Journal of Hazardous Materials 272 (2014) 10–19

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The effect of ammonium chloride and urea application on soil bacterial communities closely related to the reductive transformation of pentachlorophenol Huan-Yun Yu a , Yong-kui Wang b , Peng-cheng Chen a , Fang-bai Li a,∗ , Man-jia Chen a , Min Hu a a Guangdong Key Laboratory of Agricultural Environment Pollution Integrated Control, Guangdong Institute of Eco-Environmental and Soil Sciences, Guangzhou, PR China b Environmental Science and Engineering College, Hubei Polytechic University, Huangshi 435003, Hubei, PR China

h i g h l i g h t s • • • • •

Low concentrations of NH4 Cl/CO(NH2 )2 enhanced the reductive transformation of PCP. High concentrations of NH4 Cl/CO(NH2 )2 inhibited the reductive transformation of PCP. For the genus Shewanella, all concentrations of CO(NH2 )2 showed enhancement effects. The variation in the trends of PCP transformation and in the abundance of the genus Dehalobacter was consistent. Both NH4 Cl and CO(NH2 )2 had inhibitory effects on the growth of the genus Comamonas.

a r t i c l e

i n f o

Article history: Received 20 July 2013 Received in revised form 8 October 2013 Accepted 24 February 2014 Available online 12 March 2014 Keywords: Pentachlorophenol (PCP) Nitrogen fertilizer Reductive transformation Dissimilatory iron-reducing bacteria (DIRB) Dechlorinating bacteria

a b s t r a c t Pentachlorophenol (PCP) is widely distributed in the soil, and nitrogen fertilizer is extensively used in agricultural production. However, studies on the fate of organic contaminants as affected by nitrogen fertilizer application have been rare and superficial. The present study aimed to examine the effect of ammonium chloride (NH4 Cl) and urea (CO(NH2 )2 ) application on the reductive transformation of PCP in a paddy soil. The study showed that the addition of low concentrations of NH4 Cl/CO(NH2 )2 enhanced the transformation of PCP, while the addition of high concentrations of NH4 Cl/CO(NH2 )2 had the opposite effect. The variations in the abundance of soil microbes in response to NH4 Cl/CO(NH2 )2 addition showed that both NH4 Cl and CO(NH2 )2 had inhibitory effects on the growth of dissimilatory iron-reducing bacteria (DIRB) of the genus Comamonas. In contrast, for the genus Shewanella, low concentrations of NH4 Cl inhibited growth, and high concentrations of NH4 Cl enhanced growth, whereas all concentrations of CO(NH2 )2 showed enhancement effects. In addition, consistent patterns of variation were found between the abundances of dechlorinating bacteria in the genus Dehalobacter and PCP transformation rates under NH4 Cl/CO(NH2 )2 addition. In conclusion, nitrogen application produced variations in the structure of the soil microbial community, especially in the abundance of dissimilatory iron-reducing bacteria and dechlorinating bacteria, which, in turn, affected PCP dechlorination. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Pentachlorophenol (PCP), first manufactured commercially in 1936, was widely used as a wood preservative, herbicide, insecticide, fungicide, molluscicide and bactericide [1]. It can cause adverse effects on all forms of life because of its disruptions of the

∗ Corresponding author. Tel.: +86 20 87024721; fax: +86 20 87024123. E-mail address: [email protected] (F.-b. Li). http://dx.doi.org/10.1016/j.jhazmat.2014.02.037 0304-3894/© 2014 Elsevier B.V. All rights reserved.

integrity and function of biological membranes [2]. Human exposure to PCP occurs primarily through skin absorption, inhalation and ingestion [3,4]. Because of the possible adverse effects of PCP on humans and the eco-environment, many countries have restricted or banned the production and use of PCP. In China, the use of PCP was restricted or banned in 1997. Approximately 104 t of PCP per year were produced in China, representing approximately 20% of global production [5]. Although the production and use of PCP have been restricted or banned, PCP levels in Chinese surface water and sediments and marine vertebrates show slightly increasing trends,

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which may result from the use of Na-PCP to control schistosomiasis in certain areas [6]. Additionally, the amount of nitrogen fertilizer applied worldwide has increased rapidly, growing at an annual rate of 61% from 1961 to 1999 with the continued growth of global crop yields [7]. Chinese consumption of nitrogen fertilizer represents approximately 30% of the global total [7]. Moreover, Guangdong province, one of the most agriculturally well-developed regions of the country, consumed approximately 1/10 of the total amount of nitrogen fertilizer applied in China [8,9]. Nitrogen, as a necessary element for microorganism growth, can influence the functional diversity of soil microbes [10]. Previous reports have suggested that N addition can accelerate the degradation of organic contaminants by stimulating microbial activity [11]. In contrast, N addition can also inhibit the degradation of organic pollutants as a result of the inhibition of enzymatic systems responsible for the degradation of organic pollutants [12]. Based on these results, we speculate that soil microbes may play an important role in the transformation of organic pollutants under the condition of N addition. Previous studies have suggested that PCP is biodegradable and that bioremediation may be the most efficient and economical approach to the removal of PCP from soil [13]. Additionally, anaerobic reductive dechlorination is a crucial pathway for the degradation of highly chlorinated organic compounds such as PCP [14], because it is more difficult for bacteria to use highly chlorinated organic compounds under aerobic conditions [15]. Under anaerobic conditions, dechlorinating bacteria can use chlorinated compounds effectively as terminal electron acceptors coupled with the oxidation of available substrates. Currently, isolates capable of using chlorinated compounds as metabolic electron acceptors are primarily found in the taxa Clostridia, Deltaproteobacteria, Epsilonproteobacteria and Gammaproteobacteria [16]. In addition, dissimilatory iron-reducing bacteria (DIRB), capable of coupling the oxidation of organic substrates with the reduction of Fe(III) minerals, also play an important role in the processes of transformation of chlorinated organic contaminants [17]. Many previous reports have confirmed that dissimilatory Fe(III)-reducing bacteria can reduce a range of chlorinated compounds [18–20]. Furthermore, Fe(III) reduction and reductive dechlorination can occur concurrently in contaminated environments [18,21]. Additionally, Fe(III) reduction may actually facilitate reductive dechlorination. This effect is attributable to the promotion of the growth of DIRB, which furthers reductive dechlorination [21]. Our recent study [22] also demonstrated that Fe(III)-reducing bacteria are of great importance for PCP dechlorination. In particular, DIRB transferred electrons to iron minerals for the generation of active Fe(II) species, and the higher concentration of Fe(II) then enhanced the transformation of PCP. Therefore, DIRB, Fe(III) reduction and the reductive dechlorination of organic contaminants are highly correlated. Moreover, the soils of Guangdong province, located in subtropical and tropical zones, are red soils, which contain a large amount of free Fe oxide [23]. Therefore, the objective of the present study was to examine the effect of nitrogen, including ammonium chloride (NH4 Cl) and urea (CO(NH2 )2 ) addition, on the reductive transformation of PCP in paddy soil. The focus of the study was the variations in microbial community structure induced by nitrogen addition. In addition, the interaction between Fe(III) reduction and the reductive transformation of PCP was investigated.

2. Materials and methods 2.1. Soil sampling Paddy soil sample collected from the town of Shahu in the city of Enping (22◦ 23.11 N, 112◦ 26.85 E), located in southwestern

11

Guangdong province in southern China during September and December 2008, was used for batch incubation experiments. After transport to the laboratory, the soil was sealed in polytetrafluoroethylene (PTFE) bag and stored in glass bottles at 4 ◦ C prior to use. 2.2. Chemicals PCP (≥98% purity) and 1,4-piperazinediethanesulfonic acid (PIPES) (≥98% purity) were all purchased from Sigma-Aldrich (St. Louis, MO, USA). The other analytical-grade chemicals, including NH4 Cl and CO(NH2 )2 , were all purchased from Guangzhou Chemical Company (Guangzhou, China). Deaerated deionized water was prepared by deoxygenating ultrapure water (18 M cm, Easy Pure II RF/UV USA). 2.3. PCP transformation experiments in soils Experiments were conducted in triplicate at a constant pH of 7.0 ± 0.2 with 30 mM PIPES for the treatments with NH4 Cl and 50 mM PIPES for the treatments with CO(NH2 )2 as the buffer solution. The batch experimental procedures are as follows: the soil samples (0.5 g dry weight) were transferred into 20 mL serum bottles with silicone-lined septa and an aluminum sealing cap. A 10 mL PIPES buffer solution was then added. Six batch experiments, including a control, were conducted for the treatments with NH4 Cl: (1) control (0 mM NH4 Cl); (2) 2.5 mM NH4 Cl; (3) 5 mM NH4 Cl; (4) 10 mM NH4 Cl; (5) 20 mM NH4 Cl; and (6) 30 mM NH4 Cl. Five batch experiments, including a control, were conducted for the treatments with CO(NH2 )2 : (1) control (0 mM CO(NH2 )2 ); (2) 5 mM CO(NH2 )2 ; (3) 10 mM CO(NH2 )2 ; (4) 15 mM CO(NH2 )2 ; and (5) 20 mM CO(NH2 )2 . Subsequently, PCP and lactic acid at final concentrations of 0.0188 mM and 10 mM, respectively, were added to each vial. The mixture was then purged using O2 -free N2 for 30 min and sealed with Teflon-coated butyl rubber stoppers and crimp seals. The closed bottles were mixed on a rotary shaker and incubated at (30 ± 1) ◦ C in an anaerobic chamber. Lastly, the bottles were sampled for analysis at predetermined intervals. 2.4. Chemical analysis methods The PCP in the soil suspension was extracted with a water/ethanol mixture (50:50 in volume) on a horizontal shaker (180 rpm) for 1 h [24]. The suspension was then filtered through a 0.45 ␮m syringe filter and collected for high performance liquid chromatography analysis. The detailed analysis procedures for PCP have been described in a previous study [25]. HCl-extractable Fe(II) and dissolved Fe (II) were measured using the 1,10-phenanthroline colorimetric method at 510 nm on a UV–Vis spectrophotometer (TU-1810PC, Beijing Purkinje General Instruments, China) [26]. The concentration of NH4 + was measured by the Nesslerization method using a UV–Vis spectrophotometer (TU-1810PC, Beijing Purkinje General Instruments, China) at 420 nm. 2.5. DNA extraction, terminal restriction fragment length polymorphism (T-RFLP) analysis and clone library construction Detailed protocols for T-RFLP analysis and clone library construction have been described previously [27]. Briefly, total soil DNA was extracted using the PowerSoilTM DNA isolation kit (Mo Bio Laboratories, USA). The bacterial 16S rRNA gene was amplified using PCR primers 27F labeled with 6-FAM on the 5 end and 1492R. The labeled PCR products were then purified with a commercial PCR purification kit (OMEGA bio-tek, USA), and aliquots were digested with AluI restriction enzymes (TaKaRa Biotechnology, China). The digested PCR products were resolved by electrophoresis using an

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ABI 3730xl sequencer (Applied Biosystems, USA) with GS-500 Rox as an internal size standard. The relative abundance of the individual terminal restriction fragment (T-RF) was calculated as the percentage of its peak area covering the total peak areas in the T-RFLP profile, and only those T-RFs with a relative abundance >1% were used for the following analysis. For clone library construction, the bacterial 16S rRNA gene was first amplified with the bacterial primers 27F and 1492R. The purified PCR products were then ligated into the vector PCR2.1 TOPO (Invitrogen, USA) and transformed into Escherichia coli DH5␣ competent cells (TaKaRa Biotechnology, China). The cloned inserts were sequenced using an ABI 3730xl sequencer. After quality filtering and chimera removal, the taxonomy of the 16S rRNA gene sequence was classified using the RDP pipelines.

1.0

0 mM NH4Cl

(A)

2.5 mM NH4Cl 5 mM NH4Cl

.8

10 mM NH4Cl 20 mM NH4Cl

.6

30 mM NH4Cl

Degradation fraction of PCP

.4

2.6. Quantitative real-time PCR Quantifications of total bacteria, Shewanella spp., Geobacter spp. and Dehalobacter spp., were performed using a MyiQTM 2 Optics Module (BIO-RAD, USA). The primers that were used and PCR conditions were detailed in a previous study [27]. The quantitative real-time PCR calibration curves were generated with serial dilutions of the concentration of plasmids containing the cloned target sequences. The plasmid DNA concentration was quantified with a Qubit 2.0 Fluorometer (Invitrogen, NY, USA), and the corresponding gene copy number was calculated according to a previous study [28].

.2 0.0

1.0

0 mM CO(NH2)2 5 mM CO(NH2)2

(B)

10 mM CO(NH2)2

.8

15 mM CO(NH2)2 20 mM CO(NH2)2

.6 .4 .2 0.0 0

3

6

9

12

Reaction time (day) 2.7. Data analysis Data comparison among groups was performed using oneway ANOVA and a univariate general linear model (GLM) with statistical significance p < 0.05 in SPSS version 13.0. The obtained sequences in the present study were compared to public databases using blastn to identify closely related sequences (http://www.ncbi.nlm.nih.gov/blast/). 3. Results

Fig. 1. The variation in the degradation fractions of pentachlorophenol (PCP) in different treatments with incubation time.

concentrations of NH4 Cl/CO(NH2 )2 addition can enhance the degradation of PCP, whereas high concentrations of NH4 Cl/CO(NH2 )2 addition may inhibit the degradation of PCP. In the present study, the concentration of the added NH4 Cl/CO(NH2 )2 appeared to be an important factor affecting the degradation of PCP, and 5 mM and 10 mM were the turning points for the NH4 Cl and CO(NH2 )2 treatments, respectively.

3.1. Reductive transformation of PCP 3.2. Iron reduction The time course of the degradation of PCP was fitted to a logistic degradation curve and the maximum transformation rate (m ) of PCP were derived (Table 1). The maximum transformation rate of PCP increased gradually as the concentration of NH4 Cl varied between 0 mM and 5 mM (Fig. 1A). In contrast, the maximum transformation rate showed a decreasing trend as the concentration of NH4 Cl varied from 5 mM to 30 mM (Fig. 1A). Similarly, the maximum transformation rate of PCP increased gradually as the concentration of CO(NH2 )2 varied between 0 mM and 10 mM (Fig. 1B). In comparison, the maximum transformation rate showed a decreasing trend as the concentration of CO(NH2 )2 varied from 10 mM to 30 mM (Fig. 1B). These results indicated that low

The concentrations of dissolved and HCl-extractable Fe(II) in all treatments increased markedly during the first 3 days (Figs. 2 and 3). After 3 days, the concentration of dissolved Fe(II) continued to increase slowly in the NH4 Cl treatments and became steady after 8 days, whereas the concentration of dissolved Fe(II) first increased slowly and then decreased in the CO(NH2 )2 treatments. In contrast, the concentrations of HCl-extractable Fe(II) in all treatments became steady after 3 days (Figs. 2 and 3). In addition, there was no significant difference between most of the treatments in the concentrations of dissolved and HCl-extractable Fe(II) except those in certain CO(NH2 )2 treatments. The concentrations of

Table 1 Transformation rates of pentachlorophenol (PCP) based on a logistic degradation curve (x = K/(1 + ea – rt ) under different concentrations of NH4 Cl or CO(NH2 )2 . NH4 Cl (mM)

Ka

Rb (d−1 )

R2

m c (d−1 )

0 2.5 5 10 20 30

1.03 1.00 1.02 1.04 0.65 0.19

0.98 1.47 2.39 0.79 0.84 0.77

0.989 0.999 0.984 0.998 0.994 0.960

0.25 0.37 0.61 0.20 0.14 0.036

a b c

The maximum degradation fraction. The transformation rate constant. The maximum transformation rate (m = Kr/4).

K

r (d−1 )

R2

m (d−1 )

0

1.03

0.98

0.989

0.25

5 10 15 20

1.03 1.01 1.09 0.21

1.28 1.89 0.53 0.79

0.993 0.998 0.998 0.974

0.33 0.48 0.14 0.041

CO(NH2 )2 (mM)

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dissolved Fe(II) in the 5 mM CO(NH2 )2 treatment were higher than those in the 15 mM and 20 mM CO(NH2 )2 treatments (p < 0.05) (Fig. 3A). In addition, the concentrations of HCl-extractable Fe(II) in the 10 mM CO(NH2 )2 treatments were significantly higher than those in the 20 mM CO(NH2 )2 treatments (p < 0.05) (Fig. 3B).

.20

Dissolved Fe(II) (mM)

(A) .15

3.3. Microbial community structure

.10

.05

0.00 .8

HCl extractable Fe(II) (mM)

(B) .6

.4 0 mM NH4Cl 2.5 mM NH4Cl 5 mM NH4Cl

.2

10 mM NH4Cl 20 mM NH4Cl 30 mM NH4Cl

0.0 0

3

6

9

12

Reaction time (day) Fig. 2. The variation in the concentrations of dissolved and HCl-extractable Fe(II) in different NH4 Cl treatments with incubation time.

.16

Dissolved Fe(II) (mM)

(A) .12

.08

.04

0.00 .8

(B)

HCl extractable Fe(II) (mM)

13

.6

.4

0 mM CO(NH2)2 5 mM CO(NH2)2 10 mM CO(NH2)2

.2

15 mM CO(NH2)2 20 mM CO(NH2)2

0.0 0

3

6

9

12

Reaction time (day) Fig. 3. The variation in the concentrations of dissolved and HCl-extractable Fe(II) in different CO(NH2 )2 treatments with incubation time.

Samples incubated for 0 (control only), 5, 7 and 12 days in all NH4 Cl and CO(NH2 )2 treatments were chosen to examine the variation in microbial community structure with a T-RFLP analysis targeting the bacterial 16S rRNA genes. Approximately 20 T-RFs per treatment were detected, of which T-RFs of 69, 151, 154, 175, 213, 217, 235, 248, 253 and 443 bp were detected in all NH4 Cl treatments (Fig. S1). For all the CO(NH2 )2 treatments, T-RFs of 69, 154, 175, 213, 229, 235, 248, 253 and 443 bp were detected (Fig. S2). The phylogenetic associations of the dominant T-RFs were evaluated by constructing clone libraries from bacterial 16S rRNA genes. A total of 20 dominant T-RFs and their corresponding putative phylogenetic associations are presented in Table 2. One of the most dominant clones identified in all of the enrichments showed sequence similarity to the Comamonas genus (Figs. 4 and 5). The Comamonas genus is capable of Fe(III) and humic substance reduction coupled to the reductive dechlorination of chlorinated organic compounds under anaerobic conditions [29]. The abundance of Comamonas in the control incubated at 5 days and 7 days was higher than that in the NH4 Cl/CO(NH2 )2 treatments (p < 0.05) (Figs. 4 and 5), which indicated that NH4 Cl/CO(NH2 )2 tended to inhibit the growth of Comamonas. In addition, the abundance of Comamonas in the treatments with 2.5 mM, 5 mM and 20 mM NH4 Cl addition showed a decreasing trend at 12 days (p < 0.05), whereas the abundance increased slightly in the 30 mM NH4 Cl treatments (p < 0.05) (Fig. 4). Integrating these results with the variations in PCP transformation and in the concentration of dissolved Fe(II) (Figs. 1 and 2Figs. 1A and 2A), we speculate that Comamonas contributed substantially to Fe(III) reduction coupled to the reductive dechlorination of chlorinated organic compounds during the first 7 days. After 7 days, the concentrations of Fe(II) nearly reached a stable level (Fig. 2). Therefore, Comamonas reduced PCP directly during this period of time. Nevertheless, the reductive dechlorination ability of the microorganism is relatively weak [29]; therefore, the transformation rate of PCP in the 30 mM NH4 Cl treatment was still lower (Fig. 1), although higher abundances of Comamonas were found at 7 days and 12 days (p < 0.05) (Fig. 4). In contrast, in the control (the CO(NH2 )2 treatment), the abundances of Comamonas increased initially and then decreased (p < 0.05) (Fig. 5), whereas the concentration of dissolved Fe(II) increased continuously after 7 days (Fig. 3A). These results suggested that Fe(III) reduction coupled to reductive dechlorination was primarily performed by Comamonas during the first 7 days, and Comamonas primarily tended to reduce Fe(III) after 7 days in the control. However, no consistent pattern of variation was found between the lower abundance of Comamonas, PCP transformation and concentration of Fe(II) in the CO(NH2 )2 treatments (Figs. 1, 3 and 5), indicating a small contribution to the reductive dechlorination of PCP or Fe(III) reduction for the microorganism in the presence of CO(NH2 )2 . Other dominant clones were associated with the genus Clostridium sensu stricto (Figs. 4 and 5), and it has been reported that Clostridium sensu stricto is commonly found in the soil under anaerobic conditions [30]. A previous study [31] has demonstrated that many Clostridium species are tolerant of chlorinated solvents and can produce hydrogen in the presence of chlorinated solvents, but none of the Clostridium species isolated or obtained from culture collections in that study could reduce chlorinated solvents [31]. In contrast, many bacteria capable of reductive dechlorination require an exogenous supply of hydrogen as an electron donor

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Table 2 Predicted genus associations for dominant terminal restriction fragment (T-RF) lengths based on bacterial 16S rRNA gene sequences. Peak number

T-RF length (bp)

Predicted genus association

Accession number

Similarity (%)

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19

55 69 73 151 154 160 175 217 219 229 235 237 243 245 248 253 435 440 443

Desulfosporosinus Sedimentibacter Acetivibrio unclassified Lachnospiraceae Achromobacter Burkholderiaceae Ralstonia unclassified Veillonellaceae unclassified Veillonellaceae Clostridium sensu stricto Comamonas Clostridium sensu stricto Clostridium sensu stricto unclassified Veillonellaceae unclassified Veillonellaceae Clostridium sensu stricto Thiobacillus Clostridium sensu stricto Bacillus

AJ582756 AY673993 AB240338 JX222919 FJ195770 DQ227340 GQ417778 HQ660788 FJ799146 AJ229197 EU515237 JX133664 JX133664 HQ660792 HQ660792 FN667108 EU685841 DQ129397 U20384

97 96 95 95 95 99 98 99 98 99 98 98 98 99 99 96 100 98 96

Fig. 4. Microbial community variation in different NH4 Cl treatments with incubation time based on terminal restriction fragment length polymorphism (T-RFLP) analysis and clone library construction.

H.-Y. Yu et al. / Journal of Hazardous Materials 272 (2014) 10–19

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Fig. 5. Microbial community variation in different CO(NH2 )2 treatments with incubation time based on terminal restriction fragment length polymorphism (T-RFLP) analysis and clone library construction.

[32,33]. Therefore, Clostridum species may functionally support the dechlorination of chlorinated organic compounds through intraspecies hydrogen transfer. Furthermore, several other important clones (the genus Sedimentibacter) have been reported to assist the genus Dehalobacter in completing reductive dechlorination under anaerobic conditions [34]. Lastly, the genus Bacillus, fermentative bacteria capable of reductive dechlorination under anaerobic conditions [35], were also detected in most of the treatments. Total bacteria, dechlorinating bacteria in the genus Dehalobacter and iron-reducing bacteria in the genera Shewanella and Geobacter were quantified using quantitative real-time PCR (Figs. 6 and 7). The Geobacter genus was not detected in the present study. The copy numbers of total bacteria in all the treatments except 20 mM CO(NH2 )2 treatment all first showed an increasing trend and then decreased (Figs. 6 and 7Figs. 6A and 7A). The increasing trend may be attributable to NH4 Cl/CO(NH2 )2 addition as a nitrogen source. The copy numbers of the genus Shewanella in the control first increased, then decreased and finally increased again, whereas

they showed an increasing trend in the 30 mM NH4 Cl treatment (Fig. 6B). In contrast, the copy numbers of the genus Shewanella in all CO(NH2 )2 treatments were higher than those in the control (p < 0.05), and they first increased, then decreased and finally slowly increased again (Fig. 7B), indicating the enhanced growth of the genus Shewanella derived from CO(NH2 )2 addition. Moreover, the copy numbers of the genus Shewanella in the 5 mM CO(NH2 )2 treatment were higher than those in the other CO(NH2 )2 treatments (p < 0.05) (Fig. 7B), consistent with the variation in the concentration of dissolved Fe(II) (Fig. 3A). The abundance of Dehalobacter varied significantly between different treatments, ranging from 103 to 108 copies/g soil (Figs. 6 and 7Figs. 6C and 7C), suggesting a substantial impact of NH4 Cl/CO(NH2 )2 addition on the growth of Dehalobacter. Generally, the copy numbers of Dehalobacter in the treatments with higher concentrations of NH4 Cl/CO(NH2 )2 were lower than those in the treatments with lower concentrations of NH4 Cl/CO(NH2 )2 . For example, the copy numbers of the Dehalobacter genus in the 5 mM CO(NH2 )2 treatment were

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H.-Y. Yu et al. / Journal of Hazardous Materials 272 (2014) 10–19

Fig. 6. DNA copy numbers of the 16S rRNA genes of total bacteria (A), Shewanella spp. (B) and Dehalobacter spp. (C) in different NH4 Cl treatments with incubation time.

1 to 4 orders of magnitude higher than those in the control; in contrast, in the 20 mM CO(NH2 )2 treatment, they were 2 orders of magnitude lower than those in the control (Fig. 7C). These results indicated that the addition of low concentrations of NH4 Cl/CO(NH2 )2 can enhance the growth of Dehalobacter,

whereas the addition of high concentrations of NH4 Cl/CO(NH2 )2 can cause the opposite effect. This outcome is consistent with the conclusion that PCP degradation is enhanced by low concentrations of NH4 Cl/CO(NH2 )2 and inhibited by high concentrations of NH4 Cl/CO(NH2 )2 . In the 2.5 mM NH4 Cl treatment, the copy

Fig. 7. DNA copy numbers of the 16S rRNA genes of total bacteria (A), Shewanella spp. (B) and Dehalobacter spp. (C) in different CO(NH2 )2 treatments with incubation time.

H.-Y. Yu et al. / Journal of Hazardous Materials 272 (2014) 10–19

numbers of the Dehalobacter genus increased rapidly at 5 days and then decreased; in contrast, they increased markedly at 7 days and then decreased in the 5 mM NH4 Cl treatment but were higher than those in the 2.5 mM NH4 Cl treatment (p < 0.05) (Fig. 6C). These results were consistent with the higher transformation rate of PCP in the 5 mM NH4 Cl treatment compared to that in the 2.5 mM NH4 Cl treatment (Fig. 1A). In addition, the copy numbers of Dehalobacter in the control, higher than 105 , increased markedly at 12 days (Fig. 6C). This result is also consistent with the higher transformation rate of PCP in this treatment (Fig. 1A). Thus, Dehalobacter may play a role in the process of PCP dechlorination.

4. Discussion Nearly all the CO(NH2 )2 added was rapidly and completely hydrolyzed into NH4 + within three days, and no variation was found in the concentration of NH4 + in the NH4 Cl treatments with increasing incubation time (Fig. S3), therefore, the additions of NH4 Cl and CO(NH2 )2 showed similar effects on the transformation of PCP. In particular, the addition of low concentrations of NH4 Cl or CO(NH2 )2 promoted PCP dechlorination, whereas the application of high concentrations of NH4 Cl or CO(NH2 )2 inhibited PCP dechlorination. Similar results have been obtained by a previous study [36]. In that study, two reasonable explanations for the enhanced effect were suggested by the authors. First, the soil C/N ratio can be adjusted to a more suitable level for the degraders’ activity by the addition of a proper amount of nitrogen. Second, NH4 + –N, as a reductant, can be an electron donor and can thus, accelerate the process of reductive dechlorination. In contrast, high concentrations of NH4 Cl or CO(NH2 )2 showed an inhibitory effect, most likely because excessive N addition could alter the enzymatic systems of the soil responsible for the degradation [12,37] and thus inhibit the specific degraders. All these results are also consistent with the results of the analysis of microbial community structure in the present study. Dehalobacter, recognized as a genus of strictly organohaliderespiring anaerobes, can use many chlorinated organic compounds as electron acceptors. These compounds include chlorinated ethenes [38], chlorinated ethanes [39], ␤-hexachlorocyclohexane [40], dichlorobenzenes and monochlorobenzene [41] and chlorophenol [42]. The effects of NH4 Cl or CO(NH2 )2 addition on the growth of Dehalobacter showed trends similar to those for PCP dechlorination, i.e., enhanced effects from low concentrations of NH4 Cl or CO(NH2 )2 and inhibitory effects corresponding to high concentrations of NH4 Cl or CO(NH2 )2 , suggesting the involvement of Dehalobacter in PCP dechlorination. This result also confirmed the importance of Dehalobacter species in bioremediation as dedicated degraders of chlorinated organic compounds, as recognized by previous studies [43,44]. Although three Dehalobacter strains have been isolated and physiologically characterized [38,39,45], the isolated Dehalobacter strains have relatively slow growth rates in pure culture. Consequently, in the environment, it is very common for Dehalobacter species to show synergistic dechlorination with their co-culture or enrichment culture companions such as Sedimentibacter, Clostridium and Acetobacterium [34]. In the present study, the genera Sedimentibacter and Clostridium were all detected (Figs. 4 and 5). Metagenome analysis [34] has revealed that the pathway for producing vitamin B12, essential for reductive dehalogenase activity in organohalide respiring bacteria, is incomplete in Dehalobacter species, whereas a complete cobalamin synthesis pathway has been identified in Sedimentibacter. Therefore, Sedimentibacter may supply cobalamin to Dehalobacter species or convert cyano-cobalamin to adenosyl-cobalamin (or other corrinoids), available forms of cobalamin for Dehalobacter species. In addition, Sedimentibacter may also be involved in

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releasing simple carbon sources for Dehalobacter. Overall, coculture or enrichment culture companions of Dehalobacter may serve syntrophic roles in the process of dechlorination. The results for the concentrations of dissolved and HClextractable Fe(II) showed that no significant difference was found between most of the different NH4 Cl or CO(NH2 )2 treatments (Figs. 2 and 3). The Fe(II) species, such as sorbed Fe(II), can abiotically donate electrons to reduce chlorinated contaminants [46]. In the present study, the concentration of sorbed Fe(II) can be obtained by subtracting dissolved Fe(II) from HCl-extractable Fe(II) (Fig. S4). However, no consistent patterns of variation for the concentrations of sorbed Fe(II) and PCP transformation were found, indicating a small contribution of sorbed Fe(II) to PCP dechlorination. Nevertheless, the rapidly increasing trends in the concentrations of dissolved and HCl-extractable Fe(II) in all treatments during the first 3 days suggested the occurrence of iron reduction induced by dissimilatory iron-reducing bacteria (Figs. 2 and 3). In the present study, the genera Comamonas and Shewanella were the two dominant iron-reducing bacteria detected. Comamonas is capable of Fe(III) reduction coupled to reductive dechlorination and can also reduce chlorinated organic compounds directly (chlororespiration) [29]. However, the chlororespiration ability of Comamonas may be very weak [29]. In fact, many dissimilatory iron-reducing bacteria have been found to be capable of chlororespiration [17]. Furthermore, both halorespiring microorganisms and dissimilatory Fe(III) reducers may be more widespread than currently recognized, as evaluation of the capacity for dechlorination has not been part of the characterization of many Fe(III) reducers [17]. In addition, previous studies [18,19] have indicated that Fe(III) reduction and dechlorination may occur simultaneously in contaminated environments. In the present study, based on the variation in the abundance of Comamonas and the patterns of variation of PCP transformation and the concentration of Fe(II), Comamonas may play different roles overall at different reaction stages. Moreover, both NH4 Cl and CO(NH2 )2 , especially the latter, showed inhibitory effects on the growth of Comamonas. In addition, it appeared that there was no difference among the different concentrations of NH4 Cl or CO(NH2 )2 in terms of the resulting inhibitory effects. Shewanella has been the most intensively studied Fe(III)-reducing microorganism and can be found in a diversity of environments [17]. Similar to Comamonas, Shewanella can be capable of both Fe(III) reduction and chlororespiration, and its chlororespiration ability may also be weak [47]. Nevertheless, there is no evidence that Shewanella is prevalent and contributes substantially to iron reduction in the environments in which Fe(III) reduction is important [17]. In the present study, the addition of NH4 Cl or CO(NH2 )2 showed varying effects on the growth of Shewanella. In particular, low concentrations of NH4 Cl (2.5 mM, 5 mM and 20 mM) appeared to inhibit Shewanella’s growth, whereas high concentrations of NH4 Cl (30 mM) enhanced its growth. In contrast, all concentrations of CO(NH2 )2 showed enhanced effects. Note that all the CO(NH2 )2 -N provided was transformed to NH4 + –N in the first 3 days. Therefore, the difference was the formation of CO2 in the CO(NH2 )2 treatments compared with the NH4 Cl treatments. The effect of CO2 on the growth of Shewanella has seldom been examined. A previous study [48] has demonstrated that a mixture of 1% CO2 with N2 at total pressures of 15 or 150 psi significantly suppressed the growth of Shewanella oneidensis MR-1 (a model Shewanella strain) compared to the N2 control; at CO2 partial pressures over 15 psi, the growth of MR-1 stopped. Furthermore, after exposure to 150 psi CO2 for 5 h, no viable cells survived. All these results indicated that CO2 had a significant effect on the growth of Shewanella. Nevertheless, that study [48] also reported that MR-1 could survive better if pH-buffering minerals were present under CO2 pressure. In the present study, CO2 hydrolyzed from CO(NH2 )2 appeared to enhance the growth of Shewanella

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H.-Y. Yu et al. / Journal of Hazardous Materials 272 (2014) 10–19

relative to the control CO(NH2 )2 treatments (Fig. 7B). This result may be attributable to the pH-buffering PIPES addition. In conclusion, NH4 Cl or CO(NH2 )2 addition produced variations in soil microbial community structure, especially in the abundances of dechlorinating bacteria and dissimilatory iron-reducing bacteria, which in turn affected PCP dechlorination and iron reduction. Acknowledgments The current study was financially supported by the National Natural Science Foundation of China (nos. 41025003, 41201505, and 40901114). Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at http://dx.doi.org/10.1016/j.jhazmat. 2014.02.037. References [1] R. Eisler, in: Pentachlorophenol Hazards to Fish, Wildlife, and Invertebrates: A Synoptic Review, U.S. Fish and Wildlife Service, Laurel, 1989. [2] B.I. Escher, M. Snozzi, R.P. Schwarzenbach, Uptake, speciation, and uncoupling activity of substituted phenols in energy transducing membranes, Environ. Sci. Technol. 30 (1996) 3071–3079. [3] P.L. Williams, Pentachlorophenol, an assessment of the occupational hazard, Am. Ind. Hyg. Assoc. J. 43 (1982) 799–810. [4] A. Bevenue, H. Beckman, Pentachlorophenol: a discussion of its properties and its occurrence as a residue in human and animal tissues, Residue Rev. 19 (1967) 83–134. [5] C. Zhang, H. Xu, Q. Jiang, J. Wang, Microbial degradation and photolysis of pentachlorophenol, Chin. J. Ecol. 16 (1997) 19–22. [6] W. Zheng, X. Wang, H. Yu, X. Tao, Y. Zhou, W. Qu, Global trends and diversity in pentachlorophenol levels in the environment and in humans: a meta-analysis, Environ. Sci. Technol. 45 (2011) 4668–4675. [7] S. Peng, J. Huang, X. Zhong, J. Yang, G. Wang, Y. Zhou, F. Zhang, Q. Zhu, B. Roland, W. Christian, Reseach strategy in improving fertilizer nitrogen use efficiency of irrigated rice in China, Chin. Agric. Sci. 35 (2002) 1095–1103. [8] National Bureau of Statistics of China, in: China Statistical Yearbook, China Statistics Press, 2012, http://www.stats.gov.cn/tjsj/ndsj/2012/indexch.htm (accessed 8 October 2013). [9] Guangdong Provincial Bureau of Statistics, in: Guangdong Statistical Yearbook, Guangdong Provincial Bureau of Statistics, 2012, http://www.gdstats.gov.cn/ tjnj/2012/table/11/11-03.htm (accessed 8 October 2013) The National Bureau of Statistics Survey Office of Guangdong. [10] S.U. Sarathchandra, A. Ghani, G.W. Yeates, G. Burch, N.R. Cox, Effect of nitrogen and phosphate fertilisers on microbial and nematode diversity in pasture soils, Soil Biol. Biochem. 33 (2001) 953–964. [11] J.F. Braddock, M.L. Ruth, P.H. Catterall, J.L. Walworth, K.A. Mccarthy, Enhancement and inhibition of microbial activity in hydrocarbon-contaminated arctic soils: Implications for nutrient-amended bioremediation, Environ. Sci. Technol. 31 (1997) 2078–2084. [12] R. Abdelhafid, S. Houot, E. Barriuso, How increasing availabilities of carbon and nitrogen affect atrazine behaviour in soils, Biol. Fert. Soils 30 (2000) 333–340. [13] L. Xun, Microbial degradation of polychlorophenols, in: S.N. Singh (Ed.), Microbial Degradation of Xenobiotics, Springer-Verlag, Berlin, Heidelberg, Pullman, 2012, pp. 1–30. [14] L. Adrian, H. Gorisch, Microbial transformation of chlorinated benzenes under anaerobic conditions, Res. Microbiol. 153 (2002) 131–137. [15] E.I. Atuanya, H.J. Purohit, T. Chakrabarti, Anaerobic and aerobic biodegradation of chlorophenols using UASB and ASG bioreactors, World J. Microb. Biot. 16 (2000) 95–98. [16] F.E. Löffler, J.R. Cole, K.M. Ritalahti, J.M. Tiedje, Diversity of dechlorinating bacteria, in: M.M. Häggblom, I.D. Bossert (Eds.), Dehalogenation: Microbial Processes and Environmental Applications, Kluwer Academic Publishers, Atlanta, GA, 2003, pp. 53–87. [17] D.R. Lovley, D.E. Holmes, K.P. Nevin, Dissimilatory Fe(III) and Mn(IV) reduction, in: R.K. Poole (Ed.), Advances in Microbial Physiology, Elsevier, Amherst, 2004, pp. 219–286. [18] Y. Sung, K.M. Ritalahti, R.A. Sanford, J.W. Urbance, S.J. Flynn, J.M. Tiedje, F.E. Loffler, Characterization of two tetrachloroethene-reducing, acetate-oxidizing anaerobic bacteria and their description as Desulfuromonas michiganensis sp nov, Appl. Environ. Microbiol. 69 (2003) 2964–2974. [19] Q. He, R.A. Sanford, Characterization of Fe(III) reduction by chlororespiring Anaeromxyobacter dehalogenans, Appl. Environ. Microbiol. 69 (2003) 2712–2718.

[20] K.T. Finneran, H.M. Forbush, C.V.G. VanPraagh, D.R. Lovley, Desulfitobacterium metallireducens sp. nov., an anaerobic bacterium that couples growth to the reduction of metals and humic acids as well as chlorinated compounds, Int. J. Syst. Evol. Microbiol. 52 (2002) 1929–1935. [21] N. Wei, K.T. Finneran, Influence of ferric iron on complete dechlorination of trichloroethylene (TCE) to ethene: Fe(III) reduction does not always inhibit complete dechlorination, Environ. Sci. Technol. 45 (2011) 7422–7430. [22] M. Chen, K. Shih, M. Hu, F. Li, C. Liu, W. Wu, H. Tong, Biostimulation of indigenous microbial communities for anaerobic fransformation of pentachlorophenol in paddy soils of Southern China, J. Agric. Food Chem. 60 (2012) 2967–2975. [23] F.B. Li, X.G. Wang, C.S. Liu, Y.T. Li, F. Zeng, L. Liu, Reductive transformation of pentachlorophenol on the interface of subtropical soil colloids and water, Geoderma 148 (2008) 70–78. [24] A.P. Khodadoust, M.T. Suidan, C.M. Acheson, R.C. Brenner, Solvent extraction of pentachlorophenol from contaminated soils using water–ethanol mixtures, Chemosphere 38 (1999) 2681–2693. [25] Q. Lan, F.B. Li, C.S. Liu, X.Z. Li, Heterogeneous photodegradation of pentachlorophenol with maghemite and oxalate under UV illumination, Environ. Sci. Technol. 42 (2008) 7918–7923. [26] E. Viollier, P.W. Inglett, K. Hunter, A.N. Roychoudhury, P. Van Cappellen, The ferrozine method revisited: Fe(II)/Fe(III) determination in natural waters, Appl. Geochem. 15 (2000) 785–790. [27] H.-Y. Yu, Y.-K. Wang, P.-C. Chen, F.-B. Li, M.-J. Chen, M. Hu, Effect of nitrate addition on reductive transformation of pentachlorophenol in paddy soil in relation to iron(III) reduction, Biol. Fert. Soils (2013), Under review. [28] J.A. Whelan, N.B. Russell, M.A. Whelan, A method for the absolute quantification of cDNA using real-time PCR, J. Immunol. Methods 278 (2003) 261–269. [29] Y.B. Wang, C.Y. Wu, X.J. Wang, S.G. Zhou, The role of humic substances in the anaerobic reductive dechlorination of 2,4-dichlorophenoxyacetic acid by Comamonas koreensis strain CY01, J. Hazard. Mater. 164 (2009) 941–947. [30] U. Hengstmann, K.J. Chin, P.H. Janssen, W. Liesack, Comparative phylogenetic assignment of environmental sequences of genes encoding 16S rRNA and numerically abundant culturable bacteria from an anoxic rice paddy soil, Appl. Environ. Microbiol. 65 (1999) 5050–5058. [31] K. Bowman, Bacterial characterization of louisinan groudwater contaminated by DNAPL-containing chloroethanes and other solvents, in: Ph. D. Dissertation, Louisiana State University, Baton Rouge, 2009. [32] B.S. Ballapragada, H.D. Stensel, J.A. Puhakka, J.F. Ferguson, Effect of hydrogen on reductive dechlorination of chlorinated ethenes, Environ. Sci. Technol. 31 (1997) 1728–1734. [33] X. Maymó-Gatell, Y.T. Chien, J.M. Gossett, S.H. Zinder, Isolation of a bacterium that reductively dechlorinates tetrachloroethene to ethene, Science 276 (1997) 1568–1571. [34] F. Maphosa, M.W.J. van Passel, W.M. de Vos, H. Smidt, Metagenome analysis reveals yet unexplored reductive dechlorinating potential of Dehalobacter sp. E1 growing in co-culture with Sedimentibacter sp, Env. Microbiol. Rep. 4 (2012) 604–616. [35] D.R. Boone, Y.T. Liu, Z.J. Zhao, D.L. Balkwill, G.R. Drake, T.O. Stevens, H.C. Aldrich, Bacillus infernus sp.nov, an Fe(III)- and Mn(IV)-reducing anaerobe from the deep terrestrial subsurface, Int. J. Syst. Bacteriol. 45 (1995) 441–448. [36] C. Liu, X. Jiang, F. Wang, X. Yang, T. Wang, Hexachlorobenzene dechlorination as affected by nitrogen application in acidic paddy soil, J. Hazard. Mater. 179 (2010) 709–714. [37] J.A. Entry, K.G. Mattson, W.H. Emmingham, The influence of nitrogen on atrazine and 2,4-dichlorophenoxyacetic acid mineralization in grassland soils, Biol. Fert. Soils 16 (1993) 179–182. [38] C. Holliger, D. Hahn, H. Harmsen, W. Ludwig, W. Schumacher, B. Tindall, F. Vazquez, N. Weiss, A.J.B. Zehnder, Dehalobacter restrictus gen. nov. and sp. nov., a strictly anaerobic bacterium that reductively dechlorinates tetraand trichloroethene in an anaerobic respiration, Arch. Microbiol. 169 (1998) 313–321. [39] B.L. Sun, B.M. Griffin, H.L. Ayala-del-Rio, S.A. Hashsham, J.M. Tiedje, Microbial dehalorespiration with 1,1,1-trichloroethane, Science 298 (2002) 1023–1025. [40] W. van Doesburg, M.H.A. van Eekert, P.J.M. Middeldorp, M. Balk, G. Schraa, A.J.M. Stams, Reductive dechlorination of beta-hexachlorocyclohexane (betaHCH) by a Dehalobacter species in coculture with a Sedimentibacter sp, FEMS Microbiol. Ecol. 54 (2005) 87–95. [41] J.L. Nelson, J.M. Fung, H. Cadillo-Quiroz, X. Cheng, S.H. Zinder, A Role for Dehalobacter spp. in the reductive dehalogenation of dichlorobenzenes and monochlorobenzene, Environ. Sci. Technol. 45 (2011) 6806–6813. [42] C.F. Zhang, D. Suzuki, Z.L. Li, L.Z. Ye, A. Katayama, Polyphasic characterization of two microbial consortia with wide dechlorination spectra for chlorophenols, J. Biosci. Bioeng. 114 (2012) 512–517. [43] A. Grostern, E.A. Edwards, Growth of Dehalobacter and Dehalococcoides spp. during degradation of chlorinated ethanes, Appl. Environ. Microbiol. 72 (2006) 428–436. [44] A. Grostern, E.A. Edwards, Characterization of a Dehalobacter coculture that dechlorinates 1,2-dichloroethane to ethene and identification of the putative reductive dehalogenase Ge1e, Appl. Environ. Microbiol. 75 (2009) 2684–2693. [45] A. Wild, R. Hermann, T. Leisinger, Isolation of an anaerobic bacterium which reductively dechlorinates tetrachloroethene and trichloroethene, Biodegradation 7 (1996) 507–511.

H.-Y. Yu et al. / Journal of Hazardous Materials 272 (2014) 10–19 [46] S. Kim, F.W. Picardal, Enhanced anaerobic biotransformation of carbon tetrachloride in the presence of reduced iron oxides, Environ. Toxicol. Chem. 18 (1999) 2142–2150. [47] F.B. Li, X.M. Li, S.G. Zhou, L. Zhuang, F. Cao, D.Y. Huang, W. Xu, T.X. Liu, C.H. Feng, Enhanced reductive dechlorination of DDT in an anaerobic system of dissimilatory iron-reducing bacteria and iron oxide, Environ. Pollut. 158 (2010) 1733–1740.

19

[48] B. Wu, H.B. Shao, Z.P. Wang, Y.D. Hu, Y.J.J. Tang, Y.S. Jun, Viability and metal reduction of Shewanella oneidensis MR-1 under CO2 stress: implications for ecological effects of CO2 leakage from geologic CO2 sequestration, Environ. Sci. Technol. 44 (2010) 9213–9218.