Abstract

In this study, strain CC76, identified as Enterobacter sp., was tested for the reduction of Fe3+ and denitrification using immobilized pellets with strain CC76 as experimental group (IP) and immobilized pellets with strain CC76 and magnetite powder as experimental group (IPM) in the autotrophic denitrification immobilized systems (ADIS). Compared with IP, a higher nitrate removal rate was obtained with IPM by using three levels of influent Fe3+ (0, 5, and 10 mg/L), four levels of pH (5.0, 6.0, 7.0, and 8.0), and three levels of hydraulic retention time (HRT) (12, 14, and 16 h), respectively. Furthermore, response surface methodology (RSM) analysis demonstrated that the optimum removal ratios of nitrate of 87.21% (IP) and 96.27% (IPM) were observed under the following conditions: HRT of 12 h, pH of 7.0 and influent Fe3+ concentration of 5 mg/L (IP) and 1 mg/L (IPM).

INTRODUCTION

Groundwater is an important source of municipal water supply for domestic and industrial use (Showers et al. 2008). Nitrate is simply transported to groundwater through uncontrolled discharge of nitrate-containing sources, such as chemical fertilizers, industrial or domestic wastes, and landfill leachate (Ghafari et al. 2009). High nitrate concentration in drinking water may bring us various health effects. For instance, infants under six months fed with nitrate contaminated water could have blue baby syndrome, and if untreated, may die (Ghafari et al. 2008; Mousavi et al. 2012). The recommended nitrate concentration limit in the drinking water by WHO and the Chinese Ministry of Health is 10 mg/L (NO3-N) (Fu et al. 2014). Therefore, several nitrate removal technologies such as electro-dialysis, reverse osmosis, adsorption and chemical and biological methods have been used in water treatment (Bhatnagar & Sillanpää 2011; Loganathan et al. 2013). However, biological denitrification is considered to be the most appropriate technology compared to other techniques for treatment of nitrate-contaminated steams. Several bioreactors have been developed for the biological denitrification of wastewater.

Nitrification and denitrification processes have proved individually successful in biofilm reactors, and there are already many different biofilm systems in use, such as trickling filters, rotating biological contactors (RBCs), fixed media reactors, biofilters, fluidized bed reactors, etc. (Makkulath & Thampi 2012; Biswas et al. 2014). The sequencing batch biofilm reactor (SBBR) system, one type of biofilm technology, has attracted much attention because of its ability to take advantage of both a biofilm reactor and a sequencing batch reactor (Ding et al. 2011). In addition, biofilm reactors are also characterized by a lower biomass growth and better sedimentation properties compared to activated sludge flocs (Helness & Qdegaard 2001). Immobilization of bacteria as biofilm on the surface of the carrier can reduce the risk of biomass wash-out (Magri et al. 2012). These immobilization techniques include self-immobilization as granular biomass (Dapena-Mora et al. 2004; López et al. 2008), attachment on the surface of a carrier forming biofilm (Tsushima et al. 2007; Ni et al. 2010), and entrapment of the microbial biomass into gel pellets (Isaka et al. 2007; Furukawa et al. 2009).

Meanwhile, scholars have found that some microorganisms can utilize ferrous iron as an electron donor to convert nitrate into nitrogen gas and these microorganisms have been found in various habitats, such as swine waste lagoons, lake sediments, even freshwater (Chaudhuri et al. 2001; Straub et al. 2004; Muehe et al. 2009).

Iron-reducing bacteria (IRBs) commonly occur in anaerobic systems and play an essential role in iron cycling (Kim et al. 2014). In the present study, we aim to investigate the adaptability of iron-reducing bacterium strain CC76 under different conditions in autotrophic denitrification immobilized systems (ADIS). In this study, the recycling of iron is shown in the ADIS. Furthermore, an ADIS with magnetically immobilized CC76 cells has been designed and operated to enhance the ability for denitrification in groundwater. Herein, we discuss the abilities for nitrate removal and iron reduction through the addition of magnetite, and without it, in the immobilized pellets under different conditions.

MATERIALS AND METHODS

Bacterial strain and culture conditions

Iron reducing bacteria CC76, which has the ability to reduce nitrate and Fe3+ simultaneously, was obtained from Tang Yu oligotrophic reservoir (Shaanxi Province, China) (Su et al. 2016a).

The basal medium was comprised of the following reagents per litre: NaHCO3, 0.5 g; NaNO3, 0.05 g; KH2PO4, 0.05 g; MgSO4·7H2O, 0.05 g; Fe2(SO4)3, 0.1 g; CaCl2, 0.05 g. A trace element solution (2 mL) was added, and the final pH of the medium was adjusted to 7.0 with 1 mol/L NaOH or HCl solution. Ultra-pure water was used in this study. The trace element solution components were as follows: 0.5 mg/L FeSO4·7H2O; 0.5 mg/L CuSO4·5H2O; 0.5 mg/L MgSO4·7H2O; 1.0 mg/L EDTA; 0.2 mg/L ZnSO4; 0.1 mg/L MnCl2·4H2O and 0.2 mg/L CoCl2·6H2O (Su et al. 2016b). The medium was heated with a high pressure steam cooker to 121 °C under an anoxic atmosphere, which put it into a sterile stage. The strain was grown in 1 L bottles containing 0.9 L medium.

Experimental setup

In this study, three reactors were set up: (1) the immobilized pellets without the addition of extra bacteria as the control group; (2) the immobilized pellets with the corresponding bacteria strain CC76 as experimental group (IP); (3) the immobilized pellets with strain CC76 and magnetite powder as experimental group (IPM). Herein, 2% (m/v) sodium alginate (SA) mixed with strain CC76, strain CC76 and magnetite powder, and without either, was slowly injected into 2% (m/v) CaCl2 with a syringe needle (with a diameter of 2 mm) to form homogeneous pellets. And 0.2 g magnetite was added in 100 mL solution of SA mixed with strain CC76. Meanwhile, the reactors were maintained at 30 °C under anaerobic conditions and the water level was kept at the same point. The pH was monitored at regular time intervals through the pH monitoring system during the operation. According to the experimental requirement, the operational conditions were under different hydraulic retention times (HRT) (Phase 1); pH (Phase 2); and influent Fe3+ concentrations (Phase 3). The results of the three phases of the experiments are clearly listed in Table 1. The running of the entire experiment lasted for 1,260 h.

Table 1

Summary of the performance in the ADIS

PhaseHRT during the test [h]pH during the testInitial NO3-N [mg/L]Initial Fe (III) [mg/L]Cycle times
Phase 1 Phase 1.1 12 30 20 10 
Phase 1.2 14 30 20 10 
Phase 1.3 16 30 20 10 
Phase 2 Phase 2.1 14 30 20 10 
Phase 2.2 14 30 20 10 
Phase 2.3 14 30 20 10 
Phase 3 Phase 3.1 12 10 10 10 
Phase 3.2 12 10 10 
Phase 3.3 12 10 10 
Phase 4 Phase 4.1 12 10 
Phase 4.2 12 10 
Phase 4.3 12 10 
Phase 4.4 12 10 
Phase 5 Phase 5.1 10 10 
Phase 5.2 10 10 
Phase 5.3 10 10 
Phase 5.4 10 10 
Phase 5.5 10 10 
Phase 6 Phase 6.1 10 
Phase 6.2 10 
Phase 6.3 10 
Phase 6.4 10 
PhaseHRT during the test [h]pH during the testInitial NO3-N [mg/L]Initial Fe (III) [mg/L]Cycle times
Phase 1 Phase 1.1 12 30 20 10 
Phase 1.2 14 30 20 10 
Phase 1.3 16 30 20 10 
Phase 2 Phase 2.1 14 30 20 10 
Phase 2.2 14 30 20 10 
Phase 2.3 14 30 20 10 
Phase 3 Phase 3.1 12 10 10 10 
Phase 3.2 12 10 10 
Phase 3.3 12 10 10 
Phase 4 Phase 4.1 12 10 
Phase 4.2 12 10 
Phase 4.3 12 10 
Phase 4.4 12 10 
Phase 5 Phase 5.1 10 10 
Phase 5.2 10 10 
Phase 5.3 10 10 
Phase 5.4 10 10 
Phase 5.5 10 10 
Phase 6 Phase 6.1 10 
Phase 6.2 10 
Phase 6.3 10 
Phase 6.4 10 

Analytical methods

Water sample collection was performed every day during the 1,260 h operational running time for the control group, IP and IPM. These samples were used for testing Fe2+, nitrite (NO2-N) and nitrate (NO3-N). Meanwhile, pH was measured with a pH meter (HQ11d, HACH, USA). Nitrate-N concentration was measured by calculating the difference between OD220 and 2 × OD275 of an UV spectrophotometric screening method. Nitrite-N concentration was determined by colorimetry using the N-(1-naphthyl)-1, 2-diaminoethane dihydrochloride method at wavelengths of 540 nm. Fe2+concentration was measured spectrophotometrically with phenanthroline at 510 nm. The reduction rates of NO3-N and Fe3+ were calculated using the formula (C0-Cn)/h. C0 was the initial concentration of NO3-N. Cn was the final concentration of NO3-N at n hours, and h was the time of strain CC76 treatment. All experiments were performed at least in duplicate.

RESULTS AND DISCUSSION

Effect of HRT on denitrification performance

Figure 1(a) presents the average nitrate concentrations of effluent, which were maintained at 10.37 mg/L (IP) and 7.51 mg/L (IPM), 7.24 mg/L (IP) and 5.97 mg/L (IPM), 1.84 mg/L (IP) and 0.67 mg/L (IPM) at an HRT of 12 h, 14 h and 16 h, respectively. The maximum efficiency of 93.84% (IP) and 97.76% (IPM) was observed at an HRT of 16 h, which was higher than 63.47% (IP) and 73.57% (IPM) at an HRT of 12 h, 75.88% (IP) and 80.13% (IPM) at an HRT of 14 h. The removal rate of nitrate decreased when the HRT was changed from 16 h to 12 h (Figure 1(b)). As a result, a longer HRT could facilitate nitrate removal by such denitrifying bacteria, in accordance with Zhou et al. (2011). Moreover, the nitrate concentration of the effluent in the control group was also accompanied by a small drop compared to the influent (the removal efficiency less than 20%), this might be due to the growth of other bacteria in the reactor, which required nitrogen sources, the phenomenon which appeared in the control group also appeared in the next phase of the operation.

Figure 1

Operation performance of the ADIS: (a) changes of NO3-N concentration; (b) changes of NO3-N removal rate; (c) changes of Fe2+ and Fe3+ concentration; (d) changes of Fe3+ reducing rate; (e) changes of NO2-N concentration, IP: the immobilized pellets with corresponding bacteria strain CC76 as experimental group, IPM: the immobilized pellets with strain CC76 and magnetite powder as experimental group.

Figure 1

Operation performance of the ADIS: (a) changes of NO3-N concentration; (b) changes of NO3-N removal rate; (c) changes of Fe2+ and Fe3+ concentration; (d) changes of Fe3+ reducing rate; (e) changes of NO2-N concentration, IP: the immobilized pellets with corresponding bacteria strain CC76 as experimental group, IPM: the immobilized pellets with strain CC76 and magnetite powder as experimental group.

From Figure 1(c), it could be seen that the average concentration of Fe2+ in IP (3.40 mg/L) was lower than in IPM (5.23 mg/L). Meanwhile, the reduction rate of Fe3+ in IP and IPM had been in a state of fluctuation, especially when the HRT was 12 h (Figure 1(d)). Since Fe2+ was the electron donor for nitrate removal, and the nitrate removal rate was obviously in fluctuation, it is concluded that a higher nitrate removal rate was obtained with a higher concentration of Fe2+. The reason is that the immobilized pellets with strain CC76 can use existing Fe2+ as an electron donor for denitrification (Su et al. 2016b). In addition, Figure 1(e) shows that the concentration of nitrite in IP and IPM had remained at a low level, which was no more than 0.77 mg/L and 0.50 mg/L, and the nitrite of the control group was also maintained in the normal range (less than 0.04 mg/L). It can be concluded that pellets with magnetite were beneficial to improving the efficiency of nitrate removal, and accelerating the degradation of nitrate. Meanwhile, the nitrate removal efficiency of the effluent increased with the increase in HRT.

Effect of pH on denitrification performance

From Figure 1(a), it is shown that the nitrate was removed basically in the reactor when the HRT was 16 h. However, the extension of the HRT did not significantly improve the removal rate of nitrate. Therefore, the HRT was adjusted to 14 h in order to describe the effect of pH on the nitrate removal more accurately, and the pH was set to 5.0, 6.0, 8.0 just as Table 1 shows, pH of 7.0 had been studied at an HRT of 14 h. The highest average nitrate removal efficiency was obtained in the neutral (pH of 7.0) condition and weak acid (pH of 6.0) condition compared to other treatments, and the corresponding average removal efficiencies of 75.88% and 90.05% were obtained. It was observed from Figure 1(b) that the average nitrate removal rate of IP rose from 1.48 mg/L/h to 1.63 mg/L/h when the pH increased from 5.0 to 7.0, but then dropped to 1.50 mg/L/h at a pH of 8.0. Meanwhile, the average nitrate removal rate of IPM was higher than IP under the same pH conditions. The maximum average nitrate removal rate (1.97 mg/L/h) of IPM was observed at a pH of 6.0. Genarally, the results showed neutral or weakly acidic conditions were beneficial for the removal of nitrate, which was consistent with a previous study (Su et al. 2015).

During the pH experiment, the concentration of effluent Fe2+ was also in a state of fluctuation; this was probably caused by the iron cycle (Su et al. 2016b). This process was described as follows: strain CC76 can convert Fe3+ to Fe2+. Fe2+ which has been converted could become an electron donor for denitrification to convert Fe2+ to Fe3+ at the same time. It can be seen from Figure 1(c) that the concentration of Fe2+ decreased with the increase of pH; especially when the pH was 8.0, the average concentration of Fe2+ was only 1.35 mg/L (IP) and 1.53 mg/L (IPM), and the other phases were: 5.22 mg/L and 7.98 mg/L (pH of 5.0); 4.91 mg/L and 4.56 mg/L (pH of 6.0). Figure 1(e) shows that the concentration of effluent NO2 was 1.60 mg/L (IP) and 0.23 mg/L (IPM) when the pH was 5.0, and 1.92 mg/L (IP) and 0.095 mg/L (IPM) when the pH was 8.0. Different levels of accumulation in the effluent nitrite were obtained in acidic (pH of 5.0) and alkalescent conditions (pH of 8.0), which might be due to the nitrite reductase being inhibited in the partial acid and alkaline conditions (Sorokin et al. 2011).

Effect of the influent Fe3+ concentration on denitrification performance

Figure 1(a) shows that the concentration of Fe3+ was set to 10 mg/L (Phase 3.1), 5 mg/L (Phase 3.2), 0 mg/L (Phase 3.3) under the condition of a low concentration of nitrate (10 mg/L). As the HRT of a low concentration nitrate would be shortened, the HRT of this stage was set to 12 h. The results showed that the average nitrate removal efficiency and removal rate of IP were 98.72% and 0.81 mg/L/h (Phase 3.1), 99.79% and 0.81 mg/L/h (Phase 3.2), 18.22% and 0.15 mg/L/h (Phase 3.3); The corresponding average nitrate removal efficiency and removal rate of IPM were 99.08% and 0.81 mg/L/h, 99.75% and 0.81 mg/L/h, 23.49% and 0.19 mg/L/h, respectively. As shown in Figure 1(c), when the concentration of the influent Fe3+ decreased from 10 mg/L to 5 mg/L, the effluent Fe2+ did not vary obviously. It could be explained by sufficient concentration of Fe2+ being provided for denitrification by using Fe2+ as an electron donor when the concentration of the influent Fe3+ was 5 mg/L. When the concentration of the influent Fe3+ increased to 10 mg/L, the concentration of the effluent Fe2+ did not increase, probably due to chemical oxidation and sedimentation. It was found that the nitrate removal efficiency was increased slightly in the experimental groups (IP and IPM). The nitrate could be removed effectively in the presence of a small amount of Fe3+, but more Fe3+ did not mean the removal efficiency and rate would be higher. As in the autotrophic environment, strain CC76 can convert the oxidized Fe (III) to Fe (II), and the iron cycle was formed in this process (Su et al. 2016b).

In this study, there was no discovery of a large amount of nitrite accumulation (Figure 1(e)), which would provide a safe theoretical basis for the low concentration of nitrate in groundwater treatment.

Analysis of ADIS in the low concentration of effluent nitrate by RSM

The Box-Behnken design was used to analyze the interactive effects of important variables that significantly affect the removal of nitrate by strain CC76 at a low concentration, including HRT, pH, and the influent Fe3+ concentration as shown in Table 1. (Phase 4.1–6.4). Statistical analysis was performed using the Design-Expert (8.0.6.1) program with the SAS software package. Effluent concentration of nitrate is shown in Figure 2(a). In the IP, response surface methodology (RSM) analysis demonstrated that the maximum removal ratio (87.21%) and rate (0.70 mg/L/h) of nitrate occurred under the conditions of an HRT of 12 h, pH of 7.0, and influent Fe3+concentration of 5 mg/L; in the IPM, RSM analysis demonstrated that the maximum removal ratio (96.27%) and rate (0.79 mg/L/h) of nitrate occurred under the conditions of an HRT of 12 h, pH of 7.0, and influent Fe3+ concentration of 1 mg/L. This difference of optimal conditions could due to magnetite powder being added in the IPM. During the whole experiment, the concentration of nitrite is always maintained at a low level (Figure 2(c)). Moreover, a higher removal ratio of nitrate was obtained in the IPM, since a high concentration zone was formed by the adsorption of magnetite to increase the nitrate removal. It has been reported that magnetic nanomaterial could be used for adsorption removal of heavy metals (Pb (II), Cr (III)) with its large surface area and high adsorption capacity (Lingamdinne et al. 2017). Similarly, with added magnetite powder in the immobilized pellets, its adsorption ability could promote nitrate removal.

Figure 2

Operation performance of the reactors: changes of NO3-N concentration (a); changes of Fe2+ and Fe3+ concentration (b); changes of NO2-N concentration (c) at 1,632–2,530 h.

Figure 2

Operation performance of the reactors: changes of NO3-N concentration (a); changes of Fe2+ and Fe3+ concentration (b); changes of NO2-N concentration (c) at 1,632–2,530 h.

The response surfaces, as shown in Figure 3(a) and 3(b), show that the nitrate removal ratio increased with increasing HRT as well as pH ranging from 6.0 to 6.8. The reason is that the bacteria are able to make better use of Fe2+ as an electron donor in the acidic condition (Yang et al. 2006). In addition, it could also be concluded that the nitrate removal ratio enhanced significantly as the HRT increased. This might be due to the increasing residence times in the reactors, which allowed the bacteria to adapt to the new environment and the nitrate contaminant had enough time to degrade successfully. Zhou et al. (2011) suggested that good nitrate removal efficiency was obtained at a long HRT in a lab scale up flow biofilter. Figure 3(c) and 3(d) illustrated the effects of the interaction of initial pH and influent Fe3+concentration in the response process. As shown in Figure 3(c) and 3(d), the nitrate removal ratio increased with increasing pH and Fe3+ to attain optimum conditions, and then decreased with a further increase.

Figure 3

Design-Expert plots: Average nitrate removal ratio as a function of HRT and pH in IP (a) and IPM (b); average nitrate removal ratio as a function of pH and influent Fe3+ concentration in IP (c) and IPM (d).

Figure 3

Design-Expert plots: Average nitrate removal ratio as a function of HRT and pH in IP (a) and IPM (b); average nitrate removal ratio as a function of pH and influent Fe3+ concentration in IP (c) and IPM (d).

CONCLUSIONS

In this study, the iron reducing bacteria CC76, with the ability of removal nitrate and Fe3+, was obtained from Tang Yu oligotrophic reservoir. Immobilized pellets with corresponding bacteria strain CC76 as experimental group (IP) were compared with immobilized pellets with strain CC76 and magnetite powder as experimental group (IPM) to reflect the denitrification performance. During the whole experiment, higher nitrate removal efficiency was obtained in the IPM. It is concluded that the removal rate was increased with a longer HRT and acid conditions. RSM analysis demonstrated that the optimum nitrate removal efficiency and removal rate of 87.21% and 0.70 mg/L/h (IP), 96.27% and 0.79 mg/L/h (IPM) were observed under the conditions of HRT of 12, initial pH of 7.0 and influent Fe3+ concentration of 5 mg/L (IP) and 1 mg/L (IPM) in the low concentration of nitrate of 10 mg/L. Owing to the ability to simultaneously undertake Fe3+ reduction and nitrate removal, CC76 is a promising candidate in the extensive application of effective removal of nitrate by the iron cycle.

ACKNOWLEDGEMENT

This research work was partly supported by the National Natural Science Foundation of China (NSFC) (No. 51678471), the National Key Research and Development Project (No. 2016YFC0200706) and the Science and technology overall Plan of Shaanxi Province under Grant (No. 2016KTCG01-17).

REFERENCES

REFERENCES
Bhatnagar
A.
&
Sillanpää
M.
2011
A review of emerging adsorbents for nitrate removal from water
.
Chem. Eng.
168
,
493
504
.
Chaudhuri
S. K.
,
Lack
J. G.
&
Coates
J. D.
2001
Biogenic magnetite formation through anaerobic biooxidation of Fe(II)
.
Environ. Microbiol.
67
,
2844
2848
.
Dapena-Mora
A.
,
Campos
J. L.
,
Mosquera-Corral
A.
,
Jetten
M. S. M.
&
Méndez
R.
2004
Stability of the ANAMMOX process in a gas-lift reactor and a SBR
.
J. Biotechnol.
110
,
159
170
.
Ding
D.
,
Feng
C.
,
Jin
Y.
,
Hao
C.
,
Zhao
Y.
&
Suemura
T.
2011
Domestic sewage treatment in a sequencing batch biofilm reactor (SBBR) with an intelligent controlling system
.
Desalination
276
(
1–3
),
260
265
.
Fu
F.
,
Dionysiou
D. D.
&
Hong
L.
2014
The use of zero-valent iron for groundwater remediation and wastewater treatment
.
J. Hazard. Mater.
267
(
3
),
194
205
.
Furukawa
K.
,
Inatomi
Y.
,
Qiao
S.
,
Quan
L.
,
Yamamoto
T.
,
Isaka
K.
&
Sumino
T.
2009
Innovative treatment system for digester liquor using anammox process
.
Bioresour. Technol.
100
,
5437
5443
.
Ghafari
S.
,
Hasan
M.
&
Aroua
M. K.
2008
Bio-electrochemical removal of nitrate from water and wastewater
.
Bioresour. Technol.
99
(
10
),
3965
3974
.
Helness
H.
&
Qdegaard
H.
2001
Biological phosphorus and nitrogen removal in a sequencing batch moving bed biofilm reactor
.
Water Sci. Technol.
43
,
233
240
.
Kim
S. J.
,
Park
S. J.
,
Cha
I. T.
,
Min
D.
,
Kim
J. S.
,
Chung
W. H.
,
Chae
J. C.
,
Jeon
C. O.
&
Rhee
S. K.
2014
Metabolic versatility of toluene-degrading, iron-reducing bacteria in tidal flat sediment, characterized by stable isotope probing-based metagenomic analysis
.
Environ. Microbiol.
16
,
189
204
.
Lingamdinne
L. P.
,
Chang
Y. Y.
,
Yang
J. K.
,
Singh
J.
,
Choi
E. H.
,
Shiratani
M.
,
Koduru
R.
&
Attri
P.
2017
Biogenic reductive preparation of magnetic inverse spinel iron oxide nanoparticles for the adsorption removal of heavy metals
.
Chem. Eng. J.
307
,
74
84
.
Loganathan
P.
,
Vigneswaran
S.
&
Kandasamy
J.
2013
Enhanced removal of nitrate from water using surface modification of adsorbents
.
Environ. Manag.
131
,
363
374
.
López
H.
,
Puig
S.
,
Ganigué
R.
,
Ruscalleda
M.
,
Balaguer
M. D.
&
Colprim
J.
2008
Start-up and enrichment of a granular anammox SBR to treat high nitrogen load wastewaters
.
Chem. Technol. Biotechnol.
83
,
233
241
.
Magri
A.
,
Vanotti
M. B.
&
Szogi
A. A.
2012
Anammox sludge immobilized in polyvinyl alcohol (PVA) cryogel carriers
.
Bioresour. Technol.
114
,
231
240
.
Makkulath
G.
&
Thampi
S. G.
2012
Performance of coir geotextiles as attached media in biofilters for nutrient removal
.
Environ. Sci.
3
,
784
794
.
Mousavi
S.
,
Ibrahim
S.
,
Aroua
M. K.
&
Ghafari
S.
2012
Development of nitrate elimination by autohydrogenotrophic bacteria in bio-electrochemical reactors
.
Biochem. Eng. J.
67
(
34
),
251
264
.
Ni
S. Q.
,
Lee
P. H.
,
Fessehaie
A.
,
Gao
B. Y.
&
Sung
S. W.
2010
Enrichment and biofilm formation of Anammox bacteria in a non-woven membrane reactor
.
Bioresour. Technol.
101
,
1792
1799
.
Showers
W. J.
,
Genna
B.
,
McDade
T.
,
Bolich
R.
&
Fountain
J. C.
2008
Nitrate contamination in groundwater on an urbanized dairy farm
.
Environ. Sci. Technol.
42
,
4683
4688
.
Sorokin
D. Y.
,
Kuenen
J. G.
&
Muyzer
G.
2011
The microbial sulfur cycle at extremely haloalkaline conditions of soda lakes
.
Front. Microbiol.
2
(
1
),
111
117
.
Straub
K. L.
,
Schönhuber
W. A.
,
Buchholz-Cleven
B. E.
&
Schink
B.
2004
Diversity of ferrous iron-oxidizing: nitrate-reducing bacteria and their involvement in oxygen-independent iron cycling
.
Geomicrobiol. J.
21
,
371
378
.
Su
J. F.
,
Zheng
S. C.
,
Huang
T. L.
,
Ma
F.
,
Shao
S. C.
,
Yang
S. F.
&
Zhang
L. N.
2015
Characterization of the anaerobic denitrification bacterium Acinetobacter sp. SZ28 and its application for groundwater treatment
.
Bioresour. Technol.
192
,
654
659
.
Su
J. F.
,
Cheng
C.
,
Huang
T. L.
,
Ma
F.
,
Lu
J. S.
&
Shao
S. C.
2016a
Novel simultaneous Fe(III) reduction and ammonium oxidation of sp. FC61 under the anaerobic conditions
.
Rsc Advances
6
(
15
),
12584
12591
.
Tsushima
I.
,
Ogasawara
Y.
,
Kindaichi
T.
,
Satoh
H.
&
Okabe
S.
2007
Development of high-rate anaerobic ammonium-oxidizing (anammox) biofilm reactors
.
Water Res.
41
,
1623
1634
.
Yang
J. Y.
,
Yang
X. E.
,
He
Z. L.
,
Li
T. Q.
,
Shentu
J. L.
&
Stoffella
P. J.
2006
Effects of pH, organic acids, and inorganic ions on lead dissolution from soils
.
Environ. Poll.
143
,
9
15
.
Zhou
W.
,
Sun
Y.
,
Wu
B.
,
Zhang
Y.
,
Huang
M.
,
Miyanaga
T.
&
Zhang
Z.
2011
Autotrophic denitrification for nitrate and nitrite removal using sulfur-limestone
.
Environ. Sci.
23
(
11
),
1761
1769
.