan, K.; Baldrian, P.; Schmidt, S.; Chang, Y. S. Aerobic bacterial catabolism of persistent organic pollutants - potential impact of biotic and abiotic interaction. Curr. Opin. Biotechnol. 2016, 38, 71–78; DOI 10.1016/j.copbio. 2015.12.016. 12. Shintani, M.; Yano, H.; Habe, H.; Omori, T.; Yamane, H.; Tsuda, M.; Nojiri, H. Characterization of the replication, maintenance, and transfer features of the IncP-7 plasmid pCAR1, which carries genes involved in carbazole and dioxin degradation. Appl. Environ. Microbiol. 2006, 72 (5), 3206–3216; DOI 10.1128/aem.72.5.32063216.2006. 13. Sun, J.; Qiu, Y.; Ding, P.; Peng, P.; Yang, H.; Li, L. Conjugative transfer of dioxin-catabolic megaplasmids and bioaugmentation prospects of a Rhodococcus sp. Environ. Sci. Technol. 2017, 51 (11), 6298–6307; DOI 10.1021/acs.est.7b00188.



Page 32 of 38

14. Top, E. M.; Springael, D.; Boon, N. Catabolic mobile genetic elements and their potential use in bioaugmentation of polluted soils and waters. FEMS Microbiol. Ecol. 2002, 42 (2), 199–208; DOI 10.1016/s0168-6496(02)00370-7. 15. Ikuma, K.; Gunsch, C. K. Genetic bioaugmentation as an effective method for in situ bioremediation functionality of catabolic plasmids following conjugal transfers. Bioengineered 2012, 3 (4), 236–241; DOI 10.4161/bioe.20551. 16. Ikuma, K.; Gunsch, C. K. Functionality of the TOL plasmid under varying environmental conditions following conjugal transfer. Appl. Microbiol. Biotechnol. 2013, 97 (1), 395–408; DOI 10.1007/s00253-012-3949-8. 17. Dejonghe, W.; Goris, J.; El Fantroussi, S.; Hofte, M.; De Vos, P.; Verstraete, W.; Top, E. M. Effect of dissemination of 2,4-dichlorophenoxyacetic acid (2,4-D) degradation plasmids on 2,4-D degradation and on bacterial community structure in two different soil horizons. Appl. Environ. Microbiol. 2000, 66 (8), 3297–3304; DOI 10.1128/aem.66.8.3297-3304.2000. 18. Pepper I. L.; Gentry T.J.; Newby D. T.; Roane T. M.; Josephson K. L. The role of cell

bioaugmentation

co-contaminated

soils.

and

gene

Environ.

bioaugmentation Health

Persp.

in 2002,

the 110

remediation (6),

of

943–946;

DOI 10.1289/ehp.02110s6943. 19. Digiovanni G. D.; Neilson J. W.; Pepper I. L.; Sinclair N. A. Gene transfer of Alcaligenes eutrophus JMP134 plasmid pJP4 to indigenous soil recipients. Appl. Environ. Microbiol. 1996, 62 (7), 2521–2526. 20. Newby,

D.

T.;

Gentry,

T.

J.;

Pepper,


I.

L.

Comparison

of

Page 33 of 38


2,4-dichlorophenoxyacetic acid degradation and plasmid transfer in soil resulting from bioaugmentation with two different pJP4 donors. Appl. Environ. Microbiol. 2000, 66 (8), 3399–3407; DOI 10.1128/aem.66.8.3399-3407.2000. 21. Roane, T. M.; Josephson, K. L.; Pepper, I. L. Dual-bioaugmentation strategy to enhance remediation of cocontaminated soil. Appl. Environ. Microbiol. 2001, 67(7), 3208–3215; DOI 10.1128/AEM.67.7.3208-3215.2001. 22. Ikuma, K.; Holzem, R. M.; Gunsch, C. K. Impacts of organic carbon availability and recipient bacteria characteristics on the potential for TOL plasmid genetic bioaugmentation in soil slurries. Chemosphere 2012, 89 (2), 158–163; DOI 10.1016/j.chemosphere.2012.05.086. 23. Bathe, S. Conjugal transfer of plasmid pNB2 to activated sludge bacteria leads to 3-chloroaniline degradation in enrichment cultures. Lett. Appl. Microbiol. 2004, 38 (6), 527–531; DOI 10.1111/j.1472-765X.2004.01532.x. 24. Bathe, S.; Schwarzenbeck, N.; Hausner, M. Plasmid-mediated bioaugmentation of activated sludge bacteria in a sequencing batch moving bed reactor using pNB2. Lett. Appl. Microbiol. 2005, 41 (3), 242–247; DOI 10.1111/j.1472-765X.2005. 01754.x. 25. Bathe, S.; Schwarzenbeck, N.; Hausner, M. Bioaugmentation of activated sludge towards 3-chloroaniline removal with a mixed bacterial population carrying a degradative plasmid. Bioresour. Technol. 2009, 100 (12), 2902–2909; DOI 10.1016/j.biortech.2009.01.060. 26. Nancharaiah, Y. V.; Joshi, H. M.; Hausner, M.; Venugopalan, V. P.



Bioaugmentation of aerobic microbial granules with Pseudomonas putida carrying TOL plasmid. Chemosphere 2008, 71 (1), 30–35; DOI 10.1016/j.chemosphere.2007. 10.062. 27. De Gelder, L.; Vandecasteele, F. P. J.; Brown, C. J.; Forney, L. J.; Top, E. M. Plasmid donor affects host range of promiscuous IncP-1β plasmid pB10 in an activated-sludge microbial community. Appl. Environ. Microbiol. 2005, 71 (9), 5309– 5317; DOI 10.1128/AEM.71.9.5309-5317.2005. 28. Quan, X. C.; Ma, J. Y.; Xiong, W. C.; Yang, Z. F. Effects of gene-augmentation on the formation, characteristics and microbial community of 2,4-dichlorophenoxyacetic acid degrading aerobic microbial granules. J. Hazard. Mater. 2011, 196, 278–286; DOI 10.1016/j.jhazmat.2011.09.023. 29. Inoue, D.; Yamazaki, Y.; Tsutsui, H.; Sei, K.; Soda, S.; Fujita, M.; Ike, M. Impacts of gene bioaugmentation with pJP4-harboring bacteria of 2,4-D-contaminated soil slurry on the indigenous microbial community. Biodegradation 2012, 23 (2), 263–276; DOI 10.1007/s10532-011-9505-x. 30. Mohan, S. V.; Falkentoft, C.; Nancharaiah, Y. V.; Sturm, B. S.; Wattiau, P.; Wilderer, P. A.; Wuertz, S.; Hausner, M. Bioaugmentation of microbial communities in laboratory and pilot scale sequencing batch biofilm reactors using the TOL plasmid. Bioresour. Technol. 2009, 100 (5), 1746–1753; DOI 10.1016/j.biortech.2008.09.048. 31. Herrero, M.; Stuckey, D. C. Bioaugmentation and its application in wastewater treatment: A review. Chemosphere 2015, 140, 119–128; DOI 10.1016/j.chemosphere. 2014.10.033.


Page 34 of 38

Page 35 of 38


32. Peng, P.; Yang, H.; Jia, R.; Li, L. Biodegradation of dioxin by a newly isolated Rhodococcus sp. with the involvement of self-transmissible plasmids. Appl. Microbiol. Biotechnol. 2013, 97 (12), 5585–5595; DOI 10.1007/s00253-012-4363-y. 33. Shintani, M.; Fukushima, N.; Tezuka, M.; Yamane, H.; Nojiri, H. Conjugative transfer of the IncP-7 carbazole degradative plasmid, pCAR1, in river water samples. Biotechnol. Lett. 2008, 30 (1), 117–122; DOI 10.1007/s10529-007-9519-y. 34. Van Loosdrecht, M. C. M.; Nielsen, P. H.; Lopez-vazquez, C. M.; Brdjanovic, D., Eds. Experimental methods in wastewater treatment; IWA publishing: London, 2016. 35. Harms, G.; Layton, A. C.; Dionisi, H. M.; Gregory, I. R.; Garrett, V. M.; Hawkins, S. A.; Robinson, K. G.; Sayler, G. S. Real-time PCR quantification of nitrifying bacteria in a municipal wastewater treatment plant. Environ. Sci. Technol. 2003, 37(2), 343–351; DOI 10.1021/es0257164. 36. Yun, J. J.; Heisler, L. E.; Hwang, I. I. L.; Wilkins, O.; Lau, S. K.; Hyrcza, M. Jayabalasingham, B.; Jin, J.; McLaurin, J.; Tsao, M. S.; Der, S. D. Genomic DNA functions as a universal external standard in quantitative real-time PCR. Nucleic Acids Res. 2006, 34 (12), e85; DOI 10.1093/nar/gkl400. 37. Herlemann, D. P.; Labrenz, M.; Jürgens, K.; Bertilsson, S.; Waniek, J. J.; Andersson, A. F. Transitions in bacterial communities along the 2000km salinity gradient

of

the

Baltic

Sea.

ISME

J.

2011,

5(10),

1571–1579;

DOI

10.1038/ismej.2011.41. 38. Chiu, C. H.; Chao, A. Estimating and comparing microbial diversity in the presence of sequencing errors. Peer J. 2016, 4, e1634; DOI 10.7717/peerj.1634.



39. Garbisu, C.; Garaiyurrebaso, O.; Epelde, L.; Grohmann, E.; Alkorta, I. Plasmidmediated bioaugmentation for the bioremediation of contaminated soils. Front. Microbiol. 2017, 8, 1966; DOI 10.3389/fmicb.2017.01966. 40. San Millan, A.; MacLean, R. C. Fitness costs of plasmids: a limit to plasmid transmission. Microbiol. Spectr. 2017, 5 (5); DOI 10.1128/microbiolspec.MTBP0016-2017. 41. Werisch, M.; Berger, U.; Berendonk, T. U. Conjugative plasmids enable the maintenance of low cost non-transmissible plasmids. Plasmid 2017, 91, 96–104; DOI 10.1016/j.plasmid.2017.04.004. 42. Chakraborty, J.; Das, S. Molecular perspectives and recent advances in microbial remediation of persistent organic pollutants. Environ. Sci. Pollut. Res. 2016, 23 (17), 16883–16903; DOI 10.1007/s11356-016-6887-7. 43. Popa, O.; Dagan, T. Trends and barriers to lateral gene transfer in prokaryotes. Curr. Opin. Biotechnol. 2011, 14 (5), 615–623; DOI 10.1016/j.mib.2011.07.027. 44. Sørensen, S. J.; Bailey, M.; Hansen, L. H.; Kroer, N.; Wuertz, S. Studying plasmid horizontal transfer in situ: a critical review. Nat. Rev. Microbiol. 2005, 3(9), 700–710; DOI 10.1038/nrmicro1232. 45. Morgan-Sagastume, F.; Larsen, P.; Nielsen, J. L.; Nielsen, P. H. Characterization of the loosely attached fraction of activated sludge bacteria. Water Res. 2008, 42 (4-5), 843–854; DOI 10.1016/j.watres.2007.08.026. 46. Wittebolle, L.; Marzorati, M.; Clement, L.; Balloi, A.; Daffonchio, D.; Heylen, K.; De Vos, P.; Verstraete, W.; Boon, N. Initial community evenness favours functionality


Page 36 of 38

Page 37 of 38


under selective stress. Nature 2009, 458(7238), 623–626; DOI 10.1038/nature07840.



83x47mm (300 x 300 DPI)


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Genetic bioaugmentation of activated sludge with dioxin-catabolic

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