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Autochthonous bioaugmentation modified bacterial diversity of phenanthrene degraders in PAH-contaminated wastewater as revealed by DNA-stable isotope probing Jibing Li, Chunling Luo, Dayi Zhang, mengke song, Xixi Cai, Longfei Jiang, and Gan Zhang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b05646 • Publication Date (Web): 29 Jan 2018 Downloaded from http://pubs.acs.org on January 31, 2018

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Environmental Science & Technology

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Autochthonous bioaugmentation modified bacterial diversity of phenanthrene

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degraders in PAH-contaminated wastewater as revealed by DNA-stable isotope

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probing

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Jibing Li,†, Ѱ Chunling Luo,†, ‡* Dayi Zhang,§ Mengke Song, ‡ Xixi Cai,# Longfei Jiang,† Gan

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Zhang†

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†

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‡

Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China College of Natural Resources and Environment, South China Agricultural University, Guangzhou,

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510642, China

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§

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#

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350002, China

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School of Environment, Tsinghua University, Beijing 100084, China College of Resources and Environment, Fujian Agriculture and Forestry University, Fuzhou,

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University of Chinese Academy of Sciences, Beijing, 100049, China

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*Corresponding author: Dr. Chunling Luo

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E-mail: [email protected]

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Tel.: +86-20-85290290; Fax: +86-20-85290706

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ACS Paragon Plus Environment


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TOC

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ABSTRACT

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To reveal the mechanisms of autochthonous bioaugmentation (ABA) in wastewater contaminated

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with polycyclic aromatic hydrocarbons (PAHs), DNA-stable-isotope-probing (SIP) was used in

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the present study with the addition of an autochthonous microorganism Acinetobacter tandoii LJ-5.

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We found LJ-5 inoculum produced a significant increase in phenanthrene (PHE) mineralization,

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but LJ-5 surprisingly did not participate in indigenous PHE degradation from the SIP results. The

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improvement of PHE biodegradation was not explained by the engagement of LJ-5 but attributed

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to the remarkably altered diversity of PHE degraders. Of the major PHE degraders present in

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ambient wastewater (Rhodoplanes sp., Mycobacterium sp., Xanthomonadaceae sp. and

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Enterobacteriaceae sp.), only Mycobacterium sp. and Enterobacteriaceae sp. remained functional

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in the presence of strain LJ-5, but five new taxa Bacillus, Paenibacillus, Ammoniphilus,

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Sporosarcina and Hyphomicrobium were favoured. Rhodoplanes, Ammoniphilus, Sporosarcina

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and Hyphomicrobium were directly linked to, for the first time, indigenous PHE biodegradation.

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Sequences of functional PAH-RHDα genes from heavy fractions further proved the change in PHE

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degraders by identifying distinct PAH-ring hydroxylating dioxygenases between ambient

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degradation and ABA. Our findings indicate a new mechanism of ABA, provide new insights into

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the diversity of PHE-degrading communities, and suggest ABA as a promising in situ

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bioremediation strategy for PAH-contaminated wastewater.

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1. INTRODUCTION

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Polycyclic aromatic hydrocarbons (PAHs) are ubiquitous environmental pollutants consisting of

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two or more fused aromatic rings.1 These compounds have the potential to pose serious health

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risks to all organisms.2 They are of great environmental concern due to their high toxicity,

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mutagenicity and carcinogenicity.3 Phenanthrene (PHE) is employed as a model PAH compound

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due to its ubiquity in nature and fused-ring angular structure.4 Bioremediation is considered as a

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cost-effective and eco-friendly approach for elimination of PAHs from natural environments.5

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PAHs in a polluted environment can be partially degraded by the autochthonous microbial

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population via natural attenuation. When indigenous microbes lack the capacity to degrade PAHs,

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the introduction of PAH degraders isolated from other PAH-contaminated sites,6,

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bioaugmentation (BA), is a promising technology to encourage PAH degradation.8, 9 However, the

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efficacy of BA is variable, since the survival and the PAH-degrading ability of introduced

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microorganisms are highly dependent on environmental conditions.10,

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difficulties, autochthonous bioaugmentation (ABA) has been proposed as an alternative method.12

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ABA is defined as a BA technology that uses indigenous degraders in the polluted sites (soil,

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sediment and water) with the aim of accelerating and enhancing the biodegradation potential.13

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11

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known as

To overcome these

Many microorganisms capable of degrading PAHs have been isolated using cultivation-based

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approaches and have been evaluated for their potential application in ABA strategies.7,

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Cultivation-based methods are used to determine the phenotypes and metabolic characteristics of

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PAH degraders and the functional genes associated with PAH metabolism such as PAH-ring

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hydroxylating dioxygenase.16 Particularly, the gene encoding the alpha subunit of the PAH-ring

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hydroxylating dioxygenase (PAH-RHDα) is usually used as the biomarker to quantify the bacterial 4


14, 15

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populations capable of degrading PAH.16 However, these approaches share an critical drawback in

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that the majority of microbes are uncultivable,17 and many autochthonous PAH-degrading

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microorganisms remain resistant to traditional cultivation approaches.4,

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cultivation greatly underestimates the microbial diversity in ecosystems and fails to explain the

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complex interactions between microorganisms and the natural environment.19

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In addition, direct

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In recent years, cultivation-independent methods have been used to evaluate the prokaryotic

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diversity of complex systems and estimate the PAH degradation potential of indigenous

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consortia.20-22 High-throughput sequencing has revolutionised our ability to investigate the

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microbial communities in environmental samples by providing higher resolution of microbial taxa

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compared with conventional cloning techniques.23 Nevertheless, the metabolic features of

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organisms are difficult to infer accurately using this approach.23 Stable-isotope probing (SIP) is a

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cultivation-independent technique that circumvents the requirement of isolating an organism to

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assess metabolic responses and link its identity to function.24 This technique relies on the

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incorporation of a stable isotope-labelled substrate and the identification of active microorganisms

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by analysing isotope-enriched cellular components, such as DNA, RNA and protein.4, 25 We can

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therefore characterise the target microorganisms resistant to cultivation.18 To date, SIP has been

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used to identify a large number of indigenous bacteria capable of degrading PAHs.20, 26-28

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In our previous work, we isolated a novel PHE degrading strain, Acinetobacter tandoii LJ-5,

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from PAH-contaminated wastewater (37°49′N, 118°25′E; altitude, 37.49 m) collected in Shandong

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Province, China in September 2015.18 A. tandoii LJ-5 is a major autochthonous PHE degrader due

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to its PHE metabolism confirmed by its enrichment in the heavy DNA fraction in the presence of

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C-PHE.18 A. tandoii LJ-5 is a Gram negative, rod-shaped, obligate aerobe lacking flagella. This 5



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strain grew well in minimal medium supplemented with 100–1,000 mg/L PHE under optimal

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conditions (pH 7.0 and 30°C), and a satisfactory PHE degradation efficiency (> 80%) was

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achieved within 3 days for low PHE concentrations (100 and 200 mg/L).18

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Although the ABA strategy can significantly promote PHE mineralisation by introducing

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autochthonous strains, altering the overall microbial activity and population,15, 29 no studies have

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confirmed the degrading capacity of the reintroduced degraders, and little is known about changes

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in the diversity of indigenous PHE degraders during ABA. In the present study, wastewater

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microcosms were established to evaluate the bioremediation potential of ABA in

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PAH-contaminated wastewater with A. tandoii LJ-5 inoculum. Additionally, DNA-SIP was used to

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investigate the changes in the diversity of indigenous PHE degraders by comparing treatments in

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the presence versus absence of LJ-5, and to link the indigenous bacterial taxa with their PHE

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biodegradation phenotypes. The PHE-degrading bacteria in different wastewater treatments were

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successfully characterised using DNA-SIP and high-throughput sequencing. Furthermore, the

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functional PAH-RHD gene was investigated by analysing relevant sequences amplified from the

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heavy fractions of

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C-PHE treatments. We have provided useful information regarding the

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mechanisms of ABA in aiding bioremediation of PAH-contaminated wastewater and other

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potentially polluted environments.

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2. MATERIALS AND METHODS

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2.1. Sample collection. Water samples (~10 L) were collected in March 2016 from industrial

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wastewater at the same oil refinery (37°49′N, 118°25′E; altitude, 37.49 m) in Shandong Province

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(China) as described in our previous work.18 After transport to the laboratory, the wastewater was

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stored immediately at 4°C. Portions of the samples were stored at −80°C for subsequent DNA

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extraction, and the remainder was promptly used in PHE degradation and SIP experiments. PAHs

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identified in the wastewater are listed in the Supporting Information Table S1 (determined using

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gas chromatography–mass spectrometry as described below).

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2.2. Cultivation of bacteria. A. tandoii LJ-5 was previously isolated from wastewater in

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September 2016 at the same oil refinery, using PHE as the carbon source,18 and deposited in the

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Marine Culture Collection of China (MCCC; 1K02142). A. tandoii LJ-5 is a promising candidate

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for remediation of PAH-contaminated water, as it has been shown to be an indigenous PHE

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degrader by both DNA-SIP and direct cultivation. A. tandoii LJ-5 was stored in glycerol solution

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(15% v/v) at −80°C, thawed and re-cultivated as described previously.18 Briefly, it was cultivated

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in minimal medium supplemented with 1,000 mg/L PHE in the dark for 18–24 h with shaking at

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180 rpm at 30°C. Enriched cultures were transferred into a sterile centrifuge tube. Pelleted cells

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were obtained after centrifugation for 10 min (3000 × g; Eppendorf, Centrifuge 5810R, Hamburg,

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Germany) and resuspended in phosphate-buffered saline (PBS). Finally, the population of A.

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tandoii LJ-5 was determined to be approximately 2 × 108 colony forming units/mL (CFU/mL),

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using the dilution plate counting method.30 Details of the surface tension measurement was

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provided in the Supporting Information.

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2.3. SIP microcosms. Microcosms were established in 150 mL serum bottles with 50 mL 7



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wastewater with and without inoculation of strain LJ-5. For treatment with strain LJ-5, 100 µL of

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the cell/PBS suspension were added to the wastewater, thereby establishing an initial inoculum

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population of approximately 4 × 105 CFU/mL at the beginning of each treatment.10 The unlabelled

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PHE (99%) or

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Tewksbury, MA, USA) was added to the above bottles at a final concentration of 10 mg/L. The

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biotic treatments, which allowed comparisons of ambient degradation (AD) and ABA, are referred

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to as

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(13C-PHE without strain LJ-5), and

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PHE (non-PHE control) and treatments with unlabelled PHE in filter-sterilised wastewater (sterile

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control) were also established. Each treatment was performed in triplicate. The microcosms were

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incubated using the method described in our previous work.18 The residual PHE on days 3 and 6

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after incubation was extracted and detected. Almost all of the PHE was degraded on day 6 (Figure

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S3), indicating that the majority of PHE was mineralised and incorporated by the microbes during

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the first 3 days. Thus, samples from each treatment were collected on day 3 for PHE analysis and

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DNA extraction to avoid cross-feeding.

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C-labeled PHE (13C14-PHE, 99%, Cambridge Isotope Laboratories, Inc.,

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C_AD (12C-PHE without strain LJ-5), 13

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C_ABA (12C-PHE with strain LJ-5),

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C_AD

C_ABA (13C-PHE with strain LJ-5). Microcosms without

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2.4. Nucleic acid extraction and ultracentrifugation. DNA extraction from triplicate

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microcosms (25 mL) of each treatment was performed and quantified as described previously.18

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Ultracentrifugation of DNA was conducted as described in our previous report.31 Briefly,

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approximately 3 µg DNA from each microcosm were added to Quick-Seal polyallomer tubes (13

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× 51 mm, 5.1 mL, Beckman Coulter, Pasadena, CA, USA) and mixed with Tris-EDTA (pH

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8.0)/caesium chloride solution at a final buoyant density (BD) of ~1.77 g/mL. After balancing and

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heat-sealing, density gradient centrifugation was performed in an ultracentrifuge (Optima 8


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L-100XP, Beckman Coulter, USA) at 45,000 rpm (20°C) for 48 h. Centrifuged gradients were

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fractionated into different fractions of 400 µL each. The DNA fractions were purified after the BD

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of each fraction was measured. The relationships between BD and the fraction number or DNA

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concentration for all four biotic treatments are listed in Figures S1 and S2, respectively.

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2.5. Quantitative polymerase chain reaction (qPCR). Abundance of the bacterial 16S

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ribosomal RNA (rRNA) gene in each fraction was determined by qPCR using the universal

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bacterial primer pair Bac519F/Bac907R (Table 1). The 20 µL PCR mixture contained 10 µL

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SYBR green PCR Premix Ex Taq II (Takara Bio Inc., Japan), 0.2 µL each primer (10 µM), and 1

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µL DNA template. PCR was performed using the ABI 7500 real-time PCR system (Applied

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Biosciences, USA). Tenfold serial dilutions of known copy numbers of the plasmid DNA extracted

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from Escherichia coli were generated to produce a standard curve. The reactions were conducted

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as follows: initial denaturation at 94°C for 10 min, followed by 40 cycles at 94°C for 30 s, 55°C

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for 30 s, and 72°C for 15 s. The SYBR green signal was measured after a 20 s step at 72°C in each

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cycle. The melt curve was obtained by heating from 60 to 95°C at a rate of 0.2°C/s.

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2.6. High-throughput sequencing and computational analyses. The hypervariable V4

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region of bacterial 16S rRNA gene fragments from the DNA fractions derived from the 12C_AD,

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(Table S2), as described previously.32 Unique heptad-nucleotide sequences (12 bases) were added

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to the reverse primers as barcodes to assign sequences to the different fractions. PCR was

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performed using the method described by Song et al.33 Sequencing was conducted using 2 × 250

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base pair paired-end technology on the Illumina MiSeq sequencer in a standard pipeline. The

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qualified sequences were analysed after read filtering34,

C_ABA,

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C_AD and

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C_ABA microcosms was amplified using the 515f/806r primer set

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and assigned using an operational

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taxonomic unit (OTU)-based method to generate microbiome profiles, as described previously.36-38

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OTU assignment was based on a 97% cut-off in the present study.

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The relative abundance of each OTU was determined, and the top 100 most relatively

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abundant were selected for analysis according to previous studies.18, 39 The PHE degraders were

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identified by OTUs enriched in the heavy fractions from the

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compared with the 12C_AD and 12C_ABA samples. In the present study, we identified four OTUs

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(OTU_56, OTU_65, OTU_106 and OTU_111) from the

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(OTU_8, OTU_15, OTU_23, OTU_43, OTU_83, OTU_106 and OTU_111) from the

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treatment. These sequences were further trimmed using the Greengenes database and aligned to

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Bacillus spp. (OTU_8, accession number: MF037428), Ammoniphilus spp. (OTU_15, accession

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number: MF037425), Hyphomicrobium spp. (OTU_23, accession number: MF037426),

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Paenibacillus spp. (OTU_43, accession number: MF037429), Rhodoplanes spp. (OTU_56,

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accession number: MF037430), Xanthomonadaceae spp. (OTU_65, accession number:

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MF037431), Sporosarcina spp. (OTU_83, accession number: MF037427), Mycobacterium spp.

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(OTU_106, accession number: MF037432), and Enterobacteriaceae spp. (OTU_111, accession

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number: MF037433), respectively. Phylogenetic analysis of these sequences was performed as

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described previously.4

13

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C_AD and

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C_ABA samples

C_AD treatment and seven OTUs 13

C_ABA

2.7. Detection of PAH-RHD genes. The PAH-RHDα genes in the heavy DNA fractions of

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642f/933r) and Gram negative (610f/911r) degraders,33 respectively (Table S2). Gradient PCR and

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the amplification reactions were performed as described previously.16 However, in the present

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study, only the PAH-RHDα GP primer set produced a strong and specific amplicon and was

C-PHE treatment samples were amplified using two primer sets for Gram positive (GP;

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selected for analysis. We identified two PAH-RHDα GP genes (PAH-RHDα C1 and C15) from the

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C_AD treatment, and three PAH-RHDα GP genes (PAH-RHDα L1, L6 and C10) from the

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C_ABA treatment. These sequences are available in GenBank with the following accession

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numbers: PAH-RHDα C1: MF037434, PAH-RHDα C15: MF037435, PAH-RHDα L1: MF037436,

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PAH-RHDα L6: MF037437 and PAH-RHDα L10: MF037438.

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2.8. Chemical analysis. The concentrations of PHE and other PAHs in each microcosm were

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analysed by gas chromatography (model 7890, Agilent, Santa Clara, CA, USA) using a capillary

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column (DB-5MS, 30 m, 0.25 mm, 0.25 µm) and a mass spectrometric detector (model 5975,

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Agilent) on days 0, 3 and 6, as described previously.18 Briefly, the water sample was spiked with

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1,000 ng deuterated PAHs and extracted twice with dichloromethane. The extracted organic phase

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was concentrated to approximately 0.5 mL and purified using a silica gel/alumina column (8 mm

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i.d.). The eluent was concentrated to approximately 50 µL using a gentle stream of N2, and 1,000

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ng hexamethylbenzene were added as an internal standard to all samples before instrumental

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analysis. The components and concentrations of the deuterated PAHs, standards and the internal

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standard are listed in Table S3.

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3. RESULTS

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3.1. PHE degradation. The PHE biodegradation curves in the 12C_AD, 12C_ABA, 13C_AD

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and 13C_ABA treatments are shown in Figure S3. The recovery rates of PHE during the extraction

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procedure were 79-85% in this study. The PHE concentration in the sterile control exhibited less

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decrease than those in the biotic treatments, consistent with our previous observations. On day 3,

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there was 36.7%, 38.2%, 22.6% and 23.8% residual PHE in the 12C_AD, 13C_AD, 12C_ABA and

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13

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treatments. No significant difference (p > 0.05) was observed between either the

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13

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in the

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than those in the

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degradation efficiency increased by 13.8% when strain LJ-5 was added to the wastewater.

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Statistical analysis showed a significant difference between the AD and ABA treatments after 3

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days of incubation, suggesting that ABA strategy by A. tandoii LJ-5 inoculum could significantly

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improve the PHE biodegradation efficiency in PAH-contaminated wastewater. In addition, there

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was little loss of other PAHs in the wastewater over the 3-day period (data not shown).

C_ABA microcosms, respectively, indicating that PHE biodegradation occurred in the biotic 12

C_AD and

C_AD treatments, or the 12C_ABA and 13C_ABA treatments. Moreover, the PHE concentrations 12

C_ABA (22.6%) and 12

13

C_ABA (23.8%) microcosms were significantly lower (p < 0.05)

C_AD (36.7%) and

13

C_AD (38.2%) microcosms, suggesting that the

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3.2. PHE degraders revealed by DNA-SIP. DNA extracted from all four biotic treatments,

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after 3 days of incubation, were separated by isopycnic caesium chloride density gradient

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centrifugation, followed by high-throughput sequencing of each fraction. The relative abundance

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of the total 16S rRNA defined by the genus indicated a slight difference in indigenous microbial

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communities between the samples from either the

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12

12

C_AD and

13

C_AD treatments, or the

C_ABA and 13C_ABA treatments (Figure S4). Moreover, the addition of strain LJ-5 induced a 12


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significant change in the composition and structure of microbial communities in the ABA

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treatments, compared with the AD microcosms. Obviously, the relative abundances of the genus

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Acinetobacter (2.08%) increased significantly after addition of strain LJ-5 in the ABA treatments,

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and was much higher than that in the AD microcosms (0.05%). The majority of the abundant

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bacteria (> 5% total abundance), including members of the genera Kaistobacter, Rhodanobacter,

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Burkholderia, and unclassified Xanthomonadaceae, were enriched in the AD microcosms (Figure

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S4). The relative abundance of the genus Kaistobacter (38.4%) increased and was significantly

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higher (p < 0.05) in ABA microcosms than those in AD microcosms (22.9%), whereas the relative

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abundances of the genera Rhodanobacter, Burkholderia, and unclassified Xanthomonadaceae

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were significantly lower (p < 0.05) in ABA microcosms (15.8%, 1.9% and 3.3%, respectively)

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than those in AD microcosms (24.3%, 7.2% and 10.5%, respectively). As for rare bacteria, the

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ABA strategy stimulated growth of the genera Conexibacter and Rhodoplanes, owing to their

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higher relative abundances (2.09% and 2.01%) compared with the AD microcosms (0.76% and

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1.23%, p < 0.05). No significant changes were detected in the non-PHE controls (Figure S5).

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The bacterial 16S rRNA gene abundance was quantified by qPCR using DNA recovered

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from each fraction of all samples as the template. As shown in Figure S6, in both AD and ABA

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microcosms after the 3-day incubation, the bacterial 16S rRNA in fractions with higher BDs

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(1.7382 or 1.7556 g/mL; see asterisks in Figures 1 and 2) was significantly higher in

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treatments than those in the 12C-PHE control (marked in grey). The abundance of total 16S rRNA

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of additive bioaugmented strain LJ-5 was 4 × 105 copies/mL, accounting for 2% of that in the

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ABA treatment (1.96 × 107 copies/mL; Figure S7). Accordingly, the relative abundance of the

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genus Acinetobacter after ABA treatment increased to 2.08%, and the total 16S rRNA gene copy 13


13

C-PHE


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number of ABA microcosms was slightly higher than that of raw wastewater (Figure S7),

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indicating that the bioaugmented strain LJ-5 persisted in these treatments. The indigenous microorganisms responsible for

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13

C-PHE assimilation were detected by

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comparing the relative abundances of specific OTUs in the 12C-PHE and 13C-PHE treatments from

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each fraction. As shown in Figure 1, OTU_56, OTU_65, OTU_106 and OTU_111 were enriched

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at higher BDs (1.7382 or 1.7556 g/mL) in the

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treatments. Comparing to the relative abundances of OTU_56, OTU_65, OTU_106 and OTU_111

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in the same fractions of the

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the higher abundance in the heavy fractions from the 13C_AD sample (1.49%, 1.68%, 0.94% and

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4.45%, respectively) indicated that the microorganisms represented by the above OTUs played a

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primary role in PHE degradation. However, the inoculation of strain LJ-5 produced a significant

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change in the diversity of the indigenous PHE-degrading communities. Seven main types of

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bacteria represented by OTU_8, OTU_15, OTU_23, OTU_43, OTU_83, OTU_106 and OTU_111

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at a higher BD (1.7382 or 1.7556 g/mL) were enriched in the

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12

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OTU_43, OTU_83, OTU_106 and OTU_111 were significantly higher in the heavy fractions of

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the 13C_ABA treatment (0.93, 1.71, 1.26, 1.08, 0.11, 0.31 and 7.78%, respectively) than those in

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the

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Nevertheless, OTU_4, representing strain LJ-5, was not enriched in the heavy fractions of the

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13

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BDs, 1.7382 and 1.7556 g/mL) of the 13C_ABA microcosm (0.02 and 0.02%, respectively) than in

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the 12C_ABA microcosm (0.13 and 0.05%, respectively). Although PHE-degrading communities

12

13

C_AD microcosms, but not in the

12

C_AD

C_AD treatment (0.18%, 0.15%, 0.11% and 0.82%, respectively),

13

C_ABA sample, but not in the

C_ABA sample (Figure 2). Similarly, the relative abundances of OTU_8, OTU_15, OTU_23,

12

C_ABA microcosm (0.05, 0.08, 0.01, 0.01, 0.007, 0.13 and 1.55%, respectively).

C_ABA treatment samples because of its lower relative abundance in the heavy fractions (higher

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differed between the AD and ABA treatments, OTU_106 and OTU_111 played a vital role in PHE

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degradation under both conditions.

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Figure 3 shows phylogenetic information for the PHE degraders represented by the above

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OTUs. OTU_8, OTU_15, OTU_43 and OTU_83 belong to the genus Bacillus (family

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Bacillaceae), Ammoniphilus (family Aneurinibacillus), Paenibacillus (family Paenibacillaceae)

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and Sporosarcina (family Planococcaceae), respectively, within the same order Bacillales

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(phylum Firmicutes, class Bacilli). OTU_15 shares 100% similarity with Ammoniphilus resinae

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CC-RT-E (HM193518) and forms a subclade with a high bootstrap value of 99. OTU_43 exhibits

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98.02% similarity with the partial 16S rRNA gene sequence of strain Paenibacillus xylanisolvens

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X11-1 (AB495094). OTU_83 has 100% similarity with strains Sporosarcina ureae DSM 2281

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(AF202057), Sporosarcina aquimarina SW28 (AF202056), Sporosarcina saromensis HG645

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(AB243859) and Sporosarcina luteola Y1 (AB473560) and forms a subclade with a bootstrap

284

value of 84. OTU_23 and OTU_56 are assigned to the genus Hyphomicrobium (family

285

Hyphomicrobiaceae) and Rhodoplanes (family Bradyrhizobiaceae), respectively, within the same

286

order Rhizobiales (phylum Proteobacteria, class Alphaproteobacteria). OTU_23 exhibits 98.81%

287

similarity with the partial 16S rRNA gene sequence of strain Hyphomicrobium denitrificans ATCC

288

51888 (ACVL01000012), and forms a subclade with a high bootstrap value of 96. OTU_56 shares

289

100% similarity to the partial 16S rRNA gene sequence of uncultured Rhodoplanes sp. clone

290

Leob163 (KF226093.1), and forms a subclade with a high bootstrap value of 98. OTU_65 and

291

OTU_111 are classified in the family Xanthomonadaceae (order Xanthomonadales) and

292

Enterobacteriaceae

293

Betaproteobacteria (phylum Proteobacteria). Additionally, OTU_106 is characterised as genus

(order

Enterobacteriales),

respectively,

15


within

the

same

class


Page 16 of 36

294

Mycobacterium (phylum Actinobacteria, class Actinobacteria, order Corynebacteriales, family

295

Mycobacteriaceae) and shares 100% similarity with many strains in this genus, including

296

Mycobacterium brisbanense ATCC 49938 (AY012577), Mycobacterium elephantis 484

297

(AJ010747), Mycobacterium moriokaense DSM 44221 (AJ429044) and Mycobacterium pulveris

298

DSM 44222 (AJ429046), and forms a subclade with a high bootstrap value of 100.

299

3.3. Presence of PAH-RHDα genes involved in PHE metabolism. In the present study, the 13

13

300

PAH-RHDα GP genes were analyzed in the heavy fractions of the

301

treatments (marked with an asterisk in Figures 1 & 2). PAH-RHDα C1 and C15 were detected in

302

the heavy fractions of

303

conspicuous change in the types of PAH-RHDα GP genes, compared with the AD microcosms,

304

and PAH-RHDα L1, L6 and L10 were detected (Figure 4). Thereinto, PAH-RHDα C1 and L1 share

305

100% similarity and are the same PAH-RHDα GP gene, indicating that they exist in the heavy

306

fractions of both the

307

bootstrap value of 96 in the phylogenetic tree of amplified PAH-RHDα GP genes, and show 89%

308

similarity with the PAH-RHD genes of Mycobacterium sp. S23 (ALS30442.1). The other

309

PAH-RHD gene in the heavy fractions of the 13C_AD treatment, PAH-RHDα C15, exhibits 99%

310

similarity with the PAH-RHD genes of uncultured bacterium (AMM73080.1). In the heavy

311

fractions of the

312

genes of Mycobacterium vanbaalenii PYR-1 (AAY85176.1), Mycobacterium novocastrense

313

(GAT12202.1) and Terrabacter sp. FLO (ABA87073.1), and PAH-RHDα L10 shows 100%

314

similarity with the PAH-RHD genes of Bacillus sp. CL1-1 (AND66078.1), Enterobacter sp.

315

CL1-2 (AND66079.1) and uncultured Comamonas sp. (CAN85224.1), and forms a subclade with

13

13

13

C_AD treatment (Figure 4). The

C_AD and

13

13

C_AD and

C_ABA

C_ABA treatment produced a

C_ABA microcosms. They form a subclade with a high

C_ABA treatment, PAH-RHDα L6 reveals 99% similarity with the PAH-RHD

16


Page 17 of 36


316

a bootstrap value of 61.

17



317

Page 18 of 36

4. DISCUSSION

318

DNA-SIP has been used successfully to demonstrate that native microorganisms collected at

319

field sites are involved in PHE biodegradation.20, 27 Gutierrez et al. used DNA-SIP with fully

320

13

321

PHE in surface slicks and plume waters.27 Jones et al. performed DNA-SIP with 13C-labelled PHE

322

as part of a larger project investigating strategies for bioremediation of PAH-contaminated soil

323

from a former manufactured-gas plant site.20 Our work applied

324

DNA-SIP and revealed the mechanisms of ABA strategy in PAH-contaminated wastewater with

325

the addition of the autochthonous microorganism A. tandoii LJ-5. Although ABA strategies have

326

been successfully applied for remediation of PAH-contaminated sites,14, 15 limited studies have

327

addressed the changes that occur in the microbial community during the ABA process.14, 40 In the

328

present study, ABA with strain LJ-5 inoculum produced a significant increase in PHE

329

biodegradation efficiency in PAH-contaminated wastewater. This suggests that strain LJ-5 is

330

potentially an ABA agent encouraging the remediation of PAH-contaminated sites. Moreover, the

331

addition of LJ-5 markedly modified the bacterial community structure. By comparing the bacterial

332

community structures of microbial communities between AD and ABA, we identified three

333

bacterial genera, affiliated with Kaistobacter, Conexibacter and Rhodoplanes, that were

334

stimulated by LJ-5 inoculation. Kaistobacter demonstrated relatively high abundance in the

335

indigenous microbial communities of a tailings dump contaminated with antimony41 and the

336

ability to biodegrade both S-ethyldipropylthiocarbamate and atrazine in soils.42 Members of the

337

genus Conexibacter within the class Actinobacteria are affiliated with aromatic hydrocarbon

338

degraders and contain bph genes encoding the biphenyl degradative pathway.43 Rhodoplanes has

C-labelled PHE to link the phylogenetic identity of bacterial taxa with their ability to mineralize

18


13

C-PHE as the substrate in

Page 19 of 36


339

been proposed to accommodate taxa that are primarily phototrophic and dominant in rhizospheric

340

soils and roots.44 Furthermore, the addition of LJ-5 also affected the composition of abundant

341

bacteria, including members of the genera Kaistobacter, Rhodanobacter, Burkholderia, and

342

unclassified Xanthomonadaceae. Among them, Rhodanobacter spp. are dominant in crude oil plus

343

dispersant and rhizospheres of maize cultivars,45, 46 and Burkholderia and unclassified phylotypes

344

within the family Xanthomonadaceae such as Stenotrophomonas, are the dominant genera of PAH

345

degraders with high metabolic activities in PAH-contaminated soils and sediments.47, 48 However,

346

there was no previous evidence directly linking the above activated or abundant bacteria (except

347

Xanthomonadaceae and Burkholderia) to PHE degradation. They might play important roles in

348

stabilizing the microbial community or degrading other organic pollutants in PAH-contaminated

349

wastewater.18

350

Besieds the structure and dynamics of the microbial community, ABA might also shape the

351

composition of functional PHE-degraders community, although this has not yet been reported. Our

352

study is the first to show that ABA also influences the abundance and diversity of PAH-degrading

353

bacteria in PAH-polluted wastewater, suggesting that different PHE degradation activities can be

354

achieved by distinct microbial communities within the same environment. In the AD treatment,

355

the indigenous microorganisms responsible for PHE degradation were affiliated with

356

Rhodoplanes, Mycobacterium, Xanthomonadaceae (genus unclassified) and Enterobacteriaceae

357

(genus unclassified). The genus Rhodoplanes classified under the family Hyphomicrobiaceae of

358

the order Rhizobiales in the class Alphaproteobacteria, was first described in 1994.49 Members of

359

this genus are characterized as Gram negative, phototrophic, rod-shaped, motile and are

360

widespread in aquatic habitats.50 Rhodoplanes is the dominant genus in wastewater treatment 19



361

plants,51 and some strains in this genus can grow on organic carbon sources such as acetate,

362

pyruvate, glucose and fructose.50 However, there is still no proof of its ability to degrade PAHs,

363

including PHE, and their roles in PAH-contaminated sites remain unclear. Our results provide

364

unequivocal evidence that some microbes in this genus are primarily responsible for PHE

365

degradation in PAH-polluted wastewater. Mycobacterium is well known to degrade various

366

environmental contaminants such as chlorinated compounds, polychlorobiphenyls and PAHs.52

367

Degradation of PHE by Mycobacterium has been reported previously,53 and several strains in this

368

genus can effectively degrade high molecular weight PAHs.54 However, prior to this study, PHE

369

degradation by indigenous Mycobacterium has not been identified using DNA-SIP. The families

370

Xanthomonadaceae and Enterobacteriaceae are members of the class Betaproteobacteria.

371

Previous studies have reported that Stenotrophomonas sp. and Pseudoxanthomonas sp. in the

372

family Xanthomonadaceae possess the functions of metabolizing a wide range of PAHs, e.g.

373

naphthalene, PHE, anthracene, fluorene, pyrene and benzo[a]pyrene.55-57 Members of the family

374

Enterobacteriaceae, such as Raoultella sp. and Klebsiella sp., can also degrade PAHs efficiently,

375

including acenaphthene, fluorene, pyrene, benzo[a]pyrene, PHE and fluoranthene.58-60 Here, our

376

results provide strong evidence that some microbes in these families are active PHE degraders in

377

PAH-contaminated wastewater.

378

It should be noted that the bacterial community structure of the wastewater in this study was

379

significantly different from that of our previously reported wastewater,18 thus resulting in the

380

change in diversity of microorganisms responsible for in situ PHE degradation. The majority of

381

the abundant bacteria in wastewater collected in September 2015 were affiliated with the genera

382

Pseudomonas (23.5 ± 1.2%), unclassified Chitinophagaceae (4.7 ± 0.3%) and Comamonadaceae 20


Page 20 of 36

Page 21 of 36


383

(6.4 ± 0.3%).18 We identified four OTUs directly responsible for indigenous PHE biodegradation,

384

including the phylotypes affiliated with Acinetobacter, Sphingobium, Kouleothrix and

385

Sandaracinobacter. In the present study, among the PHE degraders identified in the AD

386

microcosms, only two taxa Mycobacterium and Enterobacteriaceae (genus unclassified) were also

387

identified as PHE degraders in the ABA treatments with strain LJ-5. Five new indigenous

388

PHE-degrading organisms became active, namely phylotypes affiliated with Bacillus,

389

Paenibacillus, Ammoniphilus, Sporosarcina and Hyphomicrobium. The genera Bacillus and

390

Paenibacillus are known to be metabolically versatile, degrading aromatic and hydroxylated

391

aromatic compounds.55, 61, 62 Moreover, these two genera can degrade PHE,55 and appear to be

392

highly competitive by the addition of root exudates.63 However, no studies have considered the

393

roles of Bacillus and Paenibacillus in the remediation of PAH-contaminated sites by ABA

394

strategy. The genus Ammoniphilus is a member of the family Aneurinibacillus in the phylum

395

Firmicutes. Members of this genus are typically Gram variable, oxidase- and catalase-positive,

396

obligately oxalotrophic, and motile by peritrichous flagella.64 Strains of the genus Ammoniphilus

397

are able to utilize high concentrations of ammonium ions.65 Similar to Ammoniphilus,

398

Sporosarcina also belongs to the phylum Firmicutes. It is usually detected in clinical specimens

399

and raw cow’s milk.66 Sporosarcina has been shown to principally degrade aflatoxin B-1, which is

400

the most potent naturally occurring carcinogen known.67 Hyphomicrobium, belonging to the

401

Alphaproteobacteria and Hyphomicrobiaceae family, has versatile metabolic capabilities such as

402

the ability to degrade dichloromethane,68 methamidophos,69 dimethyl sulphide and methanol.70

403

However, the taxa Ammoniphilus, Sporosarcina and Hyphomicrobium have not been linked

404

previously with PHE degradation; thus, their mechanisms of degradation in PAH-contaminated 21



405

sites remain unknown. Our results provide strong evidence that members of the above genera are

406

the primary indigenous PHE degraders induced by ABA strategy in wastewater, expanding our

407

knowledge of the diversity of indigenous PHE-degrading communities.

408

Despite the satisfactory performance of ABA strategies in bioremediation by the addition of

409

autochthonous degraders, no studies have confirmed the activities or degrading capacity of the

410

reintroduced strains. It is worth mentioning that, although the ABA strategy with LJ-5 inoculum

411

encouraged PHE biodegradation, the participation of A. tandoii LJ-5 in PHE degradation in situ

412

was questioned due to their limited enrichment in the heavy DNA fraction according to the

413

DNA-SIP results. Previous researchers have focused primarily on the effects of reintroduced

414

degraders on degradation efficiency and the total microbial activities or population during ABA.15,

415

29

416

strain to the polluted sites.40,

417

autochthonous strain LJ-5 in ABA strategy, modifying the diversity of indigenous PHE degraders

418

instead of participating in in situ PAH degradation. Such a mechanism has never been addressed

419

before and remains unclear in PAH-contaminated wastewater. One possible reason might be

420

attributed to the change in wastewater composition, consequently leaving strain LJ-5 from

421

functional PHE-degraders in previous wastewater to inactive microbes in the present study.

422

Another reason might be the contribution of Acinetobacter biosurfactants to the bioremediation of

423

PAHs and the shift in community composition, which has been confirmed by other researchers.72

424

For instance, a wide range of Acinetobacter sp. can produce biosurfactants (amphiphilic molecules

425

consisting of hydrophilic and hydrophobic domains)73, which might increase the bioavailability of

426

hydrophobic compounds by solubilisation and/or emulsification or alter the cell surface properties

The diversity and functions of the whole microbial community can be enhanced by adding a 71

Our work is the first study to uncover the roles of the

22


Page 22 of 36

Page 23 of 36


427

of other microorganisms.72 In the present study, A. tandoii LJ-5 could reduce the surface tension

428

of water from 70.0 to 35.2 mN/m. Our results indicated that strain LJ-5 might be capable of

429

producing biosurfactants, simultaneously improving PHE bioavailability and activating the

430

functions of other PHE-degrading candidates. Moreover, the relative abundance of the

431

bioaugmented strain LJ-5 was 4×105 copies/mL, accounting for 2% of the total 16S rRNA in ABA

432

treatments and approximately 2 times as that of the initial inoculum abundance in surface tension

433

measurements. It suggested sufficient amount of strain LJ-5 in all the bioaugmentation treatments

434

to enhance PHE degradation via the production of biosurfactants.

435

More evidence of a change in the PHE-degrading community was provided by the sequences

436

of PAH-RHDα genes between AD and ABA treatments. Only one novel PAH-RHDα GP gene,

437

PAH-RHDα C1 and L1, was detected in the heavy DNA fraction of both the

438

13

439

(genus unclassified, OTU_111) which were also both detected as functional PHE-degrading

440

bacteria in the two microcosms. In addition, the distinctive PAH-RHDα L15 gene is potentially the

441

functional gene from Rhodoplanes or Enterobacteriaceae (genus unclassified) as they were all

442

detected in the heavy DNA fraction of the

443

microcosm, PAH-RHDα L6 and L10 genes might be linked to the PHE degraders of Bacillus,

444

Paenibacillus, Ammoniphilus, Sporosarcina or Hyphomicrobium. Here, the active PHE degraders

445

might possess other functional genes that were not targeted by the primers used in this study, and

446

we could not accurately attribute one functional gene to one bacterium for the lack of available

447

database information. Indeed, little information is available regarding the functional genes of

448

certain dominant PHE-degrading genera, such as Rhodoplanes, Ammoniphilus, Sporosarcina and

13

C_AD and

C_ABA treatments, possibly associated with Mycobacterium (OTU_106) or Xanthomonadaceae

13

C_AD treatment. Accordingly, in the

23


13

C_ABA


449

Hyphomicrobium, since this is first report of PHE biodegradation for these genera.

450

ABA by A. tandoii LJ-5 was first applied as a potential strategy to enhance the remediation of

451

PAH-contaminated wastewater in this study. Besides resulting in a significant increase in the PHE

452

biodegradation efficiency, ABA remarkably modified the functional PHE degrading community of

453

the wastewater. The indigenous microorganisms responsible for PHE degradation were affiliated

454

with Rhodoplanes, Mycobacterium, Xanthomonadaceae and Enterobacteriaceae in the AD

455

treatments, whereas five new taxa (Bacillus, Paenibacillus, Ammoniphilus, Sporosarcina and

456

Hyphomicrobium) were acitvated in in situ PHE biodegradation in ABA microcosms. Of all the

457

above PHE degraders, Rhodoplanes, Ammoniphilus, Sporosarcina and Hyphomicrobium were

458

linked to indigenous PHE biodegradation for the first time. Nevertheless, LJ-5 did not participate

459

in indigenous PHE degradation. The change in PAH-RHDα gene diversity further confirmed our

460

findings by different PAH-RHDɑ genes involved in PHE metabolism in the AD and ABA

461

treatments. Collectively, our findings raise a new mechanism of ABA, provide new insights into

462

the diversity of PHE-degrading communities, and suggest ABA as a promising in situ strategy for

463

PAH-contaminated wastewater.

24


Page 24 of 36

Page 25 of 36


464

ASSOCIATED CONTENT

465

Supporting Information

466

(Table S1) Concentrations of PAHs in wastewater; (Table S2) primers used for the PCR of

467

16S rRNA and PAH-RHD gene; (Table S3) the components of deuterated PAHs, standards

468

and internal standard; (Figure S1) correlation between the fraction number and buoyant

469

density (BD; g/mL) of DNA extracted from (a)

470

12

471

(ng/µL) and buoyant density (g/mL) from DNA extracted from (a)

472

treatments, and (b) 12C-ABA and 13C-ABA treatments; (Figure S3) residual PHE percentage

473

in wastewater after 3 days of incubation; (Figure S4) relative abundance of 16S rRNA

474

defined taxa by genus in 12C_AD, 13C_AD, 12C_ABA, and 13C_ABA microcosms; (Figure S5)

475

relative abundance of 16S rRNA defined taxa by genus in non-PHE microcosms; (Figure S6)

476

the quantitative distribution of density-resolved bacterial 16S rDNA obtained from

477

wastewater samples in (a)

478

treatments; (Figure S7) the 16S rRNA gene copies in

479

13

480

Information

C-ABA and

13

12

C-AD and

13

C-AD treatments, and (b)

C-ABA treatments; (Figure S2) correlation between DNA concentration

12

C-AD and

13

C-AD treatments, and (b) 12

C_AD,

12

13

12

C-AD and

C-ABA and

C_AD,

12

13

C-AD

13

C-ABA

C_ABA, and

C_ABA microcosms. There are 12 pages, 3 tables and 7 figures in the Supporting

481 482

ACKNOWLEDGMENTS

483

Financial support was provided by the Scientific and Technological Planning Project of

484

Guangzhou, China (Nos. 201707020034), the National Natural Science Foundation of China (No.

485

41673111), and the Department of Science and Technology of Guangdong province

486

(2016TQ03Z938).

487

25



488

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52. Hormisch, D.; Hormisch, D.; Brost, I.; Kohring, G. W.; Giffhorn, F.; Kroppenstedt, R. M.; Stackebradt, E.; Färber, P.; Holzapfel, W. H., Mycobacterium fluoranthenivorans sp. nov., a Fluoranthene and Aflatoxin B1 Degrading Bacterium from Contaminated Soil of a Former Coal Gas Plant. Systematic and Applied Microbiology 2004, 27, (6), 653-660. 53. Hennessee, C. T.; Li, Q. X., Effects of Polycyclic Aromatic Hydrocarbon Mixtures on Degradation, Gene Expression, and Metabolite Production in Four Mycobacterium Species. Appl. Environ. Microbiol. 2016, 82, (11), 3357-3369. 54. Kwak, Y.; Li, Q. X.; Shin, J. H., Draft genome sequence of Mycobacterium rufum JS14(T), a polycyclic-aromatic-hydrocarbon-degrading bacterium from petroleum-contaminated soil in Hawaii. Standards in genomic sciences 2016, 11, 47. 55. Cebron, A.; Louvel, B.; Faure, P.; France-Lanord, C.; Chen, Y.; Murrell, J. C.; Leyval, C., Root exudates modify bacterial diversity of phenanthrene degraders in PAH-polluted soil but not phenanthrene degradation rates. Environ. Microbiol. 2011, 13, (3), 722-736. 56. Patel, V.; Cheturvedula, S.; Madamwar, D., Phenanthrene degradation by Pseudoxanthomonas sp. DMVP2 isolated from hydrocarbon contaminated sediment of Amlakhadi canal, Gujarat, India. J. Hazard. Mater. 2012, 201, (1), 43-51. 57. Arulazhagan, P.; Al-Shekri, K.; Huda, Q.; Godon, J.; Basahi, J.; Jeyakumar, D., Biodegradation of polycyclic aromatic hydrocarbons by an acidophilic Stenotrophomonas maltophilia strain AJH1 isolated from a mineral mining site in Saudi Arabia. Extremophiles 2017, 21, (1), 163-174. 58. Alegbeleye, O. O.; Opeolu, B. O.; Jackson, V., Bioremediation of polycyclic aromatic hydrocarbon (PAH) compounds: (acenaphthene and fluorene) in water using indigenous bacterial species isolated from the Diep and Plankenburg rivers, Western Cape, South Africa. Braz. J. Microbiol. 2016, 48, (2), 314-325. 59. Ping, L. F.; Zhang, C. R.; Zhang, C. P.; Zhu, Y. H.; He, H. M.; Wu, M.; Tang, T.; Li, Z.; Zhao, H., Isolation and characterization of pyrene and benzo a pyrene-degrading Klebsiella pneumonia PL1 and its potential use in bioremediation. Appl. Microbiol. Biotechnol. 2014, 98, (8), 3819-3828. 60. Xu, X. Y.; Chen, X.; Su, P.; Fang, F.; Hu, B. B., Biodegradation potential of polycyclic aromatic hydrocarbons by bacteria strains enriched from Yangtze River sediments. Environ. Technol. 2016, 37, (5), 513-520. 61. Raj, A.; Reddy, M. M. K.; Chandra, R., Identification of low molecular weight aromatic compounds by gas chromatography-mass spectrometry (GC-MS) from kraft lignin degradation by three Bacillus sp. Int. Biodeterior. Biodegrad. 2007, 59, (4), 292-296. 62. Daane, L. L.; Harjono, I.; Barns, S. M.; Launen, L. A.; Palleroni, N. J.; Haggblom, M. M., PAH-degradation by Paenibacillus spp. and description of Paenibacillus naphthalenovorans sp nov., a naphthalene-degrading bacterium from the rhizosphere of salt marsh plants. Int. J. Syst. Evol. Microbiol. 2002, 52, 131-139. 63. Haichar, F. Z.; Marol, C.; Berge, O.; Rangelcastro, J. I.; Prosser, J. I.; Balesdent, J.; Heulin, T.; Achouak, W., Plant host habitat and root exudates shape soil bacterial community structure. ISME J 2008, 2, (12), 1221-1230. 30


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743

31



Page 32 of 36

744

Captions

745

Figure 1 The shift tendency of OTU_56, OTU_65, OTU_111 and OTU_106 fragments. The

746

relative abundances of the OTU_56, OTU_65, OTU_111 and OTU_106 fragments over a range of

747

buoyant density (BD) of DNA extracted from the

748

standard deviation (SD) relative abundance from triplicate microcosms are shown.

749

Figure 2 The shift tendency of OTU_8, OTU_15, OTU_43, OTU_83, OTU_23, OTU_106 and

750

OTU_111 fragments. The relative abundance of the OTU_8, OTU_15, OTU_43, OTU_83,

751

OTU_23, OTU_106 and OTU_111 fragments over a range of buoyant density (BD) of DNA

752

extracted from the 12C_ABA and

753

abundance from triplicate microcosms are shown.

754

Figure 3 Phylogenetic tree of identified OTUs responsible for PHE degradation.

755

Neighbour-joining tree based on 16S rRNA gene sequences showing the phylogenetic position of

756

the bacteria corresponding to OTU_8, OTU_15, OTU_23, OTU_43, OTU_83, OTU_56, OTU_65,

757

OTU_111, OTU_106 and their representatives of other related taxa. Bootstrap values (expressed

758

as percentages of 1000 replications) > 50% are shown at the branch points. Bar 0.05 substitutions

759

per nucleotide position.

760

Figure 4 Phylogenetic tree of amplified PAH-RHDα GP genes from the heavy fractions of

761

13

762

amplified from heavy fractions of 13C_AD treatment, and PAH-RHD L1, L6 and L10 represent the

763

PAH-RHDα genes from the 13C_ABA treatment.

C_AD and

13

13

12

C_AD and

13

C_AD treatments. Mean ±

C_ABA treatments. Mean ± standard deviation (SD) relative

C_ABA microcosms. PAH-RHD C1 and C15 represent the PAH-RHDα genes

32


Page 33 of 36


764

Figure 1

3

3

Relative Abundance (%)

Rhodoplanes (OTU_56-bacteria)

Xanthomonadaceae (OTU_65-bacteria)

12C 13C

2

2

1

1

0 1.68

1.70

1.72

1.74

1.76

0 1.68

1.78

1.5

1.70

1.72

1.74

1.76

1.78

1.74

1.76

1.78

8

Mycobacterium (OTU_106-bacteria)

6

Enterobacteriaceae (OTU_111-bacteria)

1.0

4

0.5 2

0.0 1.68

1.70

1.72

1.74

1.76

1.78

0 1.68

1.70

1.72

Buoyant Density ( g/mL)

765 766

33



767

Page 34 of 36

Figure 2 1.6

Relative Abundance (%)

1.2

1.6

Bacillus (OTU_8-bacteria)

0.8

0.8

0.4

0.4

0.0 1.70

1.72

1.74

1.76

1.78

3.0

2.4

Paenibacillus (OTU_43-bacteria)

1.2

0.0 1.68

1.70

1.72

1.74

12C 13C

1.76

1.78 0.5

0.30

Ammoniphilus (OTU_15-bacteria)

Sporosarcina (OTU_83-bacteria)

0.24

0.4

Mycobacterium (OTU_106-bacteria)

0.3 1.8

0.18

1.2

0.12

0.6

0.06

0.2

0.1

0.0 0.0 1.68

1.70

1.72

1.74

1.76

1.78

2.0

1.6

0.00 1.68

1.70

1.72

1.74

1.76

10

Hyphomicrobium (OTU_23-bacteria)

8

1.2

0.8

1.68

1.78

1.70

1.72

1.74

1.76

1.78

1.76

1.78

1.0

Enterobacteriaceae (OTU_111-bacteria)

0.8

6

0.6

4

0.4

2

0.2

Acinetobacter tandoii (OTU_4-bacteria)

0.4

0.0 1.68

768 769

1.70

1.72

1.74

1.76

1.78

0 1.68

1.70

1.72

1.74

1.76

1.78

0.0 1.68

Buoyant Density (g/ mL)

34


1.70

1.72

1.74

Page 35 of 36


770

Figure 3 Salmonella enterica subsp. houtenae DSM 9221 (U92195) Escherichia coli EP5 (KU904492.1) Escherichia fergusonii ATCC 35469 (CU928158) 72

OTU_111 bacteria Shigella dysenteriae ATCC 13313 (X96966)

62 100

Kluyvera georgiana ATCC 51603 (AF047186) Raoultella planticola ATCC 33531 (JMPP01000074)

88

75

Shigella sonnei GTC 781 (AB273732)

Gammaproteobacteria

Enterobacteriales

57

Xanthomonadales

Rosenbergiella nectarea 8N4 (HQ284827) Luteibacter anthropi CCUG 25036 (FM212561) 100

OTU_65 bacteria Frateuria aurantia DSM 6220 (AGRV01000006)

82 95

Dyella soli JS12-10 (EU604272) OTU_23 bacteria

Alphaproteobacteria

96 90

Hyphomicrobium denitrificans ATCC 51888 (ACVL01000012)

100

Hyphomicrobium chloromethanicum CM2 (AF198623) Hyphomicrobium facile ATCC 27489 (Y14311)

100

98 100

OTU_56 bacteria Uncultured Rhodoplanes sp. clone Leob163 (KF226093.1)

Rhodoplanes elegans AS130 (D25311) 63 72

Rhodoplanes piscinae JA266 (AM712913) Rhodoplanes roseus 941 (D25313)

Actinobacteria

Mycobacterium brisbanense ATCC 49938 (AY012577) Mycobacterium elephantis 484 (AJ010747) 100

Mycobacterium moriokaense DSM 44221 (AJ429044) Mycobacterium pulveris DSM 44222 (AJ429046)

OTU_15 bacteria 99 Ammoniphilus Resinae CC-RT-E (HM193518) 96 Ammoniphilus oxalaticus RAOx1 (Y14578) Ammoniphilus oxalivorans RAOx-FS (Y14580) OTU_43 bacteria 94

Paenibacillus xylanisolvens X11-1 (AB495094) Paenibacillus naphthalenovorans PR-N1 (AF353681)

86

Sporosarcina aquimarina SW28 (AF202056) 84 Sporosarcina saromensis HG645 (AB243859) OTU_83 bacteria Sporosarcina luteola Y1 (AB473560) Uncultured Bacillus sp. Jel2 8B (LT601541.1) Bacillus sp. IS 1a (KU244571.1) Bacillus sp. ADMK38 (KU850994.1) 95

OTU_8 bacteria Bacillus sp. MJAU D0051 (KT350461.1)

0.02

771

35


Bacillaceae

99

Planococcaceaee

Paenibacillus soli DCY03 (DQ309072) Sporosarcina ureae DSM 2281 (AF202057)

Firmicutes

59 100

Paenibacillaceae

96

Aneurinibacillus

OTU_106 bacteria


772

Figure 4

Mycobacterium vanbaalenii PYR-1 PAH-RHD gene (AAY85176.1) PAH-RHDα L6 Mycobacterium sp. py143 NidA3 gene (ADH94645.1) Mycobacterium novocastrense PAH-RHD gene (GAT12202.1) Terrabacter sp. FLO PAH-RHD gene (ABA87073.1) Marine bacterium UBF-Py PAH-RHD gene (AFT63035.1) Mycobacterium sp. AP1 PAH-RHD gene (AFT63034.1) Uncultured bacterium PAH-RHD gene (AEW70580.1) Uncultured bacterium PAH-RHD gene (AMM73050.1) Uncultured bacterium NidA3 gene (AIC78910.1) Uncultured bacterium PAH-RHD gene (AMM73086.1) 99

Uncultured bacterium PAH-RHD gene (AEW70585.1) Uncultured bacterium PAH-RHD gene (AMM73080.1) Uncultured bacterium PAH-RHD gene (AEW70614.1) 63 PAH-RHDα C15 Uncultured bacterium PAH-RHD gene (AEW70574.1) Uncultured bacterium PAH-RHD gene (AEW70635.1)

97 PAH-RHDα L1 PAH-RHDα C1

76 87

Mycobacterium sp. S23 PAH-RHD gene (ALS30442.1) Uncultured bacterium RHD gene (AGO66328.1)

57

96 Mycobacterium sp. CH-1 PAH-RHD gene (ABD97978.1) 76 Mycobacterium sp. SNP11 PAH-RHD gene (ABK27720.1) Uncultured Comamonas sp. PAH-RHD gene (CAN85239.1) 99

Aeromonas sp. XF3-2 PAH-RHD gene (AND66083.1)

PAH-RHDα L10 70 Bacillus sp. CL1-1 PAH-RHD gene (AND66078.1) 61 Enterobacter sp. CL1-2 PAH-RHD gene (AND66079.1) Uncultured Comamonas sp. PAH-RHD gene (CAN85224.1) Uncultured bacterium NagA gene (AHZ97886.1)

773

0.1

36


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Autochthonous Bioaugmentation-Modified ... - ACS Publications

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