Diversity of the active phenanthrene degraders in PAH-polluted soil is shaped by ryegrass rhizosphere and root exudates

Accepted Manuscript Diversity of the active phenanthrene degraders in PAH-polluted soil is shaped by ryegrass rhizosphere and root exudates Jibing Li, Chunling Luo, Dayi Zhang, Xixi Cai, Longfei Jiang, Xuan Zhao, Gan Zhang PII:

S0038-0717(18)30356-0

DOI:

10.1016/j.soilbio.2018.10.008

Reference:

SBB 7312

To appear in:

Soil Biology and Biochemistry

Received Date: 8 March 2018 Revised Date:

17 September 2018

Accepted Date: 16 October 2018

Please cite this article as: Li, J., Luo, C., Zhang, D., Cai, X., Jiang, L., Zhao, X., Zhang, G., Diversity of the active phenanthrene degraders in PAH-polluted soil is shaped by ryegrass rhizosphere and root exudates, Soil Biology and Biochemistry (2018), doi: https://doi.org/10.1016/j.soilbio.2018.10.008. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

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Diversity of the active phenanthrene degraders in PAH-polluted soil is shaped by

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ryegrass rhizosphere and root exudates

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Jibing Lia,e, Chunling Luoa,b*, Dayi Zhangc, Xixi Caid, Longfei Jianga, Xuan Zhaoa, Gan Zhanga

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a

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b

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Guangzhou 510642, China

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c

School of Environment, Tsinghua University, Beijing 100084, China

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d

College of Resources and Environment, Fujian Agriculture and Forestry University, Fuzhou

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350002, China

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Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China

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College of Natural Resources and Environment, South China Agricultural University,

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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Abstract

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Root exudates can stimulate microbial degradation within the rhizosphere, but their exact roles are

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embedded within the complicated rhizospheric effects. In the present study, we applied both 12C-

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and

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phenanthrene degradation via DNA stable isotope probing (DNA-SIP). A significant increase of

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phenanthrene biodegradation efficiency (10.7%) was found in ryegrass rhizosphere compared to

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bulk soils, but not in soils supplemented with ryegrass root exudates. Results from

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high-throughput sequencing and computational analyses suggested that treatments with both

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ryegrass rhizosphere and root exudates markedly increased total bacterial populations and shaped

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the

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phenanthrene-degraders belonging to eight bacterial classes revealed by DNA-SIP, only

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Alphaproteobacteria and Nitrososphaeria were shared between bulk soils, ryegrass rhizosphere

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and soils supplemented with ryegrass root exudates. Sphingobacteriia and Actinobacteria were

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active phenanthrene-degraders within both ryegrass rhizosphere and soils supplemented with

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ryegrass root exudates, whereas others were observed only in bulk soils or soils supplemented

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with ryegrass root exudates. Most of the degraders were linked to phenanthrene degradation for

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the first time based on their incorporation of 13C-phenanthrene. In 13C-phenanthrene microcosms,

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the relative abundance of PAH-RHDα genes and active phenanthrene-degraders was strongly

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correlated with phenanthrene degradation efficiency. Compared to the rhizosphere, root exudates

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provided a minor contribution to the abundance of PAH-RHDα gene. This study helps in better

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understanding the roles of root exudates supplement in the phenanthrene biodegradation process

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within the rhizosphere and provides theoretical insights into the mechanisms of enhanced

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phenanthrene degradation via phytoremediation at PAH-contaminated sites.

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Keywords: Root exudates; rhizosphere; active phenanthrene-degraders; PAH-RHDα genes

of

the

active

phenanthrene-degrader

community.

Of

all

the

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C-phenanthrene to distinguish the effects of root exudates within ryegrass rhizosphere on

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1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are molecules containing two or more fused

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benzene rings (Haritash and Kaushik, 2009). They are ubiquitous environmental pollutants, with

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toxic, mutagenic, and carcinogenic properties, and have caused substantial environmental and

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human health concerns (Keith and Telliard, 1979; Čvančarová et al., 2013). PAHs released into the

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environment can be removed through physical, chemical, and biological approaches (Khan et al.,

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2009; Yu et al., 2011). Despite the differences among these technologies, they all suffer from low

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efficiency, operational complexity, or high cost. Phytoremediation, defined as the use of plants and

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associated microorganisms in the rhizosphere for in situ treatment of environmental pollutants, is

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recognised as an environmentally friendly, cost-effective, and socially acceptable approach for the

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remediation of PAH-contaminated soil (Ní Chadhain et al., 2006; Shahsavari et al., 2015).

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The rhizosphere has an essential role in PAH phytoremediation. Numerous studies have investigated

the involvement of rhizosphere-associated

microorganisms in

soil PAH

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bioremediation (Khan et al., 2013b; Liu et al., 2015). For example, the plant rhizosphere can

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significantly improve the dissipation of PAHs compared to unplanted soil (Cheema et al., 2010; Yu

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et al., 2011), and accelerated PAH removal is attributable mainly to the enhancement of bacterial

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activity and diversity in the rhizosphere due to improved soil aeration, permeability, and break-up

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of soil aggregates, leading to an increase in PAH bioavailability (Hamdi et al., 2007). In addition,

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compounds released by roots (i.e. root exudates) can represent high carbon input into the

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rhizosphere (And and Leyval, 2003; Cebron et al., 2011), and some exudates act as surfactants,

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increasing PAH solubility (Cebron et al., 2011). Therefore, root exudates can stimulate microbial

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degradation processes within the rhizosphere. In bulk soils supplemented with root exudates, the

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presence of root exudates can modify the bacterial diversity of PAH degraders and increase the

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abundance of bacteria containing known PAH ring hydroxylating dioxygenase (PAH-RHDα) genes

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in PAH-polluted soil (Cebron et al., 2011). Accordingly, root exudates are speculated to be a

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predominant factor leading to changes in microbial communities in the rhizosphere and a potential

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driver behind enhanced petroleum biodegradation (Martin et al., 2014). However, the actual role

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of root exudates within the rhizosphere, compared to the physical effects of the rhizosphere, in

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promoting PAH degradation and changes in the microbial community structure throughout the

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phytoremediation process remains unclear.

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ACCEPTED MANUSCRIPT Complex rhizospheric microbial communities present challenges in identifying active PAH

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degraders, which are predominantly yet-to-be-cultivated species (Rappé and Giovannoni, 2003).

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Stable-isotope probing (SIP) is a cultivation-independent technique that circumvents the

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requirement to isolate an organism to assess its metabolic responses, and provides the opportunity

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to link microbial identities to their functions (Dumont and Murrell, 2005). Using substrates

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labeled with stable isotopes (e.g., 13C or 15N) results in the formation of isotope-enriched cellular

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components, such as DNA, RNA, and protein, of the microorganisms involved in the

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mineralisation of the substrate (Jiang et al., 2015). SIP has been applied to identify many

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indigenous bacteria capable of degrading PAHs (Song et al., 2015; Li et al., 2018), and has been

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used to identify bacterial communities actively assimilating root exudates in the rhizosphere

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(Haichar et al., 2008). However, previous studies only addressed the change in PAH-degrading

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bacteria in rhizosphere or soils supplemented with root exudates alone, and no work has attempted

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to compare the diversity and functions of PAH-degraders between bulk soils, rhizosphere and soils

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supplemented with root exudates during PAH degradation process in PAH-contaminated soil via

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SIP.

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Phenanthrene (PHE) is used as a model PAH compound due to its ubiquity in nature and

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angular fused-ring structure (Li et al., 2017a). Therefore, we selected it as the target compound in

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this study to investigate the change in phenanthrene degradation efficiency, bacterial populations,

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active phenanthrene degraders, and phenanthrene-degrading genes between bulk soils, ryegrass

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rhizosphere and soils with ryegrass root exudates. To achieve these objectives, we performed

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DNA-SIP experiments to target the active degraders incorporating

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soils supplemented with ryegrass root exudates. After 12 days of phenanthrene degradation,

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high-throughput sequencing and quantitative polymerase chain reaction (qPCR) revealed

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significant changes in the active phenanthrene degraders and abundance of functional PAH-RHDα

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genes across the treatments. Our findings suggested a minor contribution of root exudates to in

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situ phenanthrene degradation compared to the rhizosphere. This information can be used to

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C-phenanthrene by adding

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C-labelled phenanthrene to bulk soils, ryegrass rhizosphere (soils planted with ryegrass), and

improve phytoremediation of PAH-contaminated sites.

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2. Materials and methods

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2.1. Sample collection Soil samples were collected in the Shengli Oil Field (37°68′N, 118°48′E) in Dongying City,

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Shandong Province, China, in September 2015, from a depth of 0–20 cm. After transport to the

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laboratory at 4°C, the soil was air dried at room temperature for 7 days and sieved to 2 mm for

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homogenisation. A portion of the soil was stored at -80°C for subsequent DNA extraction, and the

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rest was used immediately for the phenanthrene degradation and SIP experiments. Table 1 lists the

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soil characteristics and PAH concentrations. Total carbon (TC) and total nitrogen (TN) were

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measured on an element analyzer (VarioEL III, Elementar, Hanau, Germany) (Hedges and Stern,

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1984). Individual PAHs in the soil were quantified using gas chromatography–mass spectrometry

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(see Section 2.5).

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2.2. Ryegrass and root exudate collection

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Ryegrass (Lolium perenne) was used in this study for its good performance in accelerating PAH degradation in soil based on field studies (Sun et al., 2010a). After sterilisation with 30% (v/v)

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H2O2 for 30 min, ryegrass seeds were germinated in a culture dish at room temperature in the dark.

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After germination, the seedlings were transferred to Hoagland solution for further growth until the

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seedlings matured in an artificially controlled climate chamber (7-9 weeks) with a day/night

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photoperiod of 14/8 h (10,000 Lux), temperature of 28/22°C, and relative humidity of 60% before

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the phenanthrene degradation and SIP experiments. Root exudates were collected from ryegrass

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samples on week 8 in advance by submerging the roots four times in demineralised sterile water

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for 6 h. The root exudate extracts were stored at 4°C and finally gathered, filtered through a

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0.2-mm filter (Nalgene), lyophilised, and pooled (Cebron et al., 2011). The total organic carbon

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and inorganic carbon were analyzed by a Total Organic Carbon analyzer (TOC-VCPH, Shimadzu);

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total nitrogen is defined as the sum of total Kjeldahl nitrogen and nitrogenous anions, which was

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detected as previously described (Bundy et al., 2017). Calcium, potassium, magnesium, sodium

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and silicon were measured by flame atomic absorption spectrometry using a GBC Avanta

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instrument, and iron was measured by graphite furnace atomic absorption spectrometry using a

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deuterium lamp for non-atomic correction as previously described by (Mora et al., 2017). The

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nitrogen (12.7 mg/g), calcium (12.04 mg/g), iron (0.84 mg/g), potassium (102.6 mg/g),

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magnesium (13.6 mg/g), sodium (11.8 mg/g) and silicon (0.66 mg/g), similar with previous

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studies (Cebron et al., 2011).

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2.3. SIP microcosms

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Microcosms were set up in miniature planting pots with dimensions of 20 × 70 mm (diameter

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× height). After the soil was homogenised, demineralised sterile water was added to the pots to

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adjust the water content to 60% (vol./wt.) of the soil water holding capacity (WHC) before use.

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Subsequently, unlabelled phenanthrene (99%) or

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Cambridge Isotope Laboratories, Inc., Tewksbury, MA, USA) was added to the pots at a final

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phenanthrene concentration of 10 mg/kg. Treated soil was packed into the pots (5 g dry weight

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soil per pot). For the treatments planted with ryegrass (rhizosphere treatments), two uniform

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mature ryegrasses were transplanted carefully and grown in each pot. For the soil inoculated with

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root exudates, 100 mg of exudate was added to each pot. All the treated soils were watered with

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sterile distilled water throughout the experiment to keep the soil at approximately 60% of its WHC.

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In

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(12C-phenanthrene with ryegrass root exudate), 12C_RG (12C-phenanthrene with growing ryegrass),

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13

C_AD (13C-phenanthrene only),

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C_RG (13C-phenanthrene with growing ryegrass). A sterile control treatment was prepared with

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unlabelled phenanthrene in soils sterilized by a gamma-ray technique, which attempted to confirm

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the occurrence of PHE biodegradation and evaluate its contribution to PHE removal soils. All

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microcosms were incubated using the same seedling planting method described in Section 2.2.

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Each treatment was carried out in nine replicates. For the RG treatments, after digging out the

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ryegrass plant and gently shaking, the rhizospheric soils from the growing ryegrass were sampled

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by brushing the roots according to the previous study (Deng et al., 2018). On days 4, 8, and 12

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after incubation, three soil replicates from each treatment were collected for PAH analysis and

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DNA extraction.

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2.4. Nucleic acid extraction and ultracentrifugation

C-labelled phenanthrene (13C14-PHE, 99%;

six

treatments

were

included:

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12

C_AD

(12C-phenanthrene

alone),

12

C_RE

C_RE (13C-phenanthrene with ryegrass root exudate), and

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Total nucleic acids were extracted from 2 g of each collected soil using the PowerSoil DNA

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Isolation Kit (MO BIO, Carlsbad, CA, USA) according to the manufacturer’s instructions, and the

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DNA content was quantified using a ND-2000 UV-vis spectrophotometer (NanoDrop

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Technologies, Wilmington, DE, USA) (Li et al., 2017a). To separate 12C- and

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acid ultracentrifugation was conducted as described previously (Song et al., 2016). Approximately

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5 µg of DNA was added to Quick-Seal polyallomer tubes (13 × 51 mm, 5.1 mL, Beckman Coulter,

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Pasadena, CA, USA) and mixed with Tris-EDTA (pH 8.0)/CsCl solution at a final buoyant density

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(BD) of ~1.77 g/mL. After balancing and sealing, density gradient centrifugation was performed

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in an ultracentrifuge (Optima L-100XP; Beckman Coulter) at 175,000 ×g for 48 h (20°C). The

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centrifuged gradients were fractionated into different fractions of 400 µL. The DNA fractions were

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purified to remove Tris-EDTA and CsCl using the method described by Sun et al. (2010b) after

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measuring the BD of each fraction. Compared to the

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concentrations from all six biotic treatments at higher BD (1.7372–1.7784 g/mL) were higher in

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the 13C-phenanthrene microcosms, indicating that a part of DNA was labelled with the assimilated

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2.5. High-throughput sequencing and computational analyses

C-DNA, nucleic

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C-phenanthrene microcosms, the DNA

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The hypervariable V4 region of bacterial 16S rRNA gene fragments was amplified for each

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treatment using the 515f/806r primer set (Table S1), as described by Bates et al. (2011). In total,

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558 samples were amplified and sequenced. Unique heptad-nucleotide sequences (12 bases) were

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

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performed using Li’s method (Li et al., 2017a). Sequencing was conducted on an Illumina MiSeq

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sequencer in a standard pipeline using 2 × 250 bp PE technology. The qualified sequences were

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analysed as described by Schloss et al. (2009) and Caporaso et al. (2010) and then assigned using

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an operational taxonomic unit (OTU)-based method to generate microbiome profiles (Edgar, 2010;

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Mcdonald et al., 2012; Werner et al., 2012).

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The relative abundance of each OTU was determined and the top 100 relative abundances

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were selected for analysis according to previous studies (Sun et al., 2010b; Li et al., 2017a). The

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phenanthrene degraders were identified from OTUs enriched in the heavy fractions from the

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C_AD,

13

C_RE, and

13

C_RG microcosms compared to the 7

12

C_AD,

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C_RE, and

12

C_RG

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samples, respectively. In the present study, OTUs identified as active phenanthrene degraders from

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the

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phylogenetic analysis of these sequences was performed as described previously (Li et al., 2018).

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The GenBank accession numbers of the above sequences are provided in the Supporting

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Information.

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2.6. Detection of PAH-RHDα genes

C_phenanthrene treatments were further trimmed using the Greengenes database, and

The PAH-RHDα genes in the heavy DNA fractions from the

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C-phenanthrene treatments

were amplified using two primer sets for gram-positive (GP, 642f/933r) and gram-negative

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(610f/911r) degraders (Table S1) (Cebron et al., 2008). Gradient PCR and amplification reactions

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were performed as described previously (Li et al., 2017a). In the present study, only one strong

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and specific amplicon was produced with the PAH-RHDα GP primer set and selected for

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subsequent analysis. The PCR products were gel-purified using a gel extraction kit (D2500-01;

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Omega Bio-tek, Norcross, GA, USA), followed by cloning and sequencing as described

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previously (Jiang et al., 2015). Briefly, the purified fragments were ligated into pMD-19T

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(TaKaRa) and transformed into Escherichia coli DH5α. The positive clones containing right

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inserts were selected on ampicillin-containing (50 mg/L) Luria-Bertani agar plates for 12 h at

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37°C. The plasmids were finally extracted and sent for sequence. Phylogenetic dendrograms were

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prepared using the method described in section 2.5. The GenBank accession number for the partial

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PAH-RHDα gene sequence is MG659713.

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2.7. Quantitative polymerase chain reaction (qPCR)

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The abundance of bacterial 16S rRNA and PAH-RHDα GP genes in each fraction from

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pair Bac519F/Bac907R and the PAH-RHDα GP primer pair 642f/933r (Table S1). The PCR

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reactions were performed in a 20-µL mixture containing 10 µL of SYBR green PCR Premix Ex

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Taq II (TaKaRa, Japan), 0.5 µL of each primer (10 µM; BGI-Shenzhen), and 1 µL of DNA

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template ( 10 ng/uL) on an ABI 7500 real-time PCR system (Applied Biosciences, USA). Two

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standard curves were obtained by producing a 10-fold serial dilution of plasmid pGEM-T Easy

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Vector sequences (102–108 copies; Promega) containing the 16S rRNA and PAH-RHDα GP genes,

C-labelled and unlabelled DNA were determined by qPCR using the bacterial universal primer

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(Song et al., 2015; Li et al., 2017b): initial denaturation at 94°C for 10 min, followed by 40 cycles

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at 94°C for 30 s, 55°C for 30 s, and 72°C for 15 s. Finally, melt curves were obtained from 60 to

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95°C at an increment of 0.2°C/cycle.

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2.8. Chemical analysis

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Phenanthrene in soil and plant tissue from each microcosm (days 0, 4, 8, and 12 after

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incubation) was analysed by gas chromatography (model 7890; Agilent, Santa Clara, CA, USA),

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

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(model 5975; Agilent), as described previously (Khan et al., 2009; Jiang et al., 2015). Briefly, after

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sieving soil (0.2 mm) and grinding ryegrass tissue, samples were spiked with 1,000 ng of

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deuterated PAHs and extracted twice with dichloromethane. The extracted organic phase was

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concentrated to approximately 0.5 mL and then 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 was added as an internal standard to all samples before the instrumental

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analysis. Table S2 lists the components and concentrations of the deuterated PAHs, standards, and

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internal standard.

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2.9. Statistical analysis

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Data were expressed as mean ± standard deviation (SD). Statistical analyses were performed

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using Origin 8.0 (OriginLab Corporation, MA). The least significance difference (LSD) test was

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used to determine differences at α=0.05 level. Pearson correlation coefficients between the

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abundance of total 16S rRNA genes, active phenanthrene-degrading bacteria, PAH-RHDα genes,

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and phenanthrene degradation efficiency were calculated using SPSS statistical package 17.0

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(SPSS Inc., Chicago, IL).

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3. Results

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3.1. Phenanthrene degradation performance Figure 1 presents the soil residual phenanthrene concentrations in the

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12

C_RE,

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phenanthrene was detected in ryegrass tissues over the 12-day period (Table S3). The

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phenanthrene concentration in the sterile control decreased less than that in the biotic treatments,

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confirming the major contribution of biodegradation to phenanthrene elimination, which was in

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agreement with many previous studies (Li et al., 2017a; Jiang et al., 2015; Li et al., 2018). No

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significant difference (p > 0.05) was observed between

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degradation in the AD, RE, or RG treatments. As shown in Figure 1, the residual phenanthrene

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levels in the 12C_AD (53.9 ± 4.9%) and 13C_AD (53.4 ± 3.7%) microcosms were lower than those

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in the

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Thereafter, the phenanthrene biodegradation rates slowed and became similar in the AD and RE

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microcosms (56.3–56.4% and 65.7–65.9% on days 8 and 12, respectively). However, the residual

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phenanthrene levels in the

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much lower than those in the AD and RE treatments after 4 days. Moreover, residual phenanthrene

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was significantly lower (p < 0.05) in the RG microcosms (34.0 ± 3.4% and 23.3 ± 2.5% on days 8

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and 12, respectively) than in the AD microcosms (43.6 ± 4.1% and 34.4 ± 3.4% on days 8 and 12,

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respectively), suggesting that the degradation efficiency increased by 9.64% and 11.07%,

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respectively, in the presence of ryegrass. Statistical analysis showed a significant difference in

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phenanthrene degradation between the AD (bulk soils) and RG (rhizosphere) treatments, as well as

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RE (soils supplemented with root exudates alone) and RG microcosms, suggesting that the

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supplement of root exudates in PAH-contaminated soils offered only a minor contribution to the

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phenanthrene degradation efficiency.

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3.2. Abundance and community structure of total soil bacteria

C_RE, and

13

C_RG microcosms at different sampling times. Minimal

12

13

13

C-phenanthrene

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C-phenanthrene and

C_RE (59.0 ± 4.6%) treatments during the first 4 days.

C_RG (49.2 ± 4.7%) and

13

C_RG (49.2 ± 3.5%) microcosms were

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C_AD,

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C_RG,

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12

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DNA was extracted from the six biotic treatments (non-sterilized soils) after 4, 8 and 12 days

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of incubation, followed by qPCR and high-throughput sequencing. The total 16S rRNA gene

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abundance in the RG microcosms (5.76 × 108 copies/g soil) and RE microcosms (5.34 × 108 10

ACCEPTED MANUSCRIPT copies/g soil) was approximately 17-fold higher than that in the AD microcosms (3.27 × 107

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copies/g soil) after 12 days of incubation (Figure S1), indicating that both root exudates and the

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rhizosphere had significant roles in increasing the total soil bacterial population. In the AD, RE,

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and RG treatments, the relative abundance of total 16S rRNA genes defined by genus showed

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slight differences from the indigenous microbial community structures in the degradation of

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12

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composition and structure of the microbial communities exhibited significantly different

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behaviours among the AD, RE, and RG microcosms. Unclassified members of the genera

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Micrococcaceae and Comamonadaceae were predominant (>5%) in the AD, RE, and RG

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microcosms, but their abundance changed significantly with biodegradation. There was a

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concomitant decrease in the relative abundance of unclassified Micrococcaceae in the AD

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treatments, from 15.4% on day 4 to 12.9% on day 8 and 10.2% on day 12. Their relative

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abundance (19.3%) was significantly higher in the RE microcosms than in the AD microcosms in

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the first 4 days, but decreased to 12.3% and 9.14% after 8 and 12 days, respectively. In the RG

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microcosms, the relative abundance of unclassified Micrococcaceae remained lower than those in

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the AD and RE microcosms on days 4, 8, and 12 (6.09%, 6.44%, and 6.45%, respectively). There

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was a concomitant decrease in the relative abundance of Comamonadaceae in the RG treatments,

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from 12.2% on day 4 to 9.99% on day 8 and 8.98% on day 12, all significantly higher than those

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in the AD (7.06%, 5.76%, and 5.89%, respectively) and RE (6.31%, 5.67%, and 5.94%,

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respectively) microcosms. For rare bacteria, the relative abundances of unclassified Streptophyta

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(phylum Cyanobacteria, class Chloroplast) in the RG microcosms (21.8%, 9.75%, and 8.48%)

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were significantly higher than those in the AD (0.58%, 0.63%, and 1.06%) and RE (0.62%, 0.68%,

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and 0.88%) microcosms on days 4, 8, and 12, respectively.

290

3.3. Active phenanthrene degraders in the AD microcosms as revealed by DNA-SIP

13

C-phenanthrene over the 12-day period (Figure S2). However, the

AC C

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C-phenanthrene and

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267

291

Bacterial 16S rRNA gene abundance was quantified using qPCR on DNA recovered from

292

each fraction of all microcosms. On days 4, 8, and 12, the abundances of bacterial 16S rRNA

293

genes in fractions with higher BDs (1.7372–1.7784 g/mL; see the stars in Figure S4) in the

294

13

295

in grey, Figure S3). The indigenous microorganisms responsible for 13C-phenanthrene assimilation

C-AD microcosm were significantly higher than those in the 12C-phenanthrene control (marked

11

ACCEPTED MANUSCRIPT 12

296

were detected by comparing the relative abundances of specific OTUs in the

297

and

298

higher BDs (1.7372–1.7784 g/mL) and enriched in the

299

and S4). In contrast, no enrichment or similar trends were detected in the

300

addition, there was a concomitant increase in the relative abundance of OTU_3175 in the heavy

301

fractions of the

302

OTU_42987) were only enriched on days 8 and 12, but not day 4, indicating that the

303

microorganisms represented by these OTUs derived

304

incubation. Finally, OTU_32213 was enriched after 12 days of incubation. In total, five types of

305

bacteria represented by the above OTUs were detected in the AD microcosms over the 12-day

306

incubation period.

307

3.4. Active phenanthrene degraders in the RE and RG microcosms as revealed by DNA-SIP

C-phenanthrene treatments in each fraction. OTU_1476 and OTU_3175 were found at C_AD microcosm on day 4 (Figures 2 C_AD treatment. In

RI PT

12

C_AD treatment from days 4 to 12. Moreover, two OTUs (OTU_34691 and

13

C from

13

C-phenanthrene after 8 days of

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13

13

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C-phenanthrene

The bacterial 16S rRNA gene abundance in the RE and RD microcosms showed similar

309

patterns of enrichment in fractions with higher BDs (1.7575–1.7773 g/mL, see stars in Figures S5

310

and S6), as illustrated in Figure S3. Soil with root exudates produced a significant change in the

311

diversity of the indigenous phenanthrene-degrading communities (Figures 2 and S5). OTU_23284,

312

OTU_28345, and OTU_42503 were enriched in the heavy fractions of the 13C_RE microcosm on

313

day 4. On day 8, another two OTUs (OTU_3175 and OTU_34628) were enriched, and their

314

relative abundances in the heavy DNA fractions of

315

than those in the 12C_RE treatment. The other three enriched OTUs (OTU_753, OTU_34691, and

316

OTU_41564) were detected after 12 days of incubation. In total, eight OTUs were enriched in the

317

heavy fractions of the 13C_RE treatment over the 12-day period.

EP

13

C_RE treatment were significantly higher

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318

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308

The ryegrass rhizosphere also significantly changed the diversity of the active

319

phenanthrene-degrading communities (Figures 2 and S6). Three types of bacteria, represented by

320

OTU_3175,

321

phenanthrene-degrading communities over the 12-day period, as their relative abundances in the

322

heavy fractions of the

323

treatment. After 8 days of incubation, one new enriched OTU (OTU_12876) was detected. On day

324

12, more OTUs (OTU_14966, OTU_19134, OTU_34306, OTU_42987, and OTU_53992) were

OTU_5288,

13

and

OTU_15584,

were

identified

as

the

indigenous

C_RG treatment were significantly higher than those in the

12

12

C_RG

ACCEPTED MANUSCRIPT 325

enriched in the heavy fractions of the 13C_RG treatment, showing involvement of representative

326

microorganisms in phenanthrene biodegradation. Although the phenanthrene-degrading communities differed among the AD, RE, and RG

328

treatments, some active phenanthrene degraders were shared between microcosms. The

329

microorganisms represented by OTU_3175 had a role in phenanthrene metabolism in all

330

treatments. Those represented by OTU_34691 were involved in phenanthrene metabolism in both

331

the AD and RE microcosms. The microorganisms represented by OTU_42987 were identified as

332

active phenanthrene degraders in both the AD and RG treatments. Calculation of the total

333

abundance of all identified PAH-degrading bacteria in each microcosm revealed that the ryegrass

334

rhizosphere could significantly improve the abundance of PAH-degrading bacteria (5.22 ± 0.33%

335

in the heavy fraction) after 12 days of incubation compared to the AD (4.69 ± 0.25%) and RE

336

(4.71 ± 0.31%) microcosms (Figure S7, p<0.05).

337

3.5. Significant community structure alteration of soil active phenanthrene degraders among the

338

AD, RE and RG treatments

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Figure 3 presents the phylogenetic information of the phenanthrene degraders represented by

340

the identified OTUs responsible for phenanthrene metabolism. The active phenanthrene-degraders

341

belonging to eight bacterial classes (Figure 3) included Alphaproteobacteria (Sphingoaurantiacus,

342

Kaistobacter, Sphingomonas, Novosphingobium, and Blastomonas), Gammaproteobacteria

343

(Pseudomonas), Betaproteobacteria (Herbaspirillum and Ramlibacter), Sphingobacteriia

344

(Chitinophagaceae, Pedobacter, and Mucilaginibacter), Blastocatellia (Pyrinomonadaceae),

345

Spartobacteria (Chthoniobacteraceae), Actinobacteria (Mycobacterium and Micrococcaceae),

346

and Nitrososphaeria (Nitrososphaera). OTUs belonging to the classes Alphaproteobacteria and

347

Nitrososphaeria were shared by all AD, RE, and RG treatments. OTU_12876, OTU_14966,

348

OTU_41564, OTU_15584, and OTU_3175 belonged to the genera Sphingoaurantiacus,

349

Kaistobacter, Sphingomonas, Novosphingobium, and Blastomonas, respectively, within the family

350

Sphingomonadaceae (phylum Proteobacteria,

351

represented by OTU_42987, OTU_34306, and OTU_34691 were characterised as the genus

352

Nitrososphaera (kingdom Archaea, phylum Thaumarchaeota, class Nitrososphaeria), and shared

353

high similarity with many strains in this genus, such as uncultured Nitrososphaera spp. clone

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13

class Alphaproteobacteria).

The bacteria

ACCEPTED MANUSCRIPT ncaOTU35 (KY225301.1), and uncultured Archaeon clone B118 (KX061174.1). Meanwhile,

355

some bacterial classes were observed in only two treatments, including Sphingobacteriia (AD and

356

RG microcosms) and Actinobacteria (RG and RE microcosms). Microorganisms represented by

357

OTU_53992, OTU_1476, and OTU_19134 belonged to the genera Pedobacter (family

358

Sphingobacteriaceae), Mucilaginibacter (family Sphingobacteriaceae), and an unclassified

359

Chitinophagaceae, respectively, within the order Sphingobacteriales (phylum Bacteroidetes, class

360

Sphingobacteriia). OTU_5288 and OTU_42503 were classified in the family Micrococcaceae

361

(order Micrococcales) and the genus Mycobacterium (family Mycobacteriaceae, order

362

Corynebacteriales), respectively, within the class Actinobacteria. Bacteria represented by

363

OTU_28345 and OTU_32213 only appeared in the RE and AD treatments, respectively, and were

364

assigned to the families Pyrinomonadaceae (phylum Acidobacteria, class Blastocatellia, order

365

Blastocatellales) and Chthoniobacteraceae (phylum Verrucomicrobia, class Spartobacteria, order

366

Chthoniobacterales).

367

Betaproteobacteria, Gammaproteobacteria, Spartobacteria, and Blastocatellia were only

368

observed in one treatment. OTU_23284 was found only in the RE treatment and was assigned to

369

the genus Pseudomonas (class Proproteobacteria, family Pseudomonadaceae) and exhibited

370

100% similarity to the partial 16S rRNA gene sequence of strain Pseudomonas frederiksbergensis

371

(MF139035.1). OTU_34628 and OTU_753 were detected only in the RE treatment and were

372

characterised as the genera Herbaspirillum (family Oxalobacteraceae) and Ramlibacter (family

373

Comamonadaceae), respectively, within the order Burkholderiales (phylum Proteobacteria, class

374

Betaproteobacteria). The phylogenetic information suggested significant alterations in the

375

community structure of the soil active phenanthrene degraders among the AD, RE, and RG

376

treatments.

377

3.6 Occurrence and quantification of PAH-RHDα genes involved in phenanthrene metabolism

active

phenanthrene

degraders

belonging

to

the

classes

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The PAH-RHDα GP genes were analysed in the heavy fractions of the 13C_AD, 13C_RE, and

378 379

13

380

98% similarity with an uncultured strain (AEW70574.1) and 96% similarity with the genus

381

Mycobacterium (AAY85176.1), which can degrade PAHs (Figure S8).

382

C_RG treatments. In all treatments, the PAH-RHDɑ gene sequences were identical and shared

The PAH-RHDɑ genes in the AD, RE, and RG treatments were quantified against each 14

ACCEPTED MANUSCRIPT 383

density-resolved fraction (Figure 4). Remarkable enrichment of PAH-RHDɑ genes in the heavy

384

DNA fractions (highlighted in grey, Figure 4) was observed in the 13C-AD, 13C-RE, and 13C-RG

385

treatments, indicating that the PAH-RHDɑ genes were labelled with the assimilated

386

contrast, no such enrichment or similar trend was detected in the

387

treatments. Accordingly, the PAH-RHDɑ genes detected in the heavy fractions of the 13C-labelled

388

microcosms were associated with phenanthrene metabolism. In addition, the copy numbers of the

389

PAH-RHDɑ genes in the heavy fractions of the 13C-AD, 13C-RE, and 13C-RG treatments increased

390

over the 12-day period, reaching 3.34 × 104, 3.45 × 104, and 4.92 × 104 copies/ng DNA on day 12

391

(Figure S9), respectively, approximately 8–10 times higher than those on day 4. Moreover, the

392

abundance of PAH-RHDα genes in the heavy fractions of the 13C-RG treatment was significantly

393

higher than that in the

394

increased the abundance of PAH-RHDα genes in active phenanthrene degraders participating in

395

PAH biodegradation.

396

3.7 Correlation between phenanthrene degradation efficiency and microbial abundance

C-AD,

12

C RE, and

12

C. In

C RG

13

13

C-RE treatments, indicating that the ryegrass rhizosphere

M AN U

C-AD and

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12

13

We evaluated the Pearson correlation coefficients between the copy numbers of the total 16S

398

rRNA gene, abundance of active phenanthrene-degrading bacteria, abundance of PAH-RHDα

399

genes, and phenanthrene degradation efficiency after 12 days of incubation. The phenanthrene

400

degradation efficiency was positively correlated with the abundance of active phenanthrene

401

degraders (Pearson correlation coefficient = 1.000, p < 0.01) and PAH-RHDα genes (Pearson

402

correlation coefficient = 0.999, p < 0.05), but not the total bacterial 16S rRNA genes (Table 2).

403

This linear correlation confirmed the incorporation of 13C-phenanthrene into the identified active

404

phenanthrene degraders and PAH-RHDα genes during phenanthrene metabolism. The relatively

405

lower abundance of PAH-RHDα genes and phenanthrene degraders in the RE microcosms

406

suggested a minor contribution of root exudates to the community diversity of phenanthrene

407

degraders compared to the rhizosphere, consistent with the phenanthrene degradation performance

408

results.

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4. Discussion Phytoremediation has been applied successfully for the remediation of PAH-contaminated

411

sites (Johnson et al., 2005; Guo et al., 2017). In the present study, soils planted with ryegrass

412

produced a significant increase in the phenanthrene degradation efficiency, hinting at the

413

contribution of phytoremediation to phenanthrene removal. However, no significant difference

414

was observed between the phenanthrene degradation rates with and without root exudates,

415

consistent with a previous study (Cebron et al., 2011). Therefore, the improvement in

416

phenanthrene biodegradation in rhizosphere was not explained by root exudates, but was

417

attributed to other effects of the ryegrass rhizosphere, indicating that supplementing soils with root

418

exudates have a minor contribution to phenanthrene degradation compared to the rhizosphere.

419

Although many studies have reported that root exudates can significantly modify the bacterial

420

community structure and microbial functions in the rhizosphere (Baudoin et al., 2003; Haichar et

421

al., 2008; Cebron et al., 2011), only limited attempts are made to address the changes in

422

degradation performance and the active degraders for organic pollutants affected by root exudates.

423

In the present study, the ryegrass rhizosphere and its root exudates markedly increased the total

424

bacterial population and modified the bacterial community structure in PAH-contaminated soil,

425

similar to the findings of a previous report (Guo et al., 2017). Two bacterial taxa significantly

426

stimulated in the ryegrass rhizosphere were unclassified Comamonadaceae and Streptophyta.

427

Comamonadaceae dominates soil profile (Huang et al., 2013) and barley root-enriched microbiota

428

(Bulgarelli et al., 2015), and the relative abundance of Comamonadaceae is positively correlated

429

with soil-available manganese, calcium, copper, and potassium concentrations (Huang et al.,

430

2013). Members of the family Comamonadaceae can degrade different organic pollutants,

431

including phenol, toluene, naphthalene, and phenanthrene (Sun and Cupples, 2012; Reunamo et al.,

432

2017). Streptophyta is common in lichen crust, and increases with the development of biological

433

soil crust (Zhang et al., 2016). In addition, the ryegrass rhizosphere affected the composition of

434

abundant bacteria, such as members of the family Micrococcaceae. In another study,

435

Micrococcaceae was the most frequently detected family within its phylum as an important

436

cellobiose- and glucose-degrading group under oxic conditions (Schellenberger et al., 2010).

437

However, there was no evidence from the SIP results directly linking the above abundant bacteria

438

(except Micrococcaceae) to phenanthrene degradation. The low correlation between the

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16

ACCEPTED MANUSCRIPT 439

phenanthrene degradation efficiency and abundance of total 16S rRNA genes (Table 2) also

440

supported the finding that the total bacterial population had a minor contribution to phenanthrene

441

metabolism. Instead, these soil prokaryotes might function to stabilise the microbial community or

442

degrade other organic molecules in PAH-contaminated soils (Li et al., 2017a). Besides the abundance and structure of the whole microbial community, the ryegrass

444

rhizosphere might have shaped the composition of the active phenanthrene-degrader community

445

by enriching and encouraging PAH degraders to enhance contaminant degradation (Guo et al.,

446

2017). In addition, root exudates alone have been reported to modify the diversity of active

447

phenanthrene degraders in soils contaminated with PAHs (Cebron et al., 2011). However, no

448

studies have distinguished the effects of root exudates within the ryegrass rhizosphere in shaping

449

the diversity of the active phenanthrene degraders in PAH-polluted soil. In the AD treatments, the

450

indigenous microorganisms responsible for phenanthrene degradation were affiliated with

451

Blastomonas, Mucilaginibacter, Nitrososphaera, and unclassified Chthoniobacteraceae. Among

452

them, only two taxa (Blastomonas and Nitrososphaera) remained active phenanthrene degraders

453

in both the RE and RG treatments. Six additional indigenous phenanthrene-degrading

454

microorganisms became active in the RE treatments, phylotypes affiliated with Sphingomonas,

455

Pseudomonas, Herbaspirillum, Ramlibacter, Mycobacterium, and unclassified Pyrinomonadaceae.

456

In the RG treatments, another six new taxa Sphingoaurantiacus, Kaistobacter, Novosphingobium,

457

Pedobacter, Micrococcaceae (genus unclassified) and Chitinophagaceae (genus unclassified)

458

were favoured. Among all the PHE degraders, Blastomonas, Sphingoaurantiacus, Ramlibacter,

459

Mucilaginibacter,

460

Chitinophagaceae, and Chthoniobacteraceae were found, for the first time, to be directly

461

responsible for indigenous phenanthrene biodegradation in soils.

SC

M AN U

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EP Pedobacter,

Nitrososphaera,

unclassified

Pyrinomonadaceae,

AC C

462

RI PT

443

Alphaproteobacteria and Nitrososphaeria were the two key phenanthrene-degrading bacterial

463

groups observed in all AD, RE, and RG treatments (Figure S10). Of the active degraders

464

belonging to the class Alphaproteobacteria, only Sphingomonas and Novosphingobium are known

465

to degrade various environmental contaminants such as chlorinated compounds and PAHs (Tiirola

466

et al., 2002; Liu et al., 2016; Mulla et al., 2016; Segura et al., 2017), and phenanthrene degradation

467

by Sphingomonas and Novosphingobium has been reported (Liu et al., 2016; Fida et al., 2017). We

468

previously applied DNA-SIP with 13C-phenanthrene as the substrate and identified Kaistobacter as 17

ACCEPTED MANUSCRIPT an indigenous phenanthrene degrader in activated sludge (Li et al., 2017b). However, the functions

470

of Blastomonas and Sphingoaurantiacus in PAH degradation remain unknown. Since they have

471

not been linked with phenanthrene degradation, their roles remain unclear at PAH-contaminated

472

sites. Archaea have an important role in PAH degradation (Ma et al., 2015), but research on

473

archaeal PAH bioremediation mechanisms and pathways is in its infancy (Ghosal et al., 2016).

474

Some Archaea can persist in PAH-contaminated marine samples, implying the potential capability

475

to metabolise PAHs (Harada et al., 2013). In the present study, the genus Nitrososphaera of

476

Archaea was associated with phenanthrene degradation in all treated microcosms. This is the first

477

evidence of Nitrososphaera’s ability to degrade phenanthrene. Our SIP results provide direct

478

evidence that some of the above genera within the classes Alphaproteobacteria and

479

Nitrososphaeria are key functional phenanthrene-degrading microbes in PAH-polluted soil.

M AN U

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469

Some bacterial classes were observed in two treatments (Figure S10), including

481

Sphingobacteriia (AD and RG microcosms) and Actinobacteria (RG and RE microcosms). The

482

genera Mucilaginibacter and Pedobacter are members of the family Sphingobacteriaceae within

483

the class Sphingobacteriia. Members of the genus Mucilaginibacter can degrade organic matter

484

such as xylan, pectin, and laminarin (Khan et al., 2013a). Pedobacter sp. was isolated from

485

seedlings exposed to PAHs (Zhu et al., 2016) and identified as a soil alkane-degrading bacterium

486

in oil-exploration areas (Yang et al., 2015). Moreover, members of Chitinophagaceae within the

487

class Sphingobacteriia have been reported as benzo[a]pyrene-metabolising bacteria in forest soils

488

based on SIP (Song et al., 2015). However, these members of Bacteroidetes have not been

489

previously linked to phenanthrene degradation. Some bacteria assigned to Actinobacteria (e.g.

490

Mycobacterium and Micrococcaceae) have exhibited phenanthrene degradation capabilities

491

(Hennessee and Li, 2016; Li et al., 2017b). Moreover, Mycobacterium can degrade various

492

environmental contaminants such as polychlorobiphenyls and PAHs (Hormisch et al., 2004). Since

493

the above phenanthrene-degrading groups were observed in only two treatments, we speculated

494

that their roles in phenanthrene metabolism were altered by the external environment.

495

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480

Some functional phenanthrene degraders were only observed in one treatment, such as

496

Chthoniobacteraceae

(Spartobacteria)

497

(Gammaproteobacteria),

498

Pyrinomonadaceae (Blastocatellia) in the RE treatment. Bacteria closely related to

Herbaspirillum

in

the and

18

AD

microcosm

Ramlibacter

and

Pseudomonas

(Betaproteobacteria),

and

ACCEPTED MANUSCRIPT Verrucomicrobia are favoured in the ryegrass rhizosphere (Cebron et al., 2015). However,

500

Chthoniobacteraceae (class Spartobacteria) within this phylum was detected and identified as an

501

active phenanthrene degrader in soil without ryegrasses or its root exudates. Although

502

Verrucomicrobia might be important in PAH rhizoremediation (Kawasaki et al., 2012), the

503

functions of the family Chthoniobacteraceae in PAH degradation have not been documented. The

504

addition of root exudates to the soil stimulated the potential functions of three classes

505

(Gammaproteobacteria, Betaproteobacteria, and Spartobacteria) to metabolise phenanthrene.

506

Members of the genus Pseudomonas (class Gammaproteobacteria) are affiliated with PAH

507

degraders and contain PAH-RHDα genes encoding the PAH degradative pathway (Thomas et al.,

508

2016). Some Herbaspirillum strains (class Betaproteobacteria, order Burkholderiales) have been

509

described as root-associated nitrogen-fixing bacteria (Bajerski et al., 2013). Herbaspirillum was

510

able to use PAHs, such as phenanthrene, anthracene, fluoranthene and benzo[b]fluoranthene as

511

sole carbon and energy source (Xu et al., 2016). Like Herbaspirillum, Ramlibacter also belongs to

512

the order Burkholderiales. Members of the family Pyrinomonadaceae (phylum Acidobacteria,

513

class Spartobacteria) can grow in the presence of glucose, mannose, or formate (Wust et al., 2016).

514

However, no information is available on the possible functions of the genus Ramlibacter and

515

unclassified Pyrinomonadaceae in PAH degradation because they have not been linked with

516

phenanthrene degradation previously and their roles in PAH-contaminated soil remain unclear.

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499

Our results provide evidence that the ryegrass rhizosphere increased the abundance of

518

PAH-RHDα genes in active phenanthrene degraders participating in phenanthrene metabolism,

519

which was concordant with Khan’s work (2009) reporting that ryegrass rhizosphere increased the

520

number of cultivable pyrene degraders in soils (Khan et al., 2009). This suggested that the

521

rhizosphere may enhance the activities of bacteria with PAH-RHDα genes to promote

522

phenanthrene degradation, which was further supported by the results of strong correlations

523

between the abundance of PAH-RHDα genes and active phenanthrene-degraders (Table 2).

524

Although previous observations showed that rhizosphere increased the diversity of

525

aromatic-dioxygenase genes (Cebron et al., 2011), few studies have attempted to link PAH

526

degradation to the abundance of PAH-RHDα genes in the rhizosphere (Thomas and Cebron, 2016).

527

In the present study, according to correlation analysis, the abundances of PAH-RHDα genes and

528

active phenanthrene-degraders have strong correlations with phenanthrene degradation efficiency.

AC C

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517

19

ACCEPTED MANUSCRIPT The abundance of PAH-RHDα genes was previously reported to correlate with PAH degradation

530

(Ding et al., 2010), and the temporal changes in activities of PAH degraders bearing PAH-RHDα

531

genes during PAHs degradation process demonstrated its strong link with bacterial degradation

532

capability (Guo et al., 2017). Additionally, although the higher populations of total bacteria in soils

533

with rhizosphere and root exudates, compared to those in bulk soils, suggested that both of them

534

significantly stimulated bacterial growth in PAH-contaminated soils (Cebron et al., 2011; Guo et

535

al., 2017), no correlation between the total bacterial population and phenanthrene degradation

536

efficiency was observed in the present study. Our results suggested that root exudates played key

537

roles in the ryegrass rhizosphere to shape the whole microbial communities, but had minor

538

contribution to phenanthrene degradation, also supported by the difference in active phenanthrene

539

degraders and PAH-RHDα genes between the AD, RE and RG treatments.

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20

ACCEPTED MANUSCRIPT 540

5. Conclusions Our study is the first to show the different roles of the ryegrass rhizosphere and root exudates

542

supplement in influencing the abundance and diversity of the active PAH-degrading bacteria in

543

PAH-contaminated soils via SIP. Soils planted with ryegrass significantly increased the

544

phenanthrene degradation efficiencies, but no significant difference was observed between soils

545

with and without root exudates. Furthermore, PAH-RHDα gene abundance exhibited similar trends

546

and was significantly lower in the root exudate microcosms than the rhizosphere, consistent with

547

our results regarding the abundance of active phenanthrene degraders. These results suggest that

548

root exudates offer a minor contribution to accelerated in situ phenanthrene remediation within the

549

rhizosphere. Moreover, the differences in the phenanthrene degradation efficiencies were

550

attributed to the distinct communities of active phenanthrene degraders. The indigenous

551

microorganisms responsible for phenanthrene degradation belonged to eight bacterial classes, and

552

only two key classes (Alphaproteobacteria and Nitrososphaeria) were observed in all three

553

treatments. Instead, phenanthrene-degrading Sphingobacteriia and Actinobacteria were active in

554

phenanthrene degradation in two treatments (AD + RG and RG + RE, respectively), whereas the

555

other four phenanthrene-degrading microbes were only observed in the AD or RE treatments.

556

These findings suggest that active phenanthrene degraders within the same microbial community

557

might be altered by environmental changes. Of the phenanthrene degraders, Blastomonas,

558

Sphingoaurantiacus,

559

unclassified Pyrinomonadaceae, Chitinophagaceae, and Chthoniobacteraceae were found, for the

560

first time, to be directly responsible for indigenous phenanthrene biodegradation. The correlation

561

analysis further indicated that the PAH-RHDα genes and the abundance of phenanthrene degraders

562

had strong correlations with the phenanthrene degradation efficiency. This study helps in a more

563

complete understanding of the diversity of phenanthrene-degrading communities in the

564

rhizosphere during PAH degradation, and provides theoretical insights into the mechanisms of

565

enhanced phenanthrene degradation via phytoremediation at PAH-contaminated sites.

TE D

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541

Mucilaginibacter,

Pedobacter,

Nitrososphaera,

and

AC C

EP

Ramlibacter,

566

21

ACCEPTED MANUSCRIPT Acknowledgements

568

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

569

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

570

41673111), the National Postdoctoral Program for Innovative Talents (BX20180308) and the

571

Department of Science and Technology of Guangdong province (2016TQ03Z938).

RI PT

567

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22

ACCEPTED MANUSCRIPT Reference

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Environmental Microbiology 68, 173-180. Werner, J.J., Koren, O., Hugenholtz, P., Desantis, T.Z., Walters, W.A., Caporaso, J.G., Angenent, L.T., Knight, R., Ley, R.E., 2012. Impact of training sets on classification of high-throughput bacterial 16s rRNA gene surveys. Isme Journal 6, 94-103. Wust, P.K., Foesel, B.U., Geppert, A., Huber, K.J., Luckner, M., Wanner, G., Overmann, J., 2016. Brevitalea aridisoli, B-deliciosa and Arenimicrobium luteum, three novel species of Acidobacteria subdivision 4 (class Blastocatellia) isolated from savanna soil and description of the novel family Pyrinomonadaceae. International Journal of Systematic and Evolutionary Microbiology 66, 3355-3366. Xu, H.X., Li, X.H., Sun, Y.Y., Shi, X.Q., Wu, J.C., 2016. Biodegradation of Pyrene by Free and Immobilized Cells of Herbaspirillum chlorophenolicum Strain FA1. Water Air and Soil Pollution 227, 120. Yang, Y.Y., Wang, J., Liao, J.Q., Xie, S.G., Huang, Y., 2015. Abundance and diversity of soil petroleum hydrocarbon-degrading microbial communities in oil exploring areas. Applied Microbiology and Biotechnology 99, 1935-1946. Yu, X.Z., Wu, S.C., Wu, F.Y., Wong, M.H., 2011. Enhanced dissipation of PAHs from soil using mycorrhizal ryegrass and PAH-degrading bacteria. Journal of Hazardous materials 186, 1206-1217. Zhang, B., Kong, W., Wu, N., Zhang, Y., 2016. Bacterial diversity and community along the succession of biological soil crusts in the Gurbantunggut Desert, Northern China. Journal of Basic Microbiology 56, 670-679. Zhu, X.Z., Jin, L., Sun, K., Li, S., Li, X.L., Ling, W.T., 2016. Phenanthrene and Pyrene Modify the Composition and Structure of the Cultivable Endophytic Bacterial Community in Ryegrass (Lolium multiflorum Lam). International Journal of Environmental Research and Public Health 13, 1081.

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Captions

776

Figure 1. Residual PHE percentage in AD, RE and RG microcosms. C12 and C13 refer to

777

samples amended with

778

means ± standard deviation; n = 3.

779

Figure 2. Heatmap of the OTU abundance changes in the heavy fractions from the AD, RE, and

780

RG treatments. The heatmap was generated using the relative abundance of OTUs enriched in the

781

heavy fractions from the

782

12

783

ranging from green (low) to yellow (high). 4D, 8D, and 12D represent samples collected on days 4,

784

8, and 12, respectively. C12 and C13 refer to samples amended with

785

13

786

Figure 3. Phylogenetic tree of OTUs responsible for phenanthrene degradation based on the

787

neighbour-joining method using 16S rRNA gene sequences, showing the phylogenetic positions of

788

the bacteria corresponding to OTUs and representatives of some related taxa. The bar represents

789

0.05 substitutions per nucleotide position. The active phenanthrene degraders identified from the

790

AD, RE, and RG treatments are highlighted in red, green and grey, respectively. 4D, 8D, and 12D

791

represent samples collected on days 4, 8, and 12, respectively.

792

Figure 4. Relative abundance of PAH-RHDα genes from the CsCl gradient fractions from the AD,

793

RE, and RG treatments after 4, 8, and 12 days of incubation with

794

13

795

within each density gradient fraction was quantified via qPCR.

C_AD,

13

C_RE, and

13

13

C-phenanthrene, respectively. Data are

RI PT

12

13

C-phenanthrene and

C_RG treatments compared to the

12

C_AD,

C_RG treatments, respectively. The colours indicate the relative abundance,

SC

C_RE, and

12

C-phenanthrene and

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C-phenanthrene, respectively.

12

12

C-phenanthrene or

AC C

C-phenanthrene. The heavy DNA fractions are highlighted in grey. The relative abundance

28

ACCEPTED MANUSCRIPT 796

Figure 1

120

Sterile control AD-C12 AD-C13 RE-C12 RE-C13 RG-C12 RG-C13

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80

60

40

20

0

2

4

6

8

10

12

797

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Sampling time (d)

SC

Residual PHE (%)

100

798

Figure 1. Residual PHE percentage in AD, RE and RG microcosms. C12 and C13 refer to

799

samples amended with

800

means ± standard deviation; n = 3.

EP

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C-phenanthrene and

AC C

801

12

29

13

C-phenanthrene, respectively. Data are

ACCEPTED MANUSCRIPT Figure 2

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802

803

Figure 2. Heatmap of the OTU abundance changes in the heavy fractions from the AD, RE, and

805

RG treatments. The heatmap was generated using the relative abundance of OTUs enriched in the

806

heavy fractions from the

807

12

808

ranging from green (low) to yellow (high). 4D, 8D, and 12D represent samples collected on days 4,

809

8, and 12, respectively. C12 and C13 refer to samples amended with

810

13

EP

804

C_AD,

13

C_RE, and

AC C

13

C_RE, and

12

13

C_RG treatments compared to the

12

C_AD,

C_RG treatments, respectively. The colours indicate the relative abundance,

C-phenanthrene, respectively.

811

30

12

C-phenanthrene and

ACCEPTED MANUSCRIPT 812

Figure 3 OTU 12876 Sphingoaurantiacus RG-8D Sphingoaurantiacus polygranulatus MC 3718 (NR 147725.1) OTU 14966 Kaistobacter RG-12D Uncultured Kaistobacter sp. clone Plot18-G06 (FJ889333.1) OTU 41564 Sphingomonas RE-12D Sphingomonas sp. FR3 AP1 09 (KX832161.1)

Alphaproteobacteria Sphingomonadacea

OTU 3175 Blastomonas RE-8D OTU 3175 Blastomonas RG-4D OTU 3175 Blastomonas AD-4D Blastomonas natatoria M5 (LC191979.1) OTU 23284 Pseudomonas RE-4D Pseudomonas frederiksbergensis D (MF139035.1)

OTU 753 Ramlibacter RE-12D Ramlibacter sp. B534 (KY122000.1)

Gammaproteobacteria Betaproteobacteria

SC

OTU 34628 Herbaspirillum RE-8D Herbaspirillum frisingense AA6 (KX817278.1)

RI PT

OTU 15584 Novosphingobium RG-4D Novosphingobium aromaticivorans F9 (KU924009.1)

OTU 19134 Chitinophagaceae RG-12D Uncultured Chitinophagaceae clone CNY 00848 (JQ400912.1)

M AN U

OTU 53992 Pedobacter RG-12D Pedobacter humi THG S15-2 (NR 149285.1)

OTU 1476 Mucilaginibacter AD-4D Uncultured Mucilaginibacter sp. clone LWM2-4 (HQ674876.1)

Uncultured Acidobacteria bacterium clone CA8 (KJ191797.1) Uncultured Acidobacteria bacterium clone M1-027 (KF182871.1) OTU 28345 Pyrinomonadaceae RE-4D OTU 32213 Chthoniobacteraceae AD-12D Uncultured Verrucomicrobia clone RozowormcastElong3H3 (HM444706.1)

Bacteroidetes Sphingobacteriia Sphingobacteriales Blastocatellia Verrucomicrobia Spartobacteria

OTU 42503 Mycobacterium RE-4D Mycobacterium tusciae TSDHB10-85H (LT853783.1)

TE D

OTU 5288 Micrococcaceae RG-4D Pseudarthrobacter phenanthrenivorans Sphe3 (NR 074770.2) Arthrobacter globiformis SPB25 (KY082736.1) OTU 42987 Nitrososphaera RG-12D Uncultured Archaeon clone B118 (KX061174.1) OTU 42987 Nitrososphaera AD-8D OTU 34306 Nitrososphaera RG-12D Uncultured Nitrososphaera sp. clone ncaOTU35 (KY225301.1)

813

Archaea Thaumarchaeota Nitrososphaeria

OTU 34691 Nitrososphaera AD-8D Uncultured Nitrososphaera sp. clone OF-63(KP109877.1)

AC C

0.05

EP

OTU 34691 Nitrososphaera RE-12D

Actinobacteria

814

Figure 3. Phylogenetic tree of OTUs responsible for phenanthrene degradation based on the

815

neighbour-joining method using 16S rRNA gene sequences, showing the phylogenetic positions of

816

the bacteria corresponding to OTUs and representatives of some related taxa. The bar represents

817

0.05 substitutions per nucleotide position. The active phenanthrene degraders identified from the

818

AD, RE, and RG treatments are highlighted in red, green and grey, respectively. 4D, 8D, and 12D

819

represent samples collected on days 4, 8, and 12, respectively.

820 31

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Figure 4 60000

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100000

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12C-PHE 13C-PHE

40000

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80000

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Buoyant Density (g/mL)

822

Figure 4. Relative abundance of PAH-RHDα genes from the CsCl gradient fractions from the AD,

824

RE, and RG treatments after 4, 8, and 12 days of incubation with

825

13

826

within each density gradient fraction was quantified via qPCR.

12

C-phenanthrene or

C-phenanthrene. The heavy DNA fractions are highlighted in grey. The relative abundance

AC C

827

EP

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32

ACCEPTED MANUSCRIPT Table 1. Soil characteristics and PAH concentrations. Content

Clay

5.12%

Silt

48.88%

Sand

46%

pH

6.8

C/N ratio

22.7

PAHs

Concentration (µg/Kg)

Naphthalene

25.96-29.04

Acenaphthylene

0.23-0.26

Acenaphthene

4.09-4.58

56.18-62.1

Phenanthrene

3.83-4.28

Anthracene

25.32-28.32

TE D

Fluoranthene

Chrysene

SC

M AN U 9.77-10.93

Fluorene

Pyrene

RI PT

Soil characteristics

27.23-30.45 190.21-211.14 49.83-54.56

Benzo(b)fluoranthene

84.07-95.08

Benzo(k)fluoranthene

10.08-12.28

Benzo(a)pyrene

51.19-63.26

Indeno(1,2,3-cd)pyrene

8.98-10.05

Dibenzo(a,h)anthracene

18.99-19.77

Benzoperylene

46.18-51.66

EP

Benzo(a)anthracene

AC C

828

33

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Table 2. Pearson correlation coefficient between the abundance of total 16S rRNA gene, active phenanthrene-degrading bacteria, PAH-RHDɑ genes and

830

phenanthrene degradation efficiency after 12 days of incubation.

Total 16S rRNA

efficiency

gene

-

efficiency

Abundance of active degrading

0.581

1.000**

831

*. Correlation is significant at the 0.05 level.

832

**. Correlation is significant at the 0.01 level.

-

-

-

-

-

-

0.586

-

-

0.999*

0.610

1.000*

-

AC C

Abundance of PAH-RHDα genes

EP

bacteria

genes

TE D

Total 16S rRNA gene

Abundance of PAH-RHDα

bacteria

M AN U

Phenanthrene degradation

Abundance of active degrading

SC

Phenanthrene degradation

RI PT

829

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Highlights



First SIP exploration of roles of root exudates and rhizosphere in PAH degradation Most of the identified degraders were linked to PHE degradation for the first time



Strong correlation between the active PHE-degraders and PHE degradation

RI PT



efficiency 

Strong correlation between the active PAH-RHDα gene and PHE degradation

Root exudates offer a minor contribution to PAH remediation within the

EP

TE D

M AN U

rhizosphere

AC C



SC

efficiency