Nitrogen removal by the enhanced floating treatment wetlands from the secondary effluent

Bioresource Technology 234 (2017) 243–252

Contents lists available at ScienceDirect

Bioresource Technology journal homepage: www.elsevier.com/locate/biortech

Nitrogen removal by the enhanced floating treatment wetlands from the secondary effluent Lei Gao, Weili Zhou, Jungchen Huang, Shengbing He ⇑, Yijia Yan, Wenying Zhu, Suqing Wu, Xu Zhang School of Environmental Science and Engineering, Shanghai Jiaotong University, Shanghai 200240, PR China

h i g h l i g h t s
Thiosulfate-driven autotrophic denitrification was firstly introduced into EFTW.
AEFTW and HEFTW could effectively remove nitrogen from the secondary effluent.
Addition of electron donors reduced the N2O emission.
The electron donors induced distinct shifts in microbial structures.
Mixotrophic denitrification occurred in HEFTW and AEFTW.

a r t i c l e

i n f o

Article history: Received 18 January 2017 Received in revised form 2 March 2017 Accepted 5 March 2017 Available online 9 March 2017 Keywords: Secondary effluent Enhanced floating treatment wetland Autotrophic and heterotrophic denitrification N2O emission Microbial community

a b s t r a c t Three novel floating treatment wetlands, including autotrophic enhanced floating treatment wetland (AEFTW), heterotrophic enhanced floating treatment wetland (HEFTW) and enhanced floating treatment wetland (EFTW) were developed to remove nitrogen from the secondary effluent. Results showed that the analogously excellent nitrogen removal performance was achieved in AEFTW and HEFTW. About 89.4% of the total nitrogen (TN) was removed from AEFTW at a low S/N of 0.9 and 88.5% from HEFTW at a low C/N of 3.5 when the hydraulic retention time (HRT) was 1 d in summer. Higher nitrification and denitrification performance were achieved in AEFTW. Addition of electron donors effectively reduced the N2O emission, especially in summer and autumn. High-throughput sequencing analysis revealed that the electron donors distinctly induced the microbial shifts. Dechloromonas, Thiobacillus and Nitrospira became the most predominant genus in HEFTW, AEFTW and EFTW. And autotrophic and heterotrophic denitrification could simultaneously occur in HEFTW and AEFTW. Ó 2017 Elsevier Ltd. All rights reserved.

1. Introduction The excessive N, mainly from fertilizer use, crop N fixation and industrial activities, causes negative environmental problems such as eutrophication of terrestrial and aquatic systems, global acidification (Gruber and Galloway, 2008). China now faces serious problems of surface water contamination. 11.5% of major freshwater lakes and reservoirs in China were rated as severely polluted and 26.2% as slightly polluted (Li et al., 2015). The secondary effluent from the municipal sewage treatment plants entering the surface water will cause water quality deterioration and eutrophication of rivers and lakes. As the total nitrogen (TN) concentration (12–15 mg L1) in the secondary effluent is still much higher than the surface water criteria (2.0 mg L1 for grade V in Chinese National Surface Water Environmental Quality Standard ⇑ Corresponding author. E-mail addresses: [email protected], [email protected] (S. He). http://dx.doi.org/10.1016/j.biortech.2017.03.036 0960-8524/Ó 2017 Elsevier Ltd. All rights reserved.

(GB3838-2002, 2002)). Therefore, improving TN removal from secondary effluent has become an urgent need all over China. The floating treatment wetlands (FTWs), as an innovative variant of constructed wetlands, has been well applied for water quality improvement of stormwater, polluted river water, eutrophic water and secondary effluent due to its merits of low cost, occupying no additional land area and good suitability for either new construction or retrofit installation on the existing ponds and lakes (Borne et al., 2014; Karine et al., 2013). The enhanced FTWs (EFTWs), composed of the bio-carriers and emergent plants, providing additional substrate surface for the attachment of microorganisms, has been introduced to improve nutrient removal efficiency (Wu et al., 2016; Zhang et al., 2015). However, the nitrate removal performance of EFTWs is still unsatisfactory in treating the secondary effluent, mainly due to the following reasons: 1) the treatment performance is limited by the growth rate of hydrophytes, especially at low temperature; 2) the secondary effluent is lack of biodegradable organic carbon and the endoge-

244

L. Gao et al. / Bioresource Technology 234 (2017) 243–252

Nomenclature FTW EFTW HEFTW AEFTW AD TN DO

TP CW HD HRT COD RDA

floating treatment wetland enhanced floating treatment wetland heterotrophic enhanced floating treatment wetland autotrophic enhanced floating treatment wetland autotrophic denitrification total nitrogen dissolved oxygen

nous organic carbon in EFTWs from plant exudates, plant litter and microbial decomposition, which does not meet the need of complete denitrification. Therefore, when treating the secondary effluent, the external electron donors are required. Organic electron donors, such as glucose, sodium acetate, methanol and ethanol are the most frequently used. However, they may cause the secondary contamination, and add an extra operation cost on sludge disposal (Zhu and Getting, 2012). Recently, inorganic electron donors, such as the reductive sulfur compounds, have attracted significant attention, as they require no external organic carbon source and produce lower amount of excess sludge (Chung et al., 2014). In addition, thiosulfate (S2O2 3 ) has exhibited higher bioavailability and better denitrification rate, compared with elemental sulfur and sulfide (Sahinkaya and Dursun, 2015; Zhou et al., 2016). Thiosulfate-driven AD occurs according to the stoichiometric equation (Mora et al., 2014):   þ S2 O2 3 þ 1:16NO3 þ 0:035CO2 þ 0:519HCO3 þ 0:110NH4

þ 0:124H2 ! 0:110C 5 H7 O2 N þ 0:578N2 þ 0:435Hþ þ 2SO2 4 ð1Þ To our knowledge, so far there is no case introducing thiosulfate-driven AD into EFTWs. Nitrous oxide (N2O) is a powerful greenhouse gas with 298 times global warming potential of CO2 and is now increasing globally at a rate of 0.2–0.3% per year (IPCC, 2007; Ravishankara et al., 2009). N2O emissions have been extensively studied in riparian wetlands and constructed wetlands (García-Lledó et al., 2011; Mander et al., 2015), but rarely in EFTWs. For EFTWs, microbial activities play a vital role in N removal. The microbial community structure provides an approach to reveal the biological nitrogen removal mechanism in EFTWs. Different electron donors may influence the microbial composition, leading to different macroscopic nitrogen removal efficiency. However, detailed information of functional microorganisms, microbial composition and variation in EFTWs has not yet been reported. In this study, three EFTWs, including AEFTW, HEFTW and EFTW, were introduced to remove nitrogen from the secondary effluent. All three EFTWs tanks were identical. EFTW served as a control reactor with no addition of external electron donor, while AEFTW was added with thiosulfate and HEFTW with acetate as the electron donor. The objectives of this study are 1) to compare the nitrogen removal performance in AEFTW, HEFTW and EFTW; 2) to evaluate N2O emission in EFTWs and explore the influence of different electron donors on N2O emission; 3) to investigate the microbial community and biological nitrogen removal mechanisms in AEFTW, HEFTW and EFTW. 2. Materials and methods 2.1. Experimental setup and operation The experiment was operated in the botanic garden (121°260 3100 N, 31°20 500 E) of Shanghai Jiao Tong University, City of

total phosphorus constructed wetland heterotrophic denitrification hydraulic retention time chemical oxygen demand redundancy analysis

Shanghai, China. An artificial river flows through the botanic garden. Three identical EFTWs tanks (L  W  H = 0.85 m  0.40 m  0.90 m), each consisting of an inlet zone and a reaction zone (Fig. 1), were installed in the botanic garden near the river. The operational water depth was 0.8 m and the effective volume of reaction zone was 224 L. A stainless steel frame (0.6  0.38 m, L  W) with nine holes was fixed to the tank to support the plants. Hydroponic pots with inner diameter of 9.5 cm were placed in the frame. Iris pseudacorus were transplanted into the pots with a plant density of 18 rhizomes m2. The plants used in the experiment were of similar height and weight. Six strings of spherical polypropylene bio-carriers were hung under the stainless steel frame, every string was made up of five spherical fillers with the diameter of 15 cm and the specific surface area of 380 m2 m3. The simulated secondary effluent was prepared from the river water of botanic garden, supplemented with sodium nitrate (KNO3), calcium nitrate tetrahydrate (Ca(NO3)24H2O), ammonium chloride (NH4Cl) and monosodium phosphate (NaH2PO4). The influent concentration of total nitrogen (TN), nitrate nitrogen + (NO 3 -N), ammonia nitrogen (NH4-N), total phosphorus (TP) and 2 sulfate (SO4 ) were 15.13 ± 1.1 mg N L1, 11.25 ± 1.23 mg N L1, 3.28 ± 0.5 mg N L1, 0.5 ± 0.2 mg P L1 and 77.01 ± 10.16 mg L1, respectively. Moreover, sodium acetate was added into the influent of HEFTW as electron donor for the heterotrophic denitrification while sodium thiosulfate into AEFTW for the autotrophic denitrification. The ratio of external COD/TN (C/N, W/W) was set as 3.5 and 5 and the S2O2 3 /TN (S/N, M/M) 0.9, 1. The EFTWs experiment was operated in a continual flow mode for 330 days from March 15, 2015 to February 1, 2016. 2.2. Water sampling and chemical analysis Water samples were taken once every three days from the three  + EFTWs. COD, TN, NO 3 -N, nitrite (NO2 -N) and NH4-N of water sample were measured according to the standard method (APHA, 2 2005). Sulfate (SO2 4 ) and thiosulphate (S2O3 ) were analyzed by ion chromatography (Metrohom 883, Metrohom, Switzerland). DO, pH and temperature were detected by DO meter (HQ30d, HACH) and pH meter (HQ11d, HACH), respectively. All measurements of each sample were determined in three replicates. 2.3. Sludge sampling and molecular analyses In order to investigate the variation of microbial community composition in the three EFTWs, the sludge samples H1, H2; A1, A2; E1, E2 were collected from plant rhizosphere of HEFTW, AEFTW, EFTW tanks on May 29, 2015 and January 1, 2016, respectively, representing the initial and final microbial sample of experiment, respectively. The bacterial DNA was extracted using QIAamp DNA stool Mini Kit. (QIAGEN, Germany) following the manufacturer’s protocol. The extracted DNA was amplified with the primers 338F (ACTCCTACGGGAGGCAGCAG) and 533R (TTACCGCGGCTGCTGG CAC) in the V3 domain of bacterial 16S rRNA genes. PCR amplifica-

L. Gao et al. / Bioresource Technology 234 (2017) 243–252

245

Fig. 1. Schematic diagram of the enhanced floating treatment wetland.

tion program included initial denaturation at 98 °C for 2 min, followed by 28 cycles of 98 °C for 15 s, annealing at 50 °C for 30 s, and extension at 72 °C for 30 s, with a final elongation at 72 °C for 5 min (Zhou et al., 2017). Finally, High-throughput sequencing was conducted with Illumina Hiseq 2000 by Shanghai Personal Biotechnology Co., Ltd (Shanghai, China). Questionable sequences and chimeric sequences were examined and removed by calling USEARCH (v5.2.236, http://www.drive5.com/usearch/) via the Qiime software (v1.9.0, http://qiime.org/). High-quality sequences were clustered into the operational taxonomic units (OTU) by setting a 97% Sequence similarity using UCLUST, and the OTU with the abundance value less than 0.001% were eliminated (Bokulich et al., 2013). The taxonomic sequences classification was conducted using the Greengenes reference database at an 80% confidence threshold. Rarefaction curves, bacterial community richness (Chao 1 and Ace) and diversity (Shannon and Simpson) were obtained using the Mothur software package 1.9.0.

2.5. Statistical analysis All statistical analyses were conducted with SPSS version 20.0 software (SPSS Inc., Chicago, USA) and were considered significant +  at 0.05 level. The differences in TN, NO 3 -N, NH4-N, NO2 -N, pH, DO in EFTWs were tested by one-way ANOVA with Duncan’s post hoc test. Variables that were not normally distributed were transforming by the Box-cox transformation prior analyses. The Pearson correlation and Spearman rank-order correlation were used to characterize the relationship between temperature (T), HRTs and N in effluents. The relationships between bacterial community and physicochemical variables were analyzed using redundancy analysis (RDA) in Canoco version 4.5 for Windows. 2.6. Calculation of the nitrogen removal efficiency, contribution of the microbial transformation The TN removal efficiency was calculated using the flowing equation:

2.4. Gas sampling and N2O gas emission rate measurement Gas sampling was carried out once a month from April 2015 to January 2016. The static chamber technique was used to estimate the N2O emissions in three different EFTWs (Wu et al., 2009). The closed chamber was made of polymethyl methacrylate with a total volume of 340 L. Battery-driven fans were operated inside the chamber to ensure that the gas was well mixed during sampling. Five gas samples were collected at 0, 30, 60, 90, 120 min after enclosure between 8:00 and 12:00 am. The gas samples were indrawn into 500 mL gas sampling bags (Eler Inc., Shanghai, China) using a gas sampling pump. The N2O concentration was determined using an Agilent 6890 gas chromatograph (Agilent Technologies Inc., USA) with an electron capture detector (lECD) and a Poropak Q column (2 m), using 15 mL min1 pure nitrogen (99.999%) as the carrier gas. The temperatures for the column, injector, and detector were set at 70, 100, and 280 °C, respectively. The N2O flux (mg m2 d1) calculated from the slope of the linear plot of the concentration versus time by taking the chamber’s surface area into account (Uggetti et al., 2012).

TNRE ð%Þ ¼ ðTNinf  TNeff Þ=TNinf  100%

ð2Þ

where TNRE is the TN removal efficiency, TNinf represents the influent TN concentration (mg L1) and TNeff, the effluent TN concentration (mg L1). For the calculation of the contribution of the microbial transformation to total nitrogen removal in HEFTW and AEFTW, the following assumptions were made: 1) The external electron donors only increased the microbial transformation. 2) Nitrogen removal by other routes was assumed to be the same in 3 EFTWs. Therefore, it could be expressed as follows:

Microbial transformation contribution of AEFTW or HEFTW ¼ ðTNREAEFTW or HEFTW  TNREEFTW Þ=TNREAEFTW or HEFTW  100% ð3Þ

246

L. Gao et al. / Bioresource Technology 234 (2017) 243–252

where TNREAEFTW or HEFTW is the TN removal efficiency of AEFTW or HEFTW, TNREEFTW stands for the TN removal efficiency of EFTW.

3. Results and discussion 3.1. Nitrogen removal performance Fig. 2 showed the concentration of different forms of N in the influent and effluent throughout the experiment period. The average effluent TN of HEFTW (2.80 mg L1, b) and AEFTW (3.63 mg L1, b) were significantly lower than that of EFTW (12.04 mg L1, a). During the start-up stage (0–75 days in spring), C/N, S/N were set as 5 and 1 for HEFTW and AEFTW and HRT was 3 days. TN concentration in HEFTW decreased immediately from 5.3 to 2.25 mg L1 in first 30 days and then remained steady at approximately 2 mg L1 till day 220. Differently, effluent of AEFTW showed a gradual decrease from 8.59 to 4.02 mg L1 for 2 months. The phenomenon might be due to the fact that the heterotrophic denitrifying bacteria proliferated faster than the autotrophic one. The optimal effluent TN of HEFTW, AEFTW decreased to 1.23, 0.80 mg/L at low C/N (3.5), S/N (0.9) and HRT (1d) in August while peak nitrogen removal efficiency (39%) of EFTW was obtained in July at the HRT of 3 d. It was believed that the high temperature in summer accelerated the denitrifying rate, promoted the plant growing rate, leading to more uptake of nitrogen, and also promoted the production of endogenous organic matters from root exudates and died plant detritus (Borne et al., 2013). Decrease of TN removal efficiency was not found in HEFTW and

AEFTW when HRT was shortened at the temperature above 20 °C, but indeed existed in EFTW. TN effluents increased to 5.40, 6.22 and 14.6 mg L1 in HEFTW, AEFTW and EFTW, when the temperature dropped to 10 °C and below, The reason probably was that the low temperature slowed down the microbial activity (Xu et al., 2016). To improve nitrogen removal performance in cold weather, C/N and S/N were increased to 5 and 1, same as the start-up stage. Then TN decreased to 3.34 and 3.59 mg L1 in HEFTW and AEFTW, suggesting that the TN removal efficiency could be improved effectively at low temperature by increasing the electron donor supply for microorganism utilization. TN in the effluents correlated with temperature (for EFTW, r = 0.821, p < 0.001; HEFTW, r = 0.779, p < 0.001; and AEFTW, r = 0.764, p < 0.001). With the temperature dropped from 30 to 10 °C, the TN removal efficiency decreased by 34.3%, 11.2% and 16.5% in EFTW, HEFTW and AEFTW, respectively. And this indicated that the external electron donors reduced the dependence of TN removal on temperature. The effluent NO 3 -N showed a similar trend as TN (Fig. 2C) for the reason that the influent was rich in nitrate. Nitrate removal efficiency was approximately 100% from April to December as a result of high denitrifying efficiency in HEFTW and AEFTW. However, the effluent NO 3 -N in AEFTW slightly increased from day 178. The nitrate in EFTW effluent exceeded the influent from day 0 to 45 and day 231 to 330, revealing the loss of microbial capacity of denitrifiers in EFTW at low temperature, probably due to the lack of biodegradable organics. And the effluents NO 3 -N increased with water temperature decreasing (r = 0.842, p < 0.001, r = 0.758, p < 0.001 and r = 0.444, p = 0.044 for EFTW, AEFTW

+  Fig. 2. TN, NO 3 -N, NH4-N and NO2 -N concentration variations of influent and effluent and temperature change in HEFTW, AEFTW and EFTW.

L. Gao et al. / Bioresource Technology 234 (2017) 243–252

and HEFTW, respectively). This result illustrated that the denitrification rate was strongly affected by temperature and that the heterotrophic denitrifiers in HEFTW was most robust to overcome the adverse effects of temperature. As shown in Fig. 2B, from day 0 to 200, the effluents NH+4-N in HEFTW, AEFTW and EFTW showed no significant difference (p > 0.05) from each other with the average concentration of 0.63, 0.75 and 0.43 mg L1. However, the average effluent NH+4-N of HEFTW (2.01 mg L1, a) turned to be much higher than that of AEFTW (0.89 mg L1, b) and EFTW (0.54 mg L1, b) after day 200 with the highest NH+4-N concentration reaching 2.70 mg L1 and the lowest NH+4-N removal efficiency of 24%. The most likely reason was that the heterotrophic denitrifiers outcompete the nitrifiers rivals to thrive in the system after 2 months of operation at a short HRT (1 d) in the anoxic environment (as seen in Fig. S2). During the last stage of the experiment, the effluent nitrate decreased remarkably with the addition of acetate in HEFTW, suggesting that the acetate addition stimulated the thriving of heterotrophic denitrifiers, leading to much higher denitrify rate. Additionally, the effluent NH+4-N increased when HRT was shortened to 1 d, probably due to the fact that the autotrophic nitrifying bacteria required a long reproductive time to accomplish the nitrification process in EFTWs. The average concentrations of effluent NO 2 -N of 3 EFTWs were as follows: AEFTW (0.23 mg L1, b) >HEFTW (0.14 mg L1, ab) >EFTW (0.10 mg L1, a), which proved that the autotrophic denitri fiers were more effective in converting NO 3 to NO2 than the heterotrophic denitrification (Chen et al., 2009; Reyes-Avila et al., 2004). The concentration of NO 2 in AEFTW increased when HRT was decreased to 1 d, due to the slower denitrifying rate of AD than HD. Nevertheless, NO 2 -N was resumed to lower concentration after the autotrophic denitrifiers adapted to the short HRT. As seen in Fig. 3, TN removal efficiencies were 80.5, 88.5, 84.3, 74.6% in HEFTW and 59.9, 89.4, 84.1, 71.8% in AEFTW in spring, summer, autumn and winter, significantly higher than 17.5, 28.9, 20.4, 11.2% in EFTW. It could be roughly calculated that the microbial transformation contributed about 67.3–85.0% to the total nitrogen removal in HEFTW and 67.7–84.4% in AEFTW according to Eqs. (2) and (3). This finding was consistent with the previous reports that the microbial conversion was the dominant pathway to nitrogen removal (Reinhardt et al., 2006; Søvik and Mørkved, 2008). The nitrogen removal performance in HEFTW and AEFTW were vastly superior to a hybrid floating treatment bed using plant and periphyton (30% in 3 days) (Liu et al., 2016) and an enhanced ecological floating bed of only 15.3% (Wu et al., 2016).

247

In spring, the average effluent TN of AEFTW and HEFTW was 5.44 mg L1 and 2.23 mg L1. And the extra N in AEFTW was mainly NO 3 -N, implying that the autotrophic denitrifies required longer time to adapt to a new water environment. Nonetheless, the effluent TN of HEFTW and AEFTW were extremely close in autumn and winter. NH+4-N accounted for the majority of TN in HEFTW whereas NO 3 -N was overwhelming majority in AEFTW. Higher NH+4-N concentration in HEFTW indicated that nitrification was most likely the rate-limiting step, this was in accordance with the low DO in HEFTW (as discussed in Section 3.4) and the observation of low abundance of nitrifying bacteria in the biomass (as discussed in Section 3.3). The higher NO 3 -N in AEFTW might be due to the decline of the microbial activity of the autotrophic denitrifiers at lower temperature. And the effluent NO 3 -N was effectively reduced after the S/N increased to 1 from day 290 (Fig. 2C). The proportion of NO 3 -N in EFTW effluent exceeded 90% in one year operation, which illustrated that the denitrification was limited because of the shortage of labile organic compounds from the influent and plant generation. No significant difference was observed among the effluent COD of EFTW (28.8 ± 7.10 mg L1), HEFTW (30.6 ± 7.91 mg L1) and AEFTW (29.5 ± 7.22 mg L1), suggesting that the added acetate was completely consumed and the amount of external carbon would not cause the secondary contamination. The TP contents in the effluent of EFTW, HEFTW and AEFTW were 0.19 ± 0.07, 0.20 ± 0.09 and 0.17 ± 0.09 mg L1, respectively. Meanwhile, the effluent S2O2 was undetected in the most time except some 3 unstable running days and the average concentration was 0.67 ± 2.24 mg L1, and the effluent SO2 in AEFTW were 4 232.4 ± 34.9 mg L1, proving the AEFTW with thiosulfate-driven AD a promising process for treating the secondary effluent. 3.2. N2O fluxes in different seasons The mean N2O fluxes were 2.64, 1.19, 2.44, 3.67 mg N m2 d1 in HEFTW; 5.35, 1.14, 5.72, 6.48 mg N m2 d1 in AEFTW and 5.96, 5.72, 5.60, 6.61 mg N m2 d1 in EFTW from spring to winter (Fig. 4). Comparing with the previous studies which reported the rates ranging from 1.27 to 636 mg N m2 d1 in constructed wetlands (CWs) (Sims et al., 2013; Uggetti et al., 2012), the fluxes in this study were little higher than the median N2O flux (3.12 mg N m2 h1) from the CWs (Mander et al., 2014). The possible reasons are as follows: 1) the influent in this study was the simulated secondary effluent which was rich of nitrate; 2) the

Fig. 3. Concentration of N species in effluent and TN removal efficiency of HEFTW (A), AEFTW (B), and EFTW (C) in different seasons. (spr, sum, aut and win represent the seasons of spring, summer, autumn and winter).

248

L. Gao et al. / Bioresource Technology 234 (2017) 243–252

Fig. 4. N2O fluxes in HEFTW and EFTW in different seasons (Spr, sum, aut and win represent spring, summer, autumn and winter).

plant roots in EFTW were directly growing in water column rather than in substrate as in CWs, thus N2O can be transported from root zone immediately to atmosphere via vascular plants (Yavitt and Knapp, 1998). In CWs, wetland soil or substrate may cause N2O consumption exceeding N2O production and reduce the diffusion rate of N2O (Jørgensen et al., 2012). As shown in Fig. 4, the N2O Fluxes varied seasonally in HEFTW with a higher mean flux in winter than in other seasons. The possible explanation was that the low temperature in winter slowed down the overall rate of denitrification, resulting in N2O accumulation in the end products (Søvik et al., 2006). In general, EFTW had the highest fluxes in four seasons without significant change and HEFTW had the lowest fluxes in spring, autumn and winter with one exception of the lowest flux of AEFTW in summer. External carbon source and sulfur source have dramatically enhanced the nitrogen removal efficiency without negative impact on N2O emissions. Carbon (acetate) addition significantly reduced the N2O fluxes, because high C/N promoted complete denitrification in transforming NO 3 to N2 and then declined the accumulation of intermediate product. The Previous studies have demonstrated that the successful treatment performance and low N2O emission in CWs could be obtained at COD/N ratio of 5 (Wu et al., 2009). Meanwhile, high S/N ratio plays an important role in reducing N2O emission in sulfur-oxidizing autotrophic denitrification system (Yang et al., 2016). The low N2O emissions of AEFTW and HEFTW were accompanied by the high TN removal efficiency and low NO x -N concentration in summer (Fig. 3). The increased NO x -N concentration and the decreased TN removal efficiency led to more N2O emission. As described above, the major nitrogen removal process in HEFTW and AEFTW was denitrification, N2O was most likely generated from the incomplete AD and HD. The effluent nitrite was higher in AEFTW than HEFTW (Fig. 2D) in spring, autumn and winter, explaining the higher N2O flux in AEFTW. Although TN removal efficiency was similar in these two processes, high nitrite concentration during denitrification led to the accumulation of NO and N2O. 3.3. Microbial community analysis As Table 1 shows, after removing the chimera sequences and the ineffective sequences, Illumina MiSeq analysis of the samples from 3 EFTWs yielded a total of 379,315 sequences. The lowest ratio of high quality sequences was 90.47%, suggesting the results were reliable and representative. 8535–122553 OTUs were clustered with 97% similarity by performing the alignment and filter-

ing. Chao 1 and ACE richness estimators increased in HEFTW, AEFTW and EFTW implying that all microbial community became prosperous in the experiment. Comparing with E1, E2 richness (Chao1: 2439, ACE: 3391) increased greatly and was greater than the richness indexes of all the other samples. Plants in EFTW promoted the bacterial growth on fillers (Wu et al., 2016). Chen et al. (2015) found that planted CWs were of greater microbial richness than the unplanted CWs. The external electron donors in HEFTW and AEFTW promoted the growth of some specific microorganisms, thus the quantity of minority microbial species relatively reduced, which made the Chao1 and ACE richness index less than E2. The Simpson and Shannon estimators decreased from 0.9931, 8.49 to 0.9814, 8.26 in HEFTW, declaring the reduction of microbial community diversity. This was most because that the organic carbon sources supported the growth of heterotrophic denitrifiers, and other bacteria were outcompeted by the heterotrophic bacteria. On the other side, the higher Simpson and Shannon indexes in EFTW showed that the microbial diversity in EFTW was the highest among these three wetlands in the experiment. Simpson estimators were almost equal in A1 and A2, implying that the thiosulfate addition might produce less adverse impact on the microbial diversity than the acetate addition. In order to understand the phylogenetic diversity of bacterial communities HEFTW, AEFTW and EFTW systems, qualified reads were assigned to phyla, class (Fig. S1) and genus level (Fig. 5). Proteobacteria, Actinobacteria, Chloroflexi and Bacteroidetes accounted for approximately 80% in all samples. Proteobacteria was the dominant phylum in all samples accounting for 46.4, 65.2, 47.5, 62.0, 34.5 and 44.3% in H1, H2, A1, A2, E1 and E2, respectively (Fig. S1A). The Proteobacteria phylum contained many species of bacteria related to global carbon, nitrogen and sulfur cycling, which were widely found in natural and constructed wetlands (Ansola et al., 2014). The dominated classes were Betaproteobacteria, Alphaproteobacteria and Gammaproteobacteria in H1 and A1, while E1 was dominated by Anaerolineae, Alphaproteobacteria, Actinobacteria and Betaproteobacteria (Fig. S1B). After eight months running, Betaproteobacteria increased to 36.4 and 34.1% in H2 and A2. The proportion of Alphaproteobacteria increased from 15.1% in E1 to 20.8% in E2 and become the largest class, illustrating that the external carbon and sulfur source altered the structure of microbial community in EFTWs. Fig. 5 shows the relative abundance of the most dominant bacteria genera for the HEFTW, AEFTW and EFTW after a long period of operation. The most abundant genera for HEFTW were Dechloromonas, C39, Desulfobacter, Thiothrix, Sulfuricurvum, Flavobacterium and Thauera. Dechloromonas and Thauera are two denitrifying bacteria belonging to family Rhodocyclaceae which are capable of utilizing acetate under anoxic conditions (Ginige et al., 2005). Desulfobacter is a representative genus of sulfate-reducing bacteria in wetlands which could oxidize acetate into CO2 via tricarboxylic acid or acetyl-CoA (Chen et al., 2016). Although there were nearly no documented knowledge of the genus C39, it was also abundant in both H1 (2.10%) and H2 (2.89%), indicating that some species of C39 may be associated with the acetate in denitrification process. For AEFTW, the dominant genera were Thiobacillus, Sulfuricurvum, Sulfurimonas, Nitrospira, Thiothrix, Flavobacterium, Dechloromonas. Genera related to sulfur-oxidizing AD accounted for 19.8% of total bacterial community. This finding further verified that AEFTW were rich in autotrophic denitrifers and coincided to the excellent TN removal efficiency in AEFTW. Thiobacillus and Sulfurimonas were the most commonly reported autotrophic denitrifiers in CWs (Shao et al., 2010) and up-flow biofilter (Zhou et al., 2017). In EFTW, Nitrospira, Flavobacterium, Nitratireductor and devosia were the most abundant genera. Nitrospira are well-known nitrite oxidizing bacteria which transform nitrite into nitrate. Nitrospira accelerated the consumption of nitrite and then indirectly

249

L. Gao et al. / Bioresource Technology 234 (2017) 243–252 Table 1 High quality sequences, richness and diversity estimators of the sludge samples. Sample

Effective sequences

High quality sequences

Ratio (%)

OTUs

Chao1

ACE

Simpson

Shannon

H1 H2 A1 A2 E1 E2

78242 59016 67920 62552 51291 60294

72852 54381 62335 57667 46403 55579

93.1 92.1 91.8 92.2 90.2 92.2

11477 11011 10668 10995 8535 12553

1603 1967 1517 1854 1346 2439

2105 2739 2091 2509 1911 3391

0.9931 0.9814 0.9843 0.9854 0.9780 0.9946

8.49 8.26 8.01 8.25 7.24 9.33

Fig. 5. Taxonomic classification of bacterial reads retrieved from (A) H2, (B) A2, and (C) E2 at the genus level representative of matured bacterial community of HEFTW, AEFTW and EFTW. Genera making up less than 0.3% of total composition are defined as ‘‘others”.

enhanced the transformation of ammonia into nitrite in EFTW. Flavobacterium is a denitrification-related genus frequently detected in denitrification systems (Wang et al., 2014). Nitratireductor could reduce nitrate to nitrite. Devosia was related to nitrogen-fixing root-nodule symbiosis with aquatic plants (Rivas et al., 2002). Interestingly, Dechloromonas, Flavobacterium, Thiothrix and Sulfuricurvum were found abundant in HEFTW and AEFTW, indicating

that HD and AD might proceed simultaneously in these two EFTWs. Desulfobacter was the second largest genus in HEFTW, which could utilize acetate as electron donor and transform sulfate into sulfide or other sulfur compound. Then, sulfur compound promoted the growth of Thiothrix and Sulfuricurvum for denitrification. The autotrophic denitrifiers coexisted with heterotrophic denitrifiers in EFTWs, which is beneficial to the nitrogen removal, as it offers a possibility to use both acetate and thiosulfate in one EFTW

250

L. Gao et al. / Bioresource Technology 234 (2017) 243–252

to form the mixotrophic denitrification. Zhou et al. (2017) found that Thiobacillus and Dechloromonas existed in both autotrophic denitrification reactors and heterotrophic reactors. Sulfatereducing bacteria, like Desulfococcus and Desulfovibrio also coexisted with Thiobacillus in the heterotrophic system, which further proved the occurrence of mixotrophic denitrification. Table 2 exhibited the classification of functional genera related to heterotrophic denitrification, facultative autotrophic denitrification, autotrophic denitrification and Nitrification. Composition of denitrifies dramatically increased from 10.6% to 19.7% and 11.3% to and 24.2% in HEFTW and AEFTW, meanwhile the denitrifies in EFTW remained almost unchanged from 6.62 to 6.57%. This could explain why the HEFTW and AEFTW have analogously excellent nitrogen removal performance compared with EFTW. In contrast to the denitifcation-related genera, nitrification-related genera were more abundant in EFTW (4.65%) at the final stage of experiment. The proportion of nitrifiers decreased from 1.07 to 0.58% as the consequence of depressed growth of nitrifiers by tremendous amplification of heterotrophic denitrifies. On the contrary, nitrification-related genera accounted for 2.92–2.55% of total bacteria, demonstrating that the autotrophic denitrifies coexisted with the nitrifiers. This discovery revealed the reason why thiosulfatedriven AD could achieve higher nitrification performance than HD in autumn and winter. RDA analysis, a kind of constrained ordination, was used to analyze the relationship between the relative abundance of predominant bacteria (genus level) and the environmental variables. As displayed in Fig. 6, the first and second axis represented 47.0% and 40.6% of variation, respectively. It showed that the 6 samples could cluster into 3 distinct groups including group I (A2), group II (H2) and group III (A1, H1, E1 and E2). This further proved that different electron donors significantly altered the microbial compositions in AEFTW and HEFTW. The microbial composition in EFTW without external electron donor was similar during the experiment. The angles of the arrows depicted in RDA indicated that Thiobacillus, Sulfurimonas, Sulfuricurvum, Thiothrix, Flavobacterium and nitrospira were positively correlated with S/N and NO 2 -N concentration in the effluent. Dechloromonas, C39 and

Fig. 6. Redundancy analysis triplot to explore the correlation between environmental factors and bacterial community based on genus level structure.

Desulfobacter were positively correlated with C/N, pH and NH+4-N concentration. Thiobacillus, Sulfurimonas, Sulfuricurvum, Thiothrix, Dechloromonas and C39 were negatively correlated with TN. Flavobacterium and nitrospira were negatively correlated with C/N, pH and NH+4-N concentration. In summary, the addition of carbon and sulfur all contributed to the decrease of DO. Then the low DO favored the reproduction of autotrophic and heterotrophic denitrifers, leading to higher nitrate and TN removal. Thiosulfate addition accelerated the nitrification and denitrification by stimulating the growth of nitrifiers (Nitrospira) and autotrophic denitrifers (Thiobacillus, Sulfurimonas, Sulfuricurvum and Thiothrix). Acetate addition decreased the nitrification since the growth of nitrifiers

Table 2 Relative abundance of important functional genera. EFTW

HEFTW

Sample Heterotrophic denitrification related genera (Zhong et al., 2015)

Facultative autotrophic denitrification genera (Lee et al., 2013)

Autotrophic denitrification related genera (Shao et al., 2010)

Sum of the nitrification related genera

EFTW

H1

H2

A1

A2

E1

E2

Bacillus Flavobacterium Hyphomicrobium Rhodobacter Comamonas Hydrogenophaga Rubrivivax Azospira Dechloromonas Thauera Acinetobacter Pseudomonas Agrobacterium Paracoccus Thiobacillus Sulfurimonas Sulfuricurvum Thiovirga Thiothrix Arcobacter

0.11 1.79 0.27 3.64 0.01 0.23 0.02 0 1.4 0.12 1.52 0.44 0.13 0.15 0.15 0.31 0.08 0.06 0.04 0.15 10.6

0.02 1.10 0.52 0.65 0.01 0.41 0.03 0 10.75 0.97 0.22 0.09 0.17 0.09 0.24 0.24 1.38 0.12 1.86 0.84 19.7

0.08 2.07 0.29 1.27 0.02 0.18 0.02 0 0.48 2.11 0.08 0.06 0.13 0.06 2.97 0.33 0.17 0.08 0.89 0.03 11.3

0.03 1.81 0.33 0.19 0 0.20 0.03 0.01 1.19 0.08 0.1 0.06 0.09 0.03 7.77 3.00 6.93 0.11 2.13 0.11 24.2

0.05 0.63 0.44 3.96 0.01 0.15 0.01 0 0.31 0.09 0.09 0.08 0.09 0.07 0.17 0.21 0.06 0.1 0.05 0.05 6.62

0.1 2.14 0.7 0.61 0.01 0.48 0.27 0.01 0.68 0.11 0.05 0.16 0.36 0.13 0.2 0.24 0.10 0.09 0.08 0.05 6.57

Nitrosomonas Nitrospira

0.03 1.04 1.07

0.00 0.58 0.58

0.03 2.89 2.92

0.01 2.54 2.55

0.00 0.51 0.51

0.05 4.60 4.65

Sum of the denitrification related genera Nitrification related genera

AEFTW

L. Gao et al. / Bioresource Technology 234 (2017) 243–252

(Nitrospira) was depressed. Thiothirx was positively correlated with S/N and C/N, suggesting that the Thiothirx might be an indicator of mixotrophic denitrification.

3.4. Physicochemical responses During the long period study, pH and dissolved oxygen were monitored (Fig. S2). The pH in EFTW was 6.99 ± 0.17, significantly lower than the influent (7.56 ± 0.15) (P < 0.05). Lower pH in planted FTWs was observed in previous FTWs experiments (White and Cousins, 2013). This change is possibly due to the release of CO2 and acidic exudate from the plant root system respiration (Iamchaturapatr et al., 2007), or by the degradation of organic matter in the release of organic acids (Ijaz et al., 2015). High nitrification effect existing in EFTW may also reduce pH as described in Section 3.1. The pH in AEFTW (6.79 ± 0.18) was much lower than EFTW, because the thiosulfate-driven AD generated massive protons when reduing nitrate into N2 as described in equation (1). The pH in HEFTW (7.64 ± 0.26) was little higher than the influent. The alkaline pH could be explained that the acetatedriven HD produced sufficient alkalinity. DO in EFTW declined from 4.45 ± 1.68 (influent) to 1.48 ± 1.10 mg L1 (P < 0.05), and the lower oxygen concentration in EFTW was affected by many factors. First, the pervasive root system and spherical media provided huge surface for microbial populations to colonize. And the respiration of living plant roots, dead plant roots and microbial communities contributed to considerable oxygen consumption (Tanner and Headley, 2011). Further, although plants could transport the atmospheric oxygen into their rhizosphere, the EFTW obstructed the air/water interface which prevented the exchange of DO with atmospheric oxygen. DO in HEFTW (0.38 ± 0.26 mg L1) and AEFTW (0.81 ± 0.69 mg L1) were significantly lower than EFTW, which were typically anoxic conditions. The external organic acetate and inorganic thiosulfate consumed DO in water and facilitated the microbial denitrification in AEFTW and HEFTW.

4. Conclusions The carbon and sulfur sources addition effectively improved TN removal efficiency of HEFTW and AEFTW. AEFTW was more suitable for the simultaneous removal of NH+4-N and NO 3 -N. The N2O emissions reduced considerably in HEFTW and AEFTW, especially in summer. The electron donors induced distinct shifts in microbial community structures. Dechloromonas, Thiobacillus and Nitrospira became the largest genera in HEFTW, AEFTW and EFTW, respectively. Autotrophic and heterotrophic denitrification could simultaneously occur in HEFTW and AEFTW.

Acknowledgements The study was supported by the National Natural Science Foundation of China (51378306, 51678356, 51478262). The authors are grateful to Shanghai Personal Biotechnology Co., Ltd. for the technological support on high-throughput sequencing and analysis.

Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.biortech.2017.03. 036.

251

References Ansola, G., Arroyo, P., de Miera, L.E., 2014. Characterisation of the soil bacterial community structure and composition of natural and constructed wetlands. Sci. Total Environ. 473, 63–71. APHA, 2005. Standard Methods for Water and Wastewater Examination. American Public Health Association, Washington, D.C.. Bokulich, N.A., Subramanian, S., Faith, J.J., Gevers, D., 2013. Quality-filtering vastly improves diversity estimates from Illumina amplicon sequencing. Nat. Methods 10 (1), 57–59. Borne, K.E., Fassman-Beck, E.A., Tanner, C.C., 2014. Floating Treatment Wetland influences on the fate of metals in road runoff retention ponds. Water Res. 48, 430–442. Borne, K.E., Tanner, C.C., Fassman-Beck, E.A., 2013. Stormwater nitrogen removal performance of a floating treatment wetland. Water Sci. Technol. 68 (7), 1657– 1664. Chen, C., Wang, A., Ren, N., Deng, X., Lee, D.-J., 2009. Optimal process pattern for simultaneous sulfur, nitrogen and carbon removal. Water Sci. Technol. 59 (4), 833–837. Chen, Y., Wen, Y., Tang, Z., Huang, J., Zhou, Q., 2015. Effects of plant biomass on bacterial community structure in constructed wetlands used for tertiary wastewater treatment. Ecol. Eng. 84, 38–45. Chen, Y., Wen, Y., Zhou, Q., Huang, J., Vymazal, J., Kuschk, P., 2016. Sulfate removal and sulfur transformation in constructed wetlands: the roles of filling material and plant biomass. Water Res. 102, 572–581. Chung, J., Amin, K., Kim, S., Yoon, S., Kwon, K., Bae, W., 2014. Autotrophic denitrification of nitrate and nitrite using thiosulfate as an electron donor. Water Res. 58, 169–178. García-Lledó, A., Vilar-Sanz, A., Trias, R., Hallin, S., Bañeras, L., 2011. Genetic potential for N2O emissions from the sediment of a free water surface constructed wetland. Water Res. 45 (17), 5621–5632. GB3838-2002, 2002. Chinese National Environmental Quality Standards for Surface Water. Ministry of Environmental Protection, China. Ginige, M.P., Keller, J., Blackall, L.L., 2005. Investigation of an acetate-fed denitrifying microbial community by stable isotope probing, full-cycle rRNA analysis, and fluorescent in situ hybridization-microautoradiography. Appl. Environ. Microbiol. 71 (12), 8683–8691. Gruber, N., Galloway, J.N., 2008. An Earth-system perspective of the global nitrogen cycle. Nature 451 (7176), 293–296. Iamchaturapatr, J., Yi, S.W., Rhee, J.S., 2007. Nutrient removals by 21 aquatic plants for vertical free surface-flow (VFS) constructed wetland. Ecol. Eng. 29 (3), 287– 293. Ijaz, A., Shabir, G., Khan, Q.M., Afzal, M., 2015. Enhanced remediation of sewage effluent by endophyte-assisted floating treatment wetlands. Ecol. Eng. 84, 58– 66. IPCC, 2007. Climate Change 2007: The Physical Science Basis, Contribution of Working Group I to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA. Jørgensen, C., Struwe, S., Elberling, B., 2012. Temporal trends in N2O flux dynamics in a Danish wetland – effects of plant-mediated gas transport of N2O and O2 following changes in water level and soil mineral-N availability. Glob. Change Biol. 18 (1), 210–222. Karine, E.B., Elizabeth, A.F., Chris, C.T., 2013. Floating treatment wetland retrofit to improve stormwater pond performance for suspended solids, copper and zinc. Ecol. Eng., 54 Lee, D.-J.J., Pan, X., Wang, A., Ho, K.-L.L., 2013. Facultative autotrophic denitrifiers in denitrifying sulfide removal granules. Bioresour. Technol. 132, 356–360. Li, C., Dong, Y., Lei, Y., Wu, D., Xu, P., 2015. Removal of low concentration nutrients in hydroponic wetlands integrated with zeolite and calcium silicate hydrate functional substrates. Ecol. Eng. 82, 442–450. Liu, J., Wang, F., Liu, W., Tang, C., Wu, C., Wu, Y., 2016. Nutrient removal by upscaling a hybrid floating treatment bed (HFTB) using plant and periphyton: from laboratory tank to polluted river. Bioresour. Technol. 207, 142–149. Mander, Ü., Dotro, G., Ebie, Y., Towprayoon, S., Chiemchaisri, C., Nogueira, S., Jamsranjav, B., Kasak, K., Truu, J., Tournebize, J., Mitsch, W.J., 2014. Greenhouse gas emission in constructed wetlands for wastewater treatment: a review. Ecol. Eng. 66, 19–35. Mander, Ü., Maddison, M., Soosaar, K., Koger, H., Teemusk, A., Truu, J., Well, R., Sebilo, M., 2015. The impact of a pulsing water table on wastewater purification and greenhouse gas emission in a horizontal subsurface flow constructed wetland. Ecol. Eng. 80, 69–78. Mora, M., López, L.R., Gamisans, X., Gabriel, D., 2014. Coupling respirometry and titrimetry for the characterization of the biological activity of a SO-NR consortium. Chem. Eng. J. 251, 111–115. Ravishankara, A.R., Daniel, J.S., Portmann, R.W., 2009. Nitrous oxide (N2O): the dominant ozone-depleting substance emitted in the 21st century. Science 326 (5949), 123–125. Reinhardt, M., Müller, B., Gächter, R., Wehrli, B., 2006. Nitrogen removal in a small constructed wetland: an isotope mass balance approach. Environ. Sci. Technol. 40 (10), 3313–3319. Reyes-Avila, J., Razo-Flores, E.a., Gomez, J., 2004. Simultaneous biological removal of nitrogen, carbon and sulfur by denitrification. Water Res. 38 (14–15), 3313– 3321.

252

L. Gao et al. / Bioresource Technology 234 (2017) 243–252

Rivas, R., Velázquez, E., Willems, A., Vizcaíno, N., Subba-Rao, N.S., Mateos, P.F., Gillis, M., Dazzo, F.B., Martínez-Molina, E., 2002. A new species of Devosia that forms a unique nitrogen-fixing root-nodule symbiosis with the aquatic legume Neptunia natans (Lf) Druce. Appl. Environ. Microbiol. 68 (11), 5217–5222. Søvik, A.K., Augustin, J., Heikkinen, K., Huttunen, J.T., Necki, J.M., Karjalainen, S.M., Kløve, B., Liikanen, A., Mander, U., Puustinen, M., Teiter, S., Wachniew, P., 2006. Emission of the greenhouse gases nitrous oxide and methane from constructed wetlands in Europe. J. Environ. Qual. 35 (6), 2360–2373. Søvik, A.K., Mørkved, P.T., 2008. Use of stable nitrogen isotope fractionation to estimate denitrification in small constructed wetlands treating agricultural runoff. Sci. Tot. Environ. 392 (1), 157–165. Sahinkaya, E., Dursun, N., 2015. Use of elemental sulfur and thiosulfate as electron sources for water denitrification. Bioprocess Biosyst. Eng. 38 (3), 531–541. Shao, M.-F.F., Zhang, T., Fang, H.H., 2010. Sulfur-driven autotrophic denitrification: diversity, biochemistry, and engineering applications. Appl. Microbiol. Biotechnol. 88 (5), 1027–1042. Sims, A., Gajaraj, S., Hu, Z., 2013. Nutrient removal and greenhouse gas emissions in duckweed treatment ponds. Water Res. 47 (3), 1390–1398. Tanner, C.C., Headley, T.R., 2011. Components of floating emergent macrophyte treatment wetlands influencing removal of stormwater pollutants. Ecol. Eng. 37 (3), 474–486. Uggetti, E., García, J., Lind, S.E., Martikainen, P.J., Ferrer, I., 2012. Quantification of greenhouse gas emissions from sludge treatment wetlands. Water Res. 46 (6), 1755–1762. Wang, R., Zheng, P., Xing, Y.-J., Zhang, M., Ghulam, A., Zhao, Z.-Q., Li, W., Wang, L., 2014. Anaerobic ferrous oxidation by heterotrophic denitrifying enriched culture. J. Ind. Microbiol. Biotechnol. 41 (5), 803–809. White, S.A., Cousins, M.M., 2013. Floating treatment wetland aided remediation of nitrogen and phosphorus from simulated stormwater runoff. Ecol. Eng. 61, 207– 215.

Wu, J., Zhang, J., Jia, W., Xie, H., Gu, R.R., Li, C., Gao, B., 2009. Impact of COD/N ratio on nitrous oxide emission from microcosm wetlands and their performance in removing nitrogen from wastewater. Bioresour. Technol. 100 (12), 2910–2917. Wu, Q., Hu, Y., Li, S., Peng, S., Zhao, H., 2016. Microbial mechanisms of using enhanced ecological floating beds for eutrophic water improvement. Bioresour. Technol. 211, 451–456. Xu, J.-H.H., He, S.-B.B., Wu, S.-Q.Q., Huang, J.-C.C., Zhou, W.-L.L., Chen, X.-C.C., 2016. Effects of HRT and water temperature on nitrogen removal in autotrophic gravel filter. Chemosphere 147, 203–209. Yang, W., Lu, H., Khanal, S.K., Zhao, Q., Meng, L., Chen, G.-H., 2016. Granulation of sulfur-oxidizing bacteria for autotrophic denitrification. Water Res. 104, 507– 519. Yavitt, J.B., Knapp, A.K., 1998. Aspects of methane flow from sediment through emergent cattail (Typha latifolia) plants. New Phytol. 139 (3), 495–503. Zhang, L., Zhao, J., Cui, N., Dai, Y., Kong, L., Wu, J., Cheng, S., 2015. Enhancing the water purification efficiency of a floating treatment wetland using a biofilm carrier. Environ. Sci. Pollut. Res. 23 (8), 7437–7443. Zhong, F., Wu, J., Dai, Y., Yang, L., Zhang, Z., Cheng, S., Zhang, Q., 2015. Bacterial community analysis by PCR-DGGE and 454-pyrosequencing of horizontal subsurface flow constructed wetlands with front aeration. Appl. Microbiol. Biotechnol. 99 (3), 1499–1512. Zhou, W., Li, Y., Liu, X., He, S., Huang, J.C., 2017. Comparison of microbial communities in different sulfur-based autotrophic denitrification reactors. Appl. Microbiol. Biotechnol. 101, 447–453. Zhou, W., Liu, X., Dong, X., Wang, Z., Yuan, Y., 2016. Sulfur-based autotrophic denitrification from the micro-polluted water. J. Environ. Sci. 44, 180–188. Zhu, I., Getting, T., 2012. A review of nitrate reduction using inorganic materials. Environ. Technol. Rev. 1 (1), 46–58.