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Efficient nitrous oxide recovery from incineration leachate by a nosZ-deficient strain of Pseudomonas aeruginosa
Journal Pre-proofs Efficient nitrous oxide recovery from incineration leachate by a nosZ-deficient strain of Pseudomonas aeruginosa Hanbing Nie, Xinying Liu, Yan Dang, Yanan Ji, Dezhi Sun, Jessica A. Smith, Dawn E. Holmes PII: DOI: Reference:
S0960-8524(19)31601-3 https://doi.org/10.1016/j.biortech.2019.122371 BITE 122371
To appear in:
Bioresource Technology
Received Date: Revised Date: Accepted Date:
18 September 2019 1 November 2019 2 November 2019
Please cite this article as: Nie, H., Liu, X., Dang, Y., Ji, Y., Sun, D., Smith, J.A., Holmes, D.E., Efficient nitrous oxide recovery from incineration leachate by a nosZ-deficient strain of Pseudomonas aeruginosa, Bioresource Technology (2019), doi: https://doi.org/10.1016/j.biortech.2019.122371
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1
Efficient nitrous oxide recovery from incineration leachate by a nosZ-deficient
2
strain of Pseudomonas aeruginosa
3 4 5 6
Hanbing Niea, Xinying Liua, Yan Danga, Yanan Jia, Dezhi Suna*, Jessica A. Smithb, Dawn E. Holmesc
7 8 9 10 11 12 13
a
Beijing Key Laboratory for Source Control Technology of Water Pollution,
14
Engineering Research Center for Water Pollution Source Control and
15
Eco-remediation, College of Environmental Science and Engineering, Beijing
16
Forestry University, Beijing, 100083, China
17 18 19 20
b
Department of Biomolecular Sciences, Central Connecticut State University,
1615 Stanley Street, New Britain, CT 06050, USA c
Department of Physical and Biological Sciences, Western New England
University, 1215 Wilbraham Rd, Springfield, MA 01119, United States
21 22
*Corresponding authors
23
Dezhi Sun, College of Environmental Science & Engineering, Beijing
24
Forestry University, 35 Tsinghua East Road, Beijing 100083, China. (E-mail:
25
[email protected])
26
27
Abstract
28
In this study, nitrous oxide was recovered from a lab-scale moving-bed biofilm
29
reactor (MBBR) treating partial nitrification-treated leachate supplemented with a
30
nosZ-deficient strain of Pseudomonas aeruginosa. Batch culture tests with the
31
nosZ-deficient strain determined that the threshold for free nitrous acid (FNA)
32
inhibition was 0.016 mg/L and that FNA concentrations above this threshold severely
33
inhibited denitrification and transcription of genes from the dissimilatory nitrate
34
reduction pathway (narG, nirS, and norB).
35
efficiencies (> 95%) were achieved with long-term operation of this MBBR. N2O
36
accounted for the majority of biogas (80%) produced when the MBBR was fed partial
37
nitrification-treated leachate with high nitrite concentrations and the drainage ratio
38
was adjusted to 30%. Bacterial community analysis revealed that the nosZ-deficient
39
Pseudomonas strain remained metabolically active and was primarily responsible for
40
denitrification processes in the reactor. This study presents a promising method for
41
N2O recovery from incineration leachate.
High nitrite removal and N2O conversion
42 43
Keywords: incineration leachate, nitrous oxide, energy recovery, free nitrous acid,
44
Pseudomonas aeruginosa PAO1
45
1. Introduction
46
Nitrous oxide (N2O) is one of the most potent greenhouse gases (Stocker et al.,
47
2013) and is considered one of the largest ozone depleting emissions (Ravishankara et
48
al., 2009). It is also an intermediate generated during biological nitrogen removal
49
processes carried out at wastewater treatment plants (WWTPs) (Kampschreur et al.,
50
2009; Schreiber et al., 2012). A significant amount of research has focused on
51
reducing N2O production and emission from WWTPs (Frutos et al., 2018; Leix et al.,
52
2017; Massara et al., 2017; Zheng et al., 2015). However, N2O is also a strong
53
oxidizer that can serve as a potential energy source in co-combustion with other fuels
54
such as methane generated during anaerobic digestion (Zhang et al., 2019). Therefore,
55
recent studies are also trying to identify methods to harvest N2O from WWTPs (Lin et
56
al., 2018; Scherson et al., 2013; Scherson et al., 2014).
57
In
a
typical
WWTP,
ammonia
in
wastewater
is
removed
through
58
nitrification-denitrification processes in which ammonia is oxidized to nitrate or
59
nitrite followed by nitrate/nitrite reduction to nitrogen gas (N2). During denitrification,
60
N2 is formed through a series of enzymatic steps with N2O as an intermediate
61
(Holmes et al., 2019). Not all microorganisms have all of the enzymes required for
62
complete reduction of nitrate to N2, and the end-product of some of these incomplete
63
denitrifiers is N2O. Therefore, methods are being developed to stimulate the growth of
64
N2O-forming denitrifiers, while inhibiting the growth of complete denitrifiers. For
65
example, a process called Coupled Aerobic–anoxic Nitrous Decomposition Operation
66
(CANDO) involves partial nitrification of NH4+ to NO2-, which is then anaerobically
67
reduced to N2O by incomplete denitrifiers (Scherson et al., 2013).
68
Other methods that increase N2O production involve inhibition of the enzyme
69
that catalyzes the conversion of N2O to N2, nitrous oxide reductase (NosZ), with H2S,
70
Cu2+, or free nitrous acid (FNA) (Magalhaes et al., 2011; Pan et al., 2013; Wang et al.,
71
2018b). Both CANDO and NosZ inhibition methods increase N2O concentrations in
72
reactors. However, the high solubility of N2O in water means that N2O only accounts
73
for a small proportion of biogas being produced (Zhang et al., 2019). One study
74
showed that addition of poly-propylene microporous hollow fiber membranes to
75
CANDO reactors significantly increased N2O in the gas phase (Weißbach et al.,
76
2018a). However, this process still needs to be optimized before use in commercial
77
applications.
78
Two approaches should help achieve higher N2O production rates and increase
79
the proportion of N2O found in biogas: (1) increasing denitrifying loads in the
80
wastewater, and (2) increasing the proportion of N2O-forming incomplete denitrifiers
81
in the bacterial community. In a previous study, a moving-bed biofilm reactor (MBBR)
82
fed synthesized wastewater was inoculated with a nosZ-deficient strain of
83
Pseudomonas aeruginosa PAO1 (Lin et al., 2018). This reactor achieved a N2O
84
conversion ratio of 73%, and N2O accounted for 73-81% of the biogas being
85
generated. Although this study successfully increased N2O available in biogas, it was
86
done in reactors fed autoclaved synthetic wastewater. In order to determine whether
87
this organism has real-world applications, N2O production by the nosZ-deficient strain
88
grown in actual wastewater needs to be tested.
89
One type of wastewater that has high denitrifying loads is fresh incineration
90
leachate formed from municipal solid waste (MSW) incineration plants (Dang et al.,
91
2016; Dang et al., 2013; Lei et al., 2016). Fresh incineration leachate contains
92
extremely high concentrations of ammonia (600-1200 mg/L) (Lei et al., 2018) that is
93
readily converted to nitrite through a partial nitrification process (Liu et al., 2019b).
94
While nitrite can serve as an electron acceptor during denitrification to N2O, its
95
protonated form, free nitrous acid (FNA), inhibits microbial activity and impairs the
96
denitrification process (Gao et al., 2016; Hartop et al., 2017; Ye et al., 2010; Zeng et
97
al., 2016). Therefore, after partial nitrification of incineration leachate to nitrite, it is
98
necessary to adjust the pH to minimize the formation of FNA.
99
For these purposes, an MBBR treating partial nitrification-treated MSW
100
incineration-leachate was inoculated with the nosZ-deficient strain of P. aeruginosa
101
and the partial denitrification process was monitored. Thresholds for inhibition of
102
partial denitrification by FNA were identified and optimal parameters for N2O
103
production from incineration leachate were determined.
104
2. Materials and Methods
105
2.1 Experimental strain and media
106
The nosZ-deficient strain of Pseudomonas aeruginosa PAO1 was previously
107
constructed (Lin et al., 2018) and obtained from our laboratory culture collection.
108
Prior to inoculation into the MBBR, cells were grown at 37˚C for 24h under
109
denitrifying conditions in modified LB minimal medium with 100 mg/L NO2--N. For
110
batch experiments, 800 mL denitrifying culture were centrifuged at 8000 rpm for 5
111
min and resuspended in 5 ml sterile water prior to inoculation, while 200 mL of
112
denitrifying culture was inoculated into the MBBR.
113
Fresh leachate was collected from an MSW-energy incineration plant in Beijing,
114
China and stored at 4˚C. This leachate was first treated in an anaerobic digester to
115
decrease chemical oxygen demand (COD), and then treated by partial nitrification to
116
convert ~1000 mg/L NH4+-N to ~1000 mg/L NO2-N. The product generated by both
117
of these treatment steps was then used as the feedstock for this study (Liu et al., 2018).
118
Characteristics of anaerobically treated leachate and partial nitrification-treated
119
leachate are summarized in Table 1.
120
2.2 Batch-experiment design
121
Batch tests were carried out to study the effects of FNA on activity of the
122
nosZ-deficient Pseudomonas aeruginosa PAO1 strain. 50 mL of culture were grown
123
anaerobically in serum bottles with medium composed of NaH2PO4 (150 mg/L),
124
CaCl2 (100 mg/L), MgSO4 · 7H2O (40 mg/L), FeSO4 · 7H2O (20 mg/L), and yeast
125
extract (40 mg/L). Stock solutions of sodium acetate (0.98 M), sodium nitrite (0.9 M),
126
and pH regulators (1.0 N NaOH and 1.0 N HCl) were added to the medium to achieve
127
concentrations of 250 mg/L NO2--N, 1300 mg/L COD, and 0.02-0.2mg/L FNA. Each
128
bottle was incubated at 30˚C and continuously stirred at 100 rpm by a magnetic
129
stirring apparatus.
130
Concentrations of NO2--N, N2O, and COD were measured every 6h until NO2--N
131
concentrations declined to nearly 0 mg/L. FNA concentrations were calculated from
132
NO2--N, temperature, and pH measurements using the following equation (Eq. 1).
133
47
ρ(FNA) = 14
ρ(NO2- - N) EXP( - 2300/(T + 273)) × 10pH
(1)
134
The biomass of the nosZ-deficient strain was determined at the beginning and end
135
of the batch experiment using a linear regression curve constructed by plotting dry
136
weight (mg/L) against absorbance at a wavelength of 600 nm (OD600) (see
137
Supplementary Material).
138
Cells were harvested from the serum bottles and quantitative reverse
139
transcription-polymerase chain reaction (qRT-PCR) was conducted with primers
140
targeting denitrification genes in order to determine the impact that FNA could have
141
on the relative expression of genes involved in denitrification.
142
2.3 RNA extraction and quantitative reverse transcription PCR
143
RNA was collected from triplicate bottles exposed to different FNA
144
concentrations. Cells were split into 50 ml conical tubes (BD Biosciences, San Jose,
145
CA) and pelleted by centrifugation at 8000 rpm for 5 min (Liu et al., 2019a). RNA
146
was then extracted from cell pellets with the RNAprep Bacteria kit (TianGen, China)
147
and treated with DNA-free DNase (Ambion) according to the manufacturer’s
148
instructions. All RNA samples were checked for integrity by agarose gel
149
electrophoresis and had A260/A280 ratios between 2.11 to 2.16 determined by a
150
Nanodrop UV spectrophotometer (Thermo Fisher Scientific, Wilmington, DE, USA).
151
PCR was conducted with primer sets from this study with RNA that had not been
152
reverse transcribed serving as the template to ensure that the RNA samples were not
153
contaminated with DNA.
154
Complementary DNA (cDNA) was generated from total extracted RNA (1μg)
155
with the Fast Quant Reverse Transcriptase Kit (TianGen, China) according to the
156
manufacturer’s instructions. cDNA products were then used as templates for
157
qRT-PCR with previously described primers targeting narG, napA, nirS, norB, nirB,
158
nosZ and constitutively expressed proC genes (Lin et al., 2018). qRT-PCR was
159
performed in a real-time PCR system (7500 FAST, USA) using the manufacturer’s
160
guidelines. Each PCR mixture consisted of a total volume of 25 µL and contained 1.5
161
µL of the appropriate primers (stock concentrations, 15 µM) and 12.5 µL SYBR
162
green PCR master mix (Bio-Rad).
163
2.4 Long-term operation reactor design
164
To
determine
the
sustainability
of
N2O
production
from
partial
165
nitrification-treated leachate inoculated with the nosZ-deficient strain, a sequencing
166
batch MBBR with a working volume of 1 L filled with a 50% polypropylene filter
167
(v/v) was operated for more than 85 cycles. The reactor was inoculated with 100 mL
168
of ∆nosZ cells (OD600 of 0.82) and operated at 30 ± 1˚C. Filters were added to the
169
MBBR to enhance biofilm formation, which increased nosZ-deficient strain biomass
170
in the system.
171
Each operation cycle was 24 hours with the feeding period lasting 5 minutes, the
172
anaerobic stir period lasting 23.5 hours, and the settling and decant period lasting 25
173
minutes. There were three stages of operation. During the first stage (cycles 1-15), the
174
nosZ-deficient strain formed a biofilm on the filter in the reactor. The nosZ-deficient
175
strain then acclimated and started to produce high concentrations of nitrous oxide
176
during stage 2 (cycles 16-55). In the first two stages, the bioreactor was being fed
177
diluted partial nitrification-treated leachate with a drainage ratio set at 100%. The
178
NO2--N concentration in the influent was increased incrementally from 100 to 300
179
mg/L. In the third stage (cycles 56-85), the bioreactor was fed non-diluted partial
180
nitrification-treated incineration leachate (containing 1006-1090 NO2--N mg/L). The
181
drainage ratio at this stage was set at 30% in order to adjust initial NO2--N
182
concentrations at the beginning of each cycle to the same concentration (~300 mg/L).
183
Sodium acetate was provided as the electron donor for denitrification throughout the
184
operational period and the COD/N was maintained at 5.0.
185
Throughout operation, NO2--N, liquid phase N2O, gaseous phase N2O, and total
186
N2O were monitored to determine nitrogen removal efficiencies and N2O production
187
rates. The proportion of N2O in biogas was also determined to ensure the feasibility of
188
N2O recovery and bioreactor energy recovery was calculated from the co-combustion
189
of methane with produced nitrous oxide.
190
2.5 DNA extraction and sequence analysis
191
DNA was extracted from samples collected from the MBBR during cycles 1, 55,
192
and 85 with the E.Z.N.A. Soil DNA Kit (Omega Bio-Tek, Inc., Norcross, GA, USA)
193
according to the manufacturer’s instructions. A Nanodrop UV spectrophotometer
194
(Thermo Fisher Scientific, Delaware) was used to determine the DNA concentration
195
and purity of each sample.
196
Bacterial 16S rRNA gene fragments were amplified by PCR with the 338F/806R
197
primer set. Amplicons were sequenced on an Illumina Hiseq 2000 platform (Illumia,
198
San Diego, USA) by Majorbio Bio-Pharm Technology Co. Ltd. (Shanghai, China).
199
Sequences were placed into various operational taxonomic units with Pyrosequencing
200
Pipeline software (https://pyro.cme.msu.edu). Raw sequence files have been
201
submitted to the NCBI Sequence Read Archive database under accession NO.
202
PRJNA558633.
203
2.6 Analytical methods
204
NO2--N, NO3--N, NH4+-N, and COD were determined by standard methods
205
(APHA., 2005). Biomass was determined with a spectrophotometer (UV-1800,
206
Shimadzu, Japan) set at a wavelength of 600 nm. The pH was measured with a HACH
207
sensION+pH3®pH meter (HACH, USA). The N2O portion of biogas was analyzed
208
with a gas chromatograph (Agilent 7890A) equipped with an electron capture detector
209
(ECD). N2O dissolved in the liquid phase was calculated using Henrys law based on a
210
previously described method (Wang et al., 2018a). Statistical differences were
211
determined by one-way analysis of variance (ANOVA), with p＜0.05 considered to
212
be statistically significant.
213
3. Results and Discussion
214
3.1 Effects of FNA on nitrite removal and N2O conversion by the nosZ-deficient
215
mutant strain
216
A batch experiment was conducted to determine the threshold of FNA inhibition
217
and to identify optimal FNA concentrations for nitrite removal by the nosZ-deficient
218
P.
219
strain was grown in medium containing 0.206 mg/L initial FNA at pH 7.0, 0.042
220
mg/L initial FNA at pH 7.5, or 0.016 mg/L initial FNA at pH 8.0. Marked differences
221
in nitrite conversion rates were observed between the three conditions (Fig. 1a).
222
Cultures amended with an initial FNA concentration of 0.016 mg/L had the highest
aeruginosa strain during the partial denitrification process. The nosZ-deficient
223
nitrite conversion rate (6.77 mgN/(L ·h)), while nitrite reduction was minimal in the
224
other conditions. The bottles supplemented with 0.016 mg/L also formed the highest
225
concentrations of total nitrous oxide (0.398 ± 0.022 mM), while bottles grown under
226
the other conditions produced little N2O (~0.05 ± 0.022 mM) (Fig. 1b). Similar to the
227
nosZ-deficient P. aeruginosa strain, denitrification by Accumulibacter was also
228
completely inhibited by FNA concentrations of 0.2 mg/L (Zhou et al., 2010).
229
COD concentrations also decreased most rapidly in the 0.016 mg/L FNA bottles
230
during the first 18 hours of the experiment; 25.05 mg/(L·h) compared to 11.67
231
mg/(L·h) and 11.17 mg/(L · h) in the 0.042 mg/L FNA and the 0.206 mg/L FNA
232
bottles, respectively (Fig. 1c). These results are consistent with the fact that most of
233
the electron acceptor (NO2-) had been utilized by 18 h in the 0.016 mg/L FNA bottle
234
meaning that acetate could no longer serve as an electron donor for denitrification.
235
Therefore, the only way that the nosZ-deficient strain could support growth at this
236
point was via fermentation of compounds present in the yeast extract as this organism
237
cannot ferment acetate (Eschbach et al., 2004; Glasser et al., 2014; Vander Wauven et
238
al., 1984).
239
Although COD consumption in the 0.016 mg/L FNA bottle almost completely
240
leveled off by 18 h, CODs continued to steadily decline until 30 h for the 0.206 mg/L
241
FNA bottle and 36 h for the 0.042 mg/L FNA bottle (Fig. 1c). This is likely because
242
nitrite was still available as an electron acceptor at this point and although
243
denitrification
244
denitrification could support some growth. However, it appears that the majority of
245
COD consumption in the bottles with elevated FNA concentrations could be
246
attributed to fermentation and biomass conversion pathways.
enzymes
were
impaired
by
elevated
FNA
concentrations,
247
FNA concentrations in the 0.016 mg/L FNA bottle dropped down to only 0.0025
248
mg/L after 6 h and concentrations were negligible by 36 h (Fig. 1d). During the first
249
18 hours of the experiment, pH in the 0.042 mg/L bottle decreased from 7.58 to 7.11
250
which caused FNA concentrations to increase from 0.042 mg/L to 0.122 mg/L. This
251
drop in pH was most likely caused by an increase in acidic by-products formed from
252
the fermentation of yeast extract (CO2 and fatty acids) by Pseudomonas aeruginosa.
253
After 18 hours, however, FNA concentrations went back down to 0.042 mg/L and
254
this decline in FNA corresponded with an increase in nitrite consumption (Fig. 1a)
255
and nitrous oxide production (Fig. 1b). This suggests that cells started to participate in
256
denitrification once FNA concentrations had dropped. In the 0.206 mg/L FNA bottle,
257
pH was maintained at 6.95 ± 0.06 for the entire 30 h experiment (see Supplementary
258
Material). FNA concentrations in the 0.206 mg/L bottle also remained above 0.148
259
mg/L and these concentrations inhibited growth of P. aeruginosa as evidenced by the
260
finding that little denitrification or fementation occurred in this bottle.
261
Consistent with the finding that the nosZ-deficient mutant was only able to grow
262
efficiently in the 0.016 mg/L FNA bottle, biomass measurements actually declined in
263
the 0.206 mg/L FNA bottle, only slightly increased in the 0.042 mg/L FNA bottle, and
264
significantly increased in the 0.016 mg/L FNA bottle over the course of the
265
experiment (see Supplementary Material).
266
3.2 Effect of FNA on expression of denitrification pathway genes
267
In order to explore the influence of FNA on transcription patterns of
268
denitrification genes, qRT-PCR was done with primers targeting various genes from
269
the dissimilatory nitrate reduction pathway (Fig. 2). The results clearly demonstrated
270
that transcription of denitrification genes was severely inhibited by elevated FNA
271
concentrations. For example, 7.8 and 146.6 times more mRNA transcripts from the
272
gene that codes for the alpha subunit of nitrate reductase/nitrite oxidoreductase (narG)
273
were being expressed in cells grown in the 0.016 mg/L FNA bottle than the 0.042
274
mg/L FNA and 0.206 mg/L FNA bottles, respectively (Fig. 2). The protein encoded
275
by narG can catalyze both the reduction of nitrate to nitrite and the oxidation of nitrite
276
to nitrate (Gonzalez et al., 2006), and the narGHI operon is regulated by sensor
277
proteins that are activated by either nitrate or nitrite (Noriega et al., 2010). Studies
278
have also shown that elevated FNA concentrations can inhibit enzymes from the
279
denitrification pathway (Fudala-Ksiazek et al., 2014). This explains why narG was
280
being expressed in nitrite-grown cells and why the number of narG mRNA transcripts
281
was higher in the FNA 0.016 mg/L bottle than the FNA 0.042 mg/L or 0.206 mg/L
282
bottles.
283
Genes coding for the NO-forming nitrite reductase (nirS) and the beta subunit of
284
nitric oxide reductase (norB) were also more significantly expressed in the 0.016
285
mg/L FNA bottle. Cells from the 0.016 mg/L FNA bottle expressed 8.6 and 38.4
286
times more nirS mRNA transcripts and 6.5 and 35.5 times more norB transcripts than
287
the 0.042 mg/L FNA and the 0.206 mg/L FNA bottles.
288
No transcripts for genes coding for periplasmic nitrate reductase (napA) or the
289
large subunit from assimilatory nitrite reductase (nirB) were detected in the three
290
bottles. These results are consistent with previous studies that have shown that napA
291
is expressed under aerobic and microaerophilic conditions and that nirB transcription
292
is only induced under high nitrate concentrations (Lin et al., 2018). In addition, the
293
nosZ gene was not expressed, confirming the purity of the nosZ-deficient mutant
294
strain used in these studies.
295
3.3 Performance of the MBBR treating effluent from partial nitrification of
296
incineration leachate and supplemented with the nosZ-deficient P. aeruginosa
297
strain.
298
Once optimal FNA concentrations for growth of the nosZ-deficient strain were
299
identified in batch experiments, tests were done to determine whether this strain could
300
survive in long-term sequencing batch bioreactors treating effluent from partial
301
nitrification of incineration leachate, and whether N2O generated by this strain could
302
be recovered efficiently during long-term operation.
303
The denitrification MBBR received 200 mL of suspended mutant strain cells
304
(OD600 of 0.82) cultured in LB media. The reactor was operated for 85 cycles with
305
each cycle duration set at 24 h per cycle (Fig. 3). During the first stage of the
306
experiment (Stage I: cycle 1-15), NO2--N concentrations were maintained at ~100
307
mg/L and COD concentrations ranged from 500 to 550 mg/L to ensure biofilm
308
formation. During this stage, the nosZ-deficient strain gradually acclimated to
309
denitrifying conditions, and NO2--N removal efficiencies gradually increased from
310
4.85% to 55.58% for each cycle. By the end of Stage I (cycle 15), NO2--N removal
311
efficiencies were greater than 60% (Fig. 3).
312
During the second stage of the experiment (Stage II: cycles 16-55), NO2--N
313
concentrations ranged from 100 to 300 mg/L and COD concentrations ranged from
314
500-1500 mg/L to keep the COD/N at ~5.0. NO2--N removal efficiencies increased
315
further (85% to 99%), and became relatively stable after cycle 21. Approximately 2.7
316
mmol/d, 4.6 mmol/d, 6.9 mmol/d, and 10.3 mmol/d of N2O were generated by the
317
experimental reactor when it was provided with NO2--N concentrations of 100 mg/L,
318
150 mg/L, 200 mg/L, and 300 mg/L, respectively. The majority (＞70%, Fig. 3) of
319
N2O being generated in the experimental reactor during Stage II was dissolved in the
320
aqueous phase. N2O gas only accounted for 1.24-9.22% of the biogas being produced
321
during this stage (Fig. 4). N2O conversion efficiencies ranged from 66.25% to 99.77%
322
in stage II, indicating that the reactor was operating stably. A thick biofilm had also
323
formed on the bio-filter by the end of this stage (see Supplementary Material).
324
In the third stage of operation, raw partial nitrification-treated leachate with
325
~1000 mg/L NO2--N and ~5000 mg/L COD was fed into the MBBR, and the drainage
326
ratio was changed from 100% to 30% (Stage III: cycles 56-85). NO2--N efficiencies
327
were maintained at ~99% throughout Stage III, and 10.6 mmol/d of N2O was
328
generated when the reactor was provided with NO2--N concentrations of 300 mg/L.
329
The majority (>90%, Fig. 3) of N2O being produced during Stage III was found in the
330
gas phase, and N2O gas accounted for 71.52-83.57% of the biogas (Fig. 4). N2O
331
conversion efficiencies continued to be stable and ranged from 95.44 to 99.64%
332
during this final stage, significantly higher than previous studies that achieved
333
efficiencies of 80% (Table 2).
334
The proportion of N2O in biogas generated by the MBBR reactor increased
335
significantly during Stage III (after cycle 56) (Fig. 4). Prior to Stage III, N2O only
336
accounted for 15% of the biogas. However, after cycle 56, when the drainage ratio
337
was decreased from 100% to 30%, the proportion of N2O in biogas increased to 80%.
338
This can be explained by the fact that the MBBR reactor discharged 70% less leachate
339
after each cycle and that the influent was changed from diluted (<300 mg/L NO2--N)
340
to non-diluted (~1000 mg/L NO2--N) leachate (although after thoroughly stirring,
341
NO2--N concentrations at the beginning of each cycle remained at ~300 mg/L (1000 *
342
30% / 100%)). Under these conditions, the non-discharged leachate was already
343
saturated with dissolved N2O, and this left little liquid volume for any newly
344
generated N2O, concentrating N2O in the gas phase.
345
The equation of co-combustion with methane and nitrous oxide (Eq. 2) was used
346
to quantify energy recovered in N2O from the MBBR reactor. A significant amount of
347
energy (average 9206 kJ/m3) was recovered during Stage III (cycles 56 to 85), but
348
little energy (average 475 kJ/m3) was recovered during cycles 1 to 55.
349 350
°
CH4 + 4N2O →CO2+2H2O + 4N2, ∆HC = - 1219 kJ/mol
(2)
3.4 Bacterial community analysis
351
In order to ensure that microorganisms already present in the partial
352
nitrification-treated leachate used as influent for the MBBR system did not
353
competitively exclude the nosZ-deficient Pseudomonas strain, bacterial community
354
structure in the MBBR was assessed at different periods of operation (cycles 1, 55,
355
and 85) (Fig. 6). The nosZ-deficient Pseudomonas strain dominated the community at
356
the beginning of the experiment and accounted for 99.7% of the bacterial sequences
357
detected after cycle 1. As the experiment progressed, the MBBR bacterial community
358
became more diverse and Pseudomonas only accounted for 38.35% and 24.77% of
359
the bacterial sequences after cycles 55 and 85.
360
Tissierella became the most significant organism and accounted for 55.86%-58.63%
361
of the population after cycle 55. This genus has been isolated from human sewage
362
treatment plants (Harms et al., 1998) and has been a dominant member of bacterial
363
communities associated with several anaerobic digesters treating complex waste
364
(Jaenicke et al., 2011; Liu et al., 2009; Nolla-Ardèvol et al., 2015; Snell-Castro et al.,
365
2005; Ziganshina et al., 2014). Tissierella can metabolize nitrogen-containing
366
compounds but is not capable of denitrification (Harms et al., 1998). Its metabolism is
367
also stimulated by the presence of formate (Harms et al., 1998) which is a
368
fermentation by-product formed by P. aeruginosa (Eschbach et al., 2004). Therefore,
369
it is possible that production of formate by Pseudomonas helped stimulate growth of
370
Tissierella, however further investigation into this possibility is required.
371
Other bacterial genera that started to appear later in the experiment included
372
Muricomes, Alkaliphilus, Jeotaglibaca, and Clostridium (Fig. 6). Clostridium is
373
known to be capable of dissimilatory nitrate reduction to ammonia (DNRA) (Caskey
374
& Tiedje, 1980; Keith et al., 1982) and genes for ammonia forming nitrite reductase
375
(nrfAH) were found in Tissierella and Alkaliphilus genomes. The role that these other
376
microorganisms were playing in the reactor is unclear. However, it is apparent that
377
conditions in the reactor did not promote growth of traditional complete or incomplete
378
denitrifiers that could generate N2O. Therefore, it appears that under the conditions
379
used in this study, reactors need to be supplemented with N2O producing bacteria
380
such as the nosZ-deficient strain of P. aeruginosa to generate high concentrations of
381
N2O throughout the experiment. These findings are significant and show that
382
generation of N2O by the nosZ-deficient strain of P. aeruginosa and potentially other
383
denitrifying strains have real applications within the wastewater treatment field.
384 385
4. Conclusions
386
Batch experiments revealed that FNA concentrations above 0.016 mg/L severely
387
inhibited denitrification by the nosZ-deficient strain of P. aeruginosa. Addition of this
388
strain to an MBBR dramatically improved performance and resulted in nitrite removal
389
and N2O conversion efficiencies >95%. Adjustment of the drainage ratio to 30%
390
increased the proportion of N2O in biogas to levels above 80%. Bacterial community
391
analyses revealed that the nosZ-deficient Pseudomonas was primarily responsible for
392
N2O formation in the bioreactor throughout the experiment. These results demonstrate
393
that high concentrations of N2O can be recovered from incineration leachate in
394
MBBR supplemented with the nosZ-deficient Pseudomonas strain.
395 396
Note
397
The authors declare no competing interest.
398 399
Appendix A. Supplementary Data
400
E-supplementary data for this work can be found in e-version of this paper online.
401 402
Acknowledgments
403
This research was financially supported by the National Natural Science
404
Foundation
405
Natural Science Foundation (8184081) and the Major Science and Technology
406
Program for Water Pollution Control and Treatment (2017ZX07108-002).
of
China
(51678051,
51708031),
the Beijing
Municipal
407 408 409 410
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550
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554
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555
556 557
Figure captions Fig. 1. Growth of the nosZ-deficient strain of P. aeruginosa PAO1 under
558
denitrifying conditions with acetate (19.5 mM) provided as an electron donor. Cells
559
were exposed to three different initial free nitrous acid (FNA) concentrations. Nitrite
560
concentrations (a); total nitrous oxide (from liquid and gaseous phases) (b); chemical
561
oxygen demand (COD) (c), and free nitrous acid (FNA) (d) were monitored over the
562
course of 36 hours. Error bars were calculated from triplicate samples.
563
Fig. 2. The number of narG, napA, nirS, norB, nirB and nosZ mRNA transcripts
564
normalized against the number of proC mRNA transcripts detected in the
565
nosZ-deficient P. aeruginosa cells exposed to different FNA concentrations. Error
566
bars were calculated from triplicate samples.
567 568 569 570
Fig. 3. Performance of the MBBR supplemented with the nosZ-deficient P. aeruginosa strain. Fig. 4. Proportion of N2O in biogas generated by the MBBR supplemented with the nosZ-deficient P. aeruginosa strain.
571
Fig. 5. Energy recovered (calculated by nitrous oxide combustion with methane)
572
from nitrous oxide generated by the nosZ-deficient P. aeruginosa-supplemented
573
MBBR system.
574 575
Fig. 6. Proportion of various genera detected in the MBBR system supplemented with nosZ-deficient P. aeruginosa after cycles 1, 55, and 85.
576 577
Table captions
578
Table 1 Characteristics of experimental wastewater used in this study
579
Table 2 Nitrogen removal efficiencies, N2O production rates, and N2O
580
conversion efficiencies obtained with the method used in this study compared to other
581
approaches used for N2O recovery.
582
Table 1. Characteristics of experimental wastewater used in this study COD
BOD5
NH4+-N
NO3--N
NO2--N
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
1980-3660
850-2380
1140-1420
1124-1545
125-342
202-310
Item
pH
AnaerobicallyN/A
N/A
8.05-8.55
1006-1090
7.12-7.53
treated leachate Partial nitrification treated leachate 583
18.28-24.65
584
Table 2. Nitrogen removal efficiencies, N2O production rates, and N2O conversion
585
efficiencies obtained with the method used in this study compared to other approaches
586
used for N2O recovery. Nitrogen Study
N2O
N2O
removal
production rate
conversion
efficiency
(mgN/(gVSS·h))
efficiency
nosZ-deficient
This study
＞97%
119.7±5.7
＞95%
Pseudomonas
(Lin et al., 2018)
＞97%
106.1±4.9
70-80%
(Scherson et al., 2013)
＞98%
5.6
60-65%
(Scherson et al., 2014)
＞98%
25.2
75-80%
72%
2.1±0.4
65-75%
(Gao et al., 2017)
＞98%
5.1±1.6
70-80%
(Weißbach et al., 2018b)
＞95%
1.05 ± 0.20
53-63%
80%
2.38±0.26
50-65%
＞98%
N/A
71-73%
CANDO
NO inhibit Light-driven
(Myung et al., 2015)
(Yu et al., 2019) (Chen et al., 2019)
587 588 589
Highlights
590
Optimal FNA concentrations for the nosZ-deficient Pseudomonas were <0.02 mg/L
591
FNA inhibited transcription of genes from the denitrification pathway
592
N2O conversion efficiencies of MBBR supplemented with nosZ-deficient
593 594
strain were >95%
595 596 597 598 599
Drainage ratio adjustments increased the proportion of N2O in biogas to 80%
The nosZ-deficient strain remained the dominant denitrifier throughout MBBR operation
S0960-8524(19)31601-3 https://doi.org/10.1016/j.biortech.2019.122371 BITE 122371
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Bioresource Technology
Received Date: Revised Date: Accepted Date:
18 September 2019 1 November 2019 2 November 2019
Please cite this article as: Nie, H., Liu, X., Dang, Y., Ji, Y., Sun, D., Smith, J.A., Holmes, D.E., Efficient nitrous oxide recovery from incineration leachate by a nosZ-deficient strain of Pseudomonas aeruginosa, Bioresource Technology (2019), doi: https://doi.org/10.1016/j.biortech.2019.122371
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1
Efficient nitrous oxide recovery from incineration leachate by a nosZ-deficient
2
strain of Pseudomonas aeruginosa
3 4 5 6
Hanbing Niea, Xinying Liua, Yan Danga, Yanan Jia, Dezhi Suna*, Jessica A. Smithb, Dawn E. Holmesc
7 8 9 10 11 12 13
a
Beijing Key Laboratory for Source Control Technology of Water Pollution,
14
Engineering Research Center for Water Pollution Source Control and
15
Eco-remediation, College of Environmental Science and Engineering, Beijing
16
Forestry University, Beijing, 100083, China
17 18 19 20
b
Department of Biomolecular Sciences, Central Connecticut State University,
1615 Stanley Street, New Britain, CT 06050, USA c
Department of Physical and Biological Sciences, Western New England
University, 1215 Wilbraham Rd, Springfield, MA 01119, United States
21 22
*Corresponding authors
23
Dezhi Sun, College of Environmental Science & Engineering, Beijing
24
Forestry University, 35 Tsinghua East Road, Beijing 100083, China. (E-mail:
25
[email protected])
26
27
Abstract
28
In this study, nitrous oxide was recovered from a lab-scale moving-bed biofilm
29
reactor (MBBR) treating partial nitrification-treated leachate supplemented with a
30
nosZ-deficient strain of Pseudomonas aeruginosa. Batch culture tests with the
31
nosZ-deficient strain determined that the threshold for free nitrous acid (FNA)
32
inhibition was 0.016 mg/L and that FNA concentrations above this threshold severely
33
inhibited denitrification and transcription of genes from the dissimilatory nitrate
34
reduction pathway (narG, nirS, and norB).
35
efficiencies (> 95%) were achieved with long-term operation of this MBBR. N2O
36
accounted for the majority of biogas (80%) produced when the MBBR was fed partial
37
nitrification-treated leachate with high nitrite concentrations and the drainage ratio
38
was adjusted to 30%. Bacterial community analysis revealed that the nosZ-deficient
39
Pseudomonas strain remained metabolically active and was primarily responsible for
40
denitrification processes in the reactor. This study presents a promising method for
41
N2O recovery from incineration leachate.
High nitrite removal and N2O conversion
42 43
Keywords: incineration leachate, nitrous oxide, energy recovery, free nitrous acid,
44
Pseudomonas aeruginosa PAO1
45
1. Introduction
46
Nitrous oxide (N2O) is one of the most potent greenhouse gases (Stocker et al.,
47
2013) and is considered one of the largest ozone depleting emissions (Ravishankara et
48
al., 2009). It is also an intermediate generated during biological nitrogen removal
49
processes carried out at wastewater treatment plants (WWTPs) (Kampschreur et al.,
50
2009; Schreiber et al., 2012). A significant amount of research has focused on
51
reducing N2O production and emission from WWTPs (Frutos et al., 2018; Leix et al.,
52
2017; Massara et al., 2017; Zheng et al., 2015). However, N2O is also a strong
53
oxidizer that can serve as a potential energy source in co-combustion with other fuels
54
such as methane generated during anaerobic digestion (Zhang et al., 2019). Therefore,
55
recent studies are also trying to identify methods to harvest N2O from WWTPs (Lin et
56
al., 2018; Scherson et al., 2013; Scherson et al., 2014).
57
In
a
typical
WWTP,
ammonia
in
wastewater
is
removed
through
58
nitrification-denitrification processes in which ammonia is oxidized to nitrate or
59
nitrite followed by nitrate/nitrite reduction to nitrogen gas (N2). During denitrification,
60
N2 is formed through a series of enzymatic steps with N2O as an intermediate
61
(Holmes et al., 2019). Not all microorganisms have all of the enzymes required for
62
complete reduction of nitrate to N2, and the end-product of some of these incomplete
63
denitrifiers is N2O. Therefore, methods are being developed to stimulate the growth of
64
N2O-forming denitrifiers, while inhibiting the growth of complete denitrifiers. For
65
example, a process called Coupled Aerobic–anoxic Nitrous Decomposition Operation
66
(CANDO) involves partial nitrification of NH4+ to NO2-, which is then anaerobically
67
reduced to N2O by incomplete denitrifiers (Scherson et al., 2013).
68
Other methods that increase N2O production involve inhibition of the enzyme
69
that catalyzes the conversion of N2O to N2, nitrous oxide reductase (NosZ), with H2S,
70
Cu2+, or free nitrous acid (FNA) (Magalhaes et al., 2011; Pan et al., 2013; Wang et al.,
71
2018b). Both CANDO and NosZ inhibition methods increase N2O concentrations in
72
reactors. However, the high solubility of N2O in water means that N2O only accounts
73
for a small proportion of biogas being produced (Zhang et al., 2019). One study
74
showed that addition of poly-propylene microporous hollow fiber membranes to
75
CANDO reactors significantly increased N2O in the gas phase (Weißbach et al.,
76
2018a). However, this process still needs to be optimized before use in commercial
77
applications.
78
Two approaches should help achieve higher N2O production rates and increase
79
the proportion of N2O found in biogas: (1) increasing denitrifying loads in the
80
wastewater, and (2) increasing the proportion of N2O-forming incomplete denitrifiers
81
in the bacterial community. In a previous study, a moving-bed biofilm reactor (MBBR)
82
fed synthesized wastewater was inoculated with a nosZ-deficient strain of
83
Pseudomonas aeruginosa PAO1 (Lin et al., 2018). This reactor achieved a N2O
84
conversion ratio of 73%, and N2O accounted for 73-81% of the biogas being
85
generated. Although this study successfully increased N2O available in biogas, it was
86
done in reactors fed autoclaved synthetic wastewater. In order to determine whether
87
this organism has real-world applications, N2O production by the nosZ-deficient strain
88
grown in actual wastewater needs to be tested.
89
One type of wastewater that has high denitrifying loads is fresh incineration
90
leachate formed from municipal solid waste (MSW) incineration plants (Dang et al.,
91
2016; Dang et al., 2013; Lei et al., 2016). Fresh incineration leachate contains
92
extremely high concentrations of ammonia (600-1200 mg/L) (Lei et al., 2018) that is
93
readily converted to nitrite through a partial nitrification process (Liu et al., 2019b).
94
While nitrite can serve as an electron acceptor during denitrification to N2O, its
95
protonated form, free nitrous acid (FNA), inhibits microbial activity and impairs the
96
denitrification process (Gao et al., 2016; Hartop et al., 2017; Ye et al., 2010; Zeng et
97
al., 2016). Therefore, after partial nitrification of incineration leachate to nitrite, it is
98
necessary to adjust the pH to minimize the formation of FNA.
99
For these purposes, an MBBR treating partial nitrification-treated MSW
100
incineration-leachate was inoculated with the nosZ-deficient strain of P. aeruginosa
101
and the partial denitrification process was monitored. Thresholds for inhibition of
102
partial denitrification by FNA were identified and optimal parameters for N2O
103
production from incineration leachate were determined.
104
2. Materials and Methods
105
2.1 Experimental strain and media
106
The nosZ-deficient strain of Pseudomonas aeruginosa PAO1 was previously
107
constructed (Lin et al., 2018) and obtained from our laboratory culture collection.
108
Prior to inoculation into the MBBR, cells were grown at 37˚C for 24h under
109
denitrifying conditions in modified LB minimal medium with 100 mg/L NO2--N. For
110
batch experiments, 800 mL denitrifying culture were centrifuged at 8000 rpm for 5
111
min and resuspended in 5 ml sterile water prior to inoculation, while 200 mL of
112
denitrifying culture was inoculated into the MBBR.
113
Fresh leachate was collected from an MSW-energy incineration plant in Beijing,
114
China and stored at 4˚C. This leachate was first treated in an anaerobic digester to
115
decrease chemical oxygen demand (COD), and then treated by partial nitrification to
116
convert ~1000 mg/L NH4+-N to ~1000 mg/L NO2-N. The product generated by both
117
of these treatment steps was then used as the feedstock for this study (Liu et al., 2018).
118
Characteristics of anaerobically treated leachate and partial nitrification-treated
119
leachate are summarized in Table 1.
120
2.2 Batch-experiment design
121
Batch tests were carried out to study the effects of FNA on activity of the
122
nosZ-deficient Pseudomonas aeruginosa PAO1 strain. 50 mL of culture were grown
123
anaerobically in serum bottles with medium composed of NaH2PO4 (150 mg/L),
124
CaCl2 (100 mg/L), MgSO4 · 7H2O (40 mg/L), FeSO4 · 7H2O (20 mg/L), and yeast
125
extract (40 mg/L). Stock solutions of sodium acetate (0.98 M), sodium nitrite (0.9 M),
126
and pH regulators (1.0 N NaOH and 1.0 N HCl) were added to the medium to achieve
127
concentrations of 250 mg/L NO2--N, 1300 mg/L COD, and 0.02-0.2mg/L FNA. Each
128
bottle was incubated at 30˚C and continuously stirred at 100 rpm by a magnetic
129
stirring apparatus.
130
Concentrations of NO2--N, N2O, and COD were measured every 6h until NO2--N
131
concentrations declined to nearly 0 mg/L. FNA concentrations were calculated from
132
NO2--N, temperature, and pH measurements using the following equation (Eq. 1).
133
47
ρ(FNA) = 14
ρ(NO2- - N) EXP( - 2300/(T + 273)) × 10pH
(1)
134
The biomass of the nosZ-deficient strain was determined at the beginning and end
135
of the batch experiment using a linear regression curve constructed by plotting dry
136
weight (mg/L) against absorbance at a wavelength of 600 nm (OD600) (see
137
Supplementary Material).
138
Cells were harvested from the serum bottles and quantitative reverse
139
transcription-polymerase chain reaction (qRT-PCR) was conducted with primers
140
targeting denitrification genes in order to determine the impact that FNA could have
141
on the relative expression of genes involved in denitrification.
142
2.3 RNA extraction and quantitative reverse transcription PCR
143
RNA was collected from triplicate bottles exposed to different FNA
144
concentrations. Cells were split into 50 ml conical tubes (BD Biosciences, San Jose,
145
CA) and pelleted by centrifugation at 8000 rpm for 5 min (Liu et al., 2019a). RNA
146
was then extracted from cell pellets with the RNAprep Bacteria kit (TianGen, China)
147
and treated with DNA-free DNase (Ambion) according to the manufacturer’s
148
instructions. All RNA samples were checked for integrity by agarose gel
149
electrophoresis and had A260/A280 ratios between 2.11 to 2.16 determined by a
150
Nanodrop UV spectrophotometer (Thermo Fisher Scientific, Wilmington, DE, USA).
151
PCR was conducted with primer sets from this study with RNA that had not been
152
reverse transcribed serving as the template to ensure that the RNA samples were not
153
contaminated with DNA.
154
Complementary DNA (cDNA) was generated from total extracted RNA (1μg)
155
with the Fast Quant Reverse Transcriptase Kit (TianGen, China) according to the
156
manufacturer’s instructions. cDNA products were then used as templates for
157
qRT-PCR with previously described primers targeting narG, napA, nirS, norB, nirB,
158
nosZ and constitutively expressed proC genes (Lin et al., 2018). qRT-PCR was
159
performed in a real-time PCR system (7500 FAST, USA) using the manufacturer’s
160
guidelines. Each PCR mixture consisted of a total volume of 25 µL and contained 1.5
161
µL of the appropriate primers (stock concentrations, 15 µM) and 12.5 µL SYBR
162
green PCR master mix (Bio-Rad).
163
2.4 Long-term operation reactor design
164
To
determine
the
sustainability
of
N2O
production
from
partial
165
nitrification-treated leachate inoculated with the nosZ-deficient strain, a sequencing
166
batch MBBR with a working volume of 1 L filled with a 50% polypropylene filter
167
(v/v) was operated for more than 85 cycles. The reactor was inoculated with 100 mL
168
of ∆nosZ cells (OD600 of 0.82) and operated at 30 ± 1˚C. Filters were added to the
169
MBBR to enhance biofilm formation, which increased nosZ-deficient strain biomass
170
in the system.
171
Each operation cycle was 24 hours with the feeding period lasting 5 minutes, the
172
anaerobic stir period lasting 23.5 hours, and the settling and decant period lasting 25
173
minutes. There were three stages of operation. During the first stage (cycles 1-15), the
174
nosZ-deficient strain formed a biofilm on the filter in the reactor. The nosZ-deficient
175
strain then acclimated and started to produce high concentrations of nitrous oxide
176
during stage 2 (cycles 16-55). In the first two stages, the bioreactor was being fed
177
diluted partial nitrification-treated leachate with a drainage ratio set at 100%. The
178
NO2--N concentration in the influent was increased incrementally from 100 to 300
179
mg/L. In the third stage (cycles 56-85), the bioreactor was fed non-diluted partial
180
nitrification-treated incineration leachate (containing 1006-1090 NO2--N mg/L). The
181
drainage ratio at this stage was set at 30% in order to adjust initial NO2--N
182
concentrations at the beginning of each cycle to the same concentration (~300 mg/L).
183
Sodium acetate was provided as the electron donor for denitrification throughout the
184
operational period and the COD/N was maintained at 5.0.
185
Throughout operation, NO2--N, liquid phase N2O, gaseous phase N2O, and total
186
N2O were monitored to determine nitrogen removal efficiencies and N2O production
187
rates. The proportion of N2O in biogas was also determined to ensure the feasibility of
188
N2O recovery and bioreactor energy recovery was calculated from the co-combustion
189
of methane with produced nitrous oxide.
190
2.5 DNA extraction and sequence analysis
191
DNA was extracted from samples collected from the MBBR during cycles 1, 55,
192
and 85 with the E.Z.N.A. Soil DNA Kit (Omega Bio-Tek, Inc., Norcross, GA, USA)
193
according to the manufacturer’s instructions. A Nanodrop UV spectrophotometer
194
(Thermo Fisher Scientific, Delaware) was used to determine the DNA concentration
195
and purity of each sample.
196
Bacterial 16S rRNA gene fragments were amplified by PCR with the 338F/806R
197
primer set. Amplicons were sequenced on an Illumina Hiseq 2000 platform (Illumia,
198
San Diego, USA) by Majorbio Bio-Pharm Technology Co. Ltd. (Shanghai, China).
199
Sequences were placed into various operational taxonomic units with Pyrosequencing
200
Pipeline software (https://pyro.cme.msu.edu). Raw sequence files have been
201
submitted to the NCBI Sequence Read Archive database under accession NO.
202
PRJNA558633.
203
2.6 Analytical methods
204
NO2--N, NO3--N, NH4+-N, and COD were determined by standard methods
205
(APHA., 2005). Biomass was determined with a spectrophotometer (UV-1800,
206
Shimadzu, Japan) set at a wavelength of 600 nm. The pH was measured with a HACH
207
sensION+pH3®pH meter (HACH, USA). The N2O portion of biogas was analyzed
208
with a gas chromatograph (Agilent 7890A) equipped with an electron capture detector
209
(ECD). N2O dissolved in the liquid phase was calculated using Henrys law based on a
210
previously described method (Wang et al., 2018a). Statistical differences were
211
determined by one-way analysis of variance (ANOVA), with p＜0.05 considered to
212
be statistically significant.
213
3. Results and Discussion
214
3.1 Effects of FNA on nitrite removal and N2O conversion by the nosZ-deficient
215
mutant strain
216
A batch experiment was conducted to determine the threshold of FNA inhibition
217
and to identify optimal FNA concentrations for nitrite removal by the nosZ-deficient
218
P.
219
strain was grown in medium containing 0.206 mg/L initial FNA at pH 7.0, 0.042
220
mg/L initial FNA at pH 7.5, or 0.016 mg/L initial FNA at pH 8.0. Marked differences
221
in nitrite conversion rates were observed between the three conditions (Fig. 1a).
222
Cultures amended with an initial FNA concentration of 0.016 mg/L had the highest
aeruginosa strain during the partial denitrification process. The nosZ-deficient
223
nitrite conversion rate (6.77 mgN/(L ·h)), while nitrite reduction was minimal in the
224
other conditions. The bottles supplemented with 0.016 mg/L also formed the highest
225
concentrations of total nitrous oxide (0.398 ± 0.022 mM), while bottles grown under
226
the other conditions produced little N2O (~0.05 ± 0.022 mM) (Fig. 1b). Similar to the
227
nosZ-deficient P. aeruginosa strain, denitrification by Accumulibacter was also
228
completely inhibited by FNA concentrations of 0.2 mg/L (Zhou et al., 2010).
229
COD concentrations also decreased most rapidly in the 0.016 mg/L FNA bottles
230
during the first 18 hours of the experiment; 25.05 mg/(L·h) compared to 11.67
231
mg/(L·h) and 11.17 mg/(L · h) in the 0.042 mg/L FNA and the 0.206 mg/L FNA
232
bottles, respectively (Fig. 1c). These results are consistent with the fact that most of
233
the electron acceptor (NO2-) had been utilized by 18 h in the 0.016 mg/L FNA bottle
234
meaning that acetate could no longer serve as an electron donor for denitrification.
235
Therefore, the only way that the nosZ-deficient strain could support growth at this
236
point was via fermentation of compounds present in the yeast extract as this organism
237
cannot ferment acetate (Eschbach et al., 2004; Glasser et al., 2014; Vander Wauven et
238
al., 1984).
239
Although COD consumption in the 0.016 mg/L FNA bottle almost completely
240
leveled off by 18 h, CODs continued to steadily decline until 30 h for the 0.206 mg/L
241
FNA bottle and 36 h for the 0.042 mg/L FNA bottle (Fig. 1c). This is likely because
242
nitrite was still available as an electron acceptor at this point and although
243
denitrification
244
denitrification could support some growth. However, it appears that the majority of
245
COD consumption in the bottles with elevated FNA concentrations could be
246
attributed to fermentation and biomass conversion pathways.
enzymes
were
impaired
by
elevated
FNA
concentrations,
247
FNA concentrations in the 0.016 mg/L FNA bottle dropped down to only 0.0025
248
mg/L after 6 h and concentrations were negligible by 36 h (Fig. 1d). During the first
249
18 hours of the experiment, pH in the 0.042 mg/L bottle decreased from 7.58 to 7.11
250
which caused FNA concentrations to increase from 0.042 mg/L to 0.122 mg/L. This
251
drop in pH was most likely caused by an increase in acidic by-products formed from
252
the fermentation of yeast extract (CO2 and fatty acids) by Pseudomonas aeruginosa.
253
After 18 hours, however, FNA concentrations went back down to 0.042 mg/L and
254
this decline in FNA corresponded with an increase in nitrite consumption (Fig. 1a)
255
and nitrous oxide production (Fig. 1b). This suggests that cells started to participate in
256
denitrification once FNA concentrations had dropped. In the 0.206 mg/L FNA bottle,
257
pH was maintained at 6.95 ± 0.06 for the entire 30 h experiment (see Supplementary
258
Material). FNA concentrations in the 0.206 mg/L bottle also remained above 0.148
259
mg/L and these concentrations inhibited growth of P. aeruginosa as evidenced by the
260
finding that little denitrification or fementation occurred in this bottle.
261
Consistent with the finding that the nosZ-deficient mutant was only able to grow
262
efficiently in the 0.016 mg/L FNA bottle, biomass measurements actually declined in
263
the 0.206 mg/L FNA bottle, only slightly increased in the 0.042 mg/L FNA bottle, and
264
significantly increased in the 0.016 mg/L FNA bottle over the course of the
265
experiment (see Supplementary Material).
266
3.2 Effect of FNA on expression of denitrification pathway genes
267
In order to explore the influence of FNA on transcription patterns of
268
denitrification genes, qRT-PCR was done with primers targeting various genes from
269
the dissimilatory nitrate reduction pathway (Fig. 2). The results clearly demonstrated
270
that transcription of denitrification genes was severely inhibited by elevated FNA
271
concentrations. For example, 7.8 and 146.6 times more mRNA transcripts from the
272
gene that codes for the alpha subunit of nitrate reductase/nitrite oxidoreductase (narG)
273
were being expressed in cells grown in the 0.016 mg/L FNA bottle than the 0.042
274
mg/L FNA and 0.206 mg/L FNA bottles, respectively (Fig. 2). The protein encoded
275
by narG can catalyze both the reduction of nitrate to nitrite and the oxidation of nitrite
276
to nitrate (Gonzalez et al., 2006), and the narGHI operon is regulated by sensor
277
proteins that are activated by either nitrate or nitrite (Noriega et al., 2010). Studies
278
have also shown that elevated FNA concentrations can inhibit enzymes from the
279
denitrification pathway (Fudala-Ksiazek et al., 2014). This explains why narG was
280
being expressed in nitrite-grown cells and why the number of narG mRNA transcripts
281
was higher in the FNA 0.016 mg/L bottle than the FNA 0.042 mg/L or 0.206 mg/L
282
bottles.
283
Genes coding for the NO-forming nitrite reductase (nirS) and the beta subunit of
284
nitric oxide reductase (norB) were also more significantly expressed in the 0.016
285
mg/L FNA bottle. Cells from the 0.016 mg/L FNA bottle expressed 8.6 and 38.4
286
times more nirS mRNA transcripts and 6.5 and 35.5 times more norB transcripts than
287
the 0.042 mg/L FNA and the 0.206 mg/L FNA bottles.
288
No transcripts for genes coding for periplasmic nitrate reductase (napA) or the
289
large subunit from assimilatory nitrite reductase (nirB) were detected in the three
290
bottles. These results are consistent with previous studies that have shown that napA
291
is expressed under aerobic and microaerophilic conditions and that nirB transcription
292
is only induced under high nitrate concentrations (Lin et al., 2018). In addition, the
293
nosZ gene was not expressed, confirming the purity of the nosZ-deficient mutant
294
strain used in these studies.
295
3.3 Performance of the MBBR treating effluent from partial nitrification of
296
incineration leachate and supplemented with the nosZ-deficient P. aeruginosa
297
strain.
298
Once optimal FNA concentrations for growth of the nosZ-deficient strain were
299
identified in batch experiments, tests were done to determine whether this strain could
300
survive in long-term sequencing batch bioreactors treating effluent from partial
301
nitrification of incineration leachate, and whether N2O generated by this strain could
302
be recovered efficiently during long-term operation.
303
The denitrification MBBR received 200 mL of suspended mutant strain cells
304
(OD600 of 0.82) cultured in LB media. The reactor was operated for 85 cycles with
305
each cycle duration set at 24 h per cycle (Fig. 3). During the first stage of the
306
experiment (Stage I: cycle 1-15), NO2--N concentrations were maintained at ~100
307
mg/L and COD concentrations ranged from 500 to 550 mg/L to ensure biofilm
308
formation. During this stage, the nosZ-deficient strain gradually acclimated to
309
denitrifying conditions, and NO2--N removal efficiencies gradually increased from
310
4.85% to 55.58% for each cycle. By the end of Stage I (cycle 15), NO2--N removal
311
efficiencies were greater than 60% (Fig. 3).
312
During the second stage of the experiment (Stage II: cycles 16-55), NO2--N
313
concentrations ranged from 100 to 300 mg/L and COD concentrations ranged from
314
500-1500 mg/L to keep the COD/N at ~5.0. NO2--N removal efficiencies increased
315
further (85% to 99%), and became relatively stable after cycle 21. Approximately 2.7
316
mmol/d, 4.6 mmol/d, 6.9 mmol/d, and 10.3 mmol/d of N2O were generated by the
317
experimental reactor when it was provided with NO2--N concentrations of 100 mg/L,
318
150 mg/L, 200 mg/L, and 300 mg/L, respectively. The majority (＞70%, Fig. 3) of
319
N2O being generated in the experimental reactor during Stage II was dissolved in the
320
aqueous phase. N2O gas only accounted for 1.24-9.22% of the biogas being produced
321
during this stage (Fig. 4). N2O conversion efficiencies ranged from 66.25% to 99.77%
322
in stage II, indicating that the reactor was operating stably. A thick biofilm had also
323
formed on the bio-filter by the end of this stage (see Supplementary Material).
324
In the third stage of operation, raw partial nitrification-treated leachate with
325
~1000 mg/L NO2--N and ~5000 mg/L COD was fed into the MBBR, and the drainage
326
ratio was changed from 100% to 30% (Stage III: cycles 56-85). NO2--N efficiencies
327
were maintained at ~99% throughout Stage III, and 10.6 mmol/d of N2O was
328
generated when the reactor was provided with NO2--N concentrations of 300 mg/L.
329
The majority (>90%, Fig. 3) of N2O being produced during Stage III was found in the
330
gas phase, and N2O gas accounted for 71.52-83.57% of the biogas (Fig. 4). N2O
331
conversion efficiencies continued to be stable and ranged from 95.44 to 99.64%
332
during this final stage, significantly higher than previous studies that achieved
333
efficiencies of 80% (Table 2).
334
The proportion of N2O in biogas generated by the MBBR reactor increased
335
significantly during Stage III (after cycle 56) (Fig. 4). Prior to Stage III, N2O only
336
accounted for 15% of the biogas. However, after cycle 56, when the drainage ratio
337
was decreased from 100% to 30%, the proportion of N2O in biogas increased to 80%.
338
This can be explained by the fact that the MBBR reactor discharged 70% less leachate
339
after each cycle and that the influent was changed from diluted (<300 mg/L NO2--N)
340
to non-diluted (~1000 mg/L NO2--N) leachate (although after thoroughly stirring,
341
NO2--N concentrations at the beginning of each cycle remained at ~300 mg/L (1000 *
342
30% / 100%)). Under these conditions, the non-discharged leachate was already
343
saturated with dissolved N2O, and this left little liquid volume for any newly
344
generated N2O, concentrating N2O in the gas phase.
345
The equation of co-combustion with methane and nitrous oxide (Eq. 2) was used
346
to quantify energy recovered in N2O from the MBBR reactor. A significant amount of
347
energy (average 9206 kJ/m3) was recovered during Stage III (cycles 56 to 85), but
348
little energy (average 475 kJ/m3) was recovered during cycles 1 to 55.
349 350
°
CH4 + 4N2O →CO2+2H2O + 4N2, ∆HC = - 1219 kJ/mol
(2)
3.4 Bacterial community analysis
351
In order to ensure that microorganisms already present in the partial
352
nitrification-treated leachate used as influent for the MBBR system did not
353
competitively exclude the nosZ-deficient Pseudomonas strain, bacterial community
354
structure in the MBBR was assessed at different periods of operation (cycles 1, 55,
355
and 85) (Fig. 6). The nosZ-deficient Pseudomonas strain dominated the community at
356
the beginning of the experiment and accounted for 99.7% of the bacterial sequences
357
detected after cycle 1. As the experiment progressed, the MBBR bacterial community
358
became more diverse and Pseudomonas only accounted for 38.35% and 24.77% of
359
the bacterial sequences after cycles 55 and 85.
360
Tissierella became the most significant organism and accounted for 55.86%-58.63%
361
of the population after cycle 55. This genus has been isolated from human sewage
362
treatment plants (Harms et al., 1998) and has been a dominant member of bacterial
363
communities associated with several anaerobic digesters treating complex waste
364
(Jaenicke et al., 2011; Liu et al., 2009; Nolla-Ardèvol et al., 2015; Snell-Castro et al.,
365
2005; Ziganshina et al., 2014). Tissierella can metabolize nitrogen-containing
366
compounds but is not capable of denitrification (Harms et al., 1998). Its metabolism is
367
also stimulated by the presence of formate (Harms et al., 1998) which is a
368
fermentation by-product formed by P. aeruginosa (Eschbach et al., 2004). Therefore,
369
it is possible that production of formate by Pseudomonas helped stimulate growth of
370
Tissierella, however further investigation into this possibility is required.
371
Other bacterial genera that started to appear later in the experiment included
372
Muricomes, Alkaliphilus, Jeotaglibaca, and Clostridium (Fig. 6). Clostridium is
373
known to be capable of dissimilatory nitrate reduction to ammonia (DNRA) (Caskey
374
& Tiedje, 1980; Keith et al., 1982) and genes for ammonia forming nitrite reductase
375
(nrfAH) were found in Tissierella and Alkaliphilus genomes. The role that these other
376
microorganisms were playing in the reactor is unclear. However, it is apparent that
377
conditions in the reactor did not promote growth of traditional complete or incomplete
378
denitrifiers that could generate N2O. Therefore, it appears that under the conditions
379
used in this study, reactors need to be supplemented with N2O producing bacteria
380
such as the nosZ-deficient strain of P. aeruginosa to generate high concentrations of
381
N2O throughout the experiment. These findings are significant and show that
382
generation of N2O by the nosZ-deficient strain of P. aeruginosa and potentially other
383
denitrifying strains have real applications within the wastewater treatment field.
384 385
4. Conclusions
386
Batch experiments revealed that FNA concentrations above 0.016 mg/L severely
387
inhibited denitrification by the nosZ-deficient strain of P. aeruginosa. Addition of this
388
strain to an MBBR dramatically improved performance and resulted in nitrite removal
389
and N2O conversion efficiencies >95%. Adjustment of the drainage ratio to 30%
390
increased the proportion of N2O in biogas to levels above 80%. Bacterial community
391
analyses revealed that the nosZ-deficient Pseudomonas was primarily responsible for
392
N2O formation in the bioreactor throughout the experiment. These results demonstrate
393
that high concentrations of N2O can be recovered from incineration leachate in
394
MBBR supplemented with the nosZ-deficient Pseudomonas strain.
395 396
Note
397
The authors declare no competing interest.
398 399
Appendix A. Supplementary Data
400
E-supplementary data for this work can be found in e-version of this paper online.
401 402
Acknowledgments
403
This research was financially supported by the National Natural Science
404
Foundation
405
Natural Science Foundation (8184081) and the Major Science and Technology
406
Program for Water Pollution Control and Treatment (2017ZX07108-002).
of
China
(51678051,
51708031),
the Beijing
Municipal
407 408 409 410
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555
556 557
Figure captions Fig. 1. Growth of the nosZ-deficient strain of P. aeruginosa PAO1 under
558
denitrifying conditions with acetate (19.5 mM) provided as an electron donor. Cells
559
were exposed to three different initial free nitrous acid (FNA) concentrations. Nitrite
560
concentrations (a); total nitrous oxide (from liquid and gaseous phases) (b); chemical
561
oxygen demand (COD) (c), and free nitrous acid (FNA) (d) were monitored over the
562
course of 36 hours. Error bars were calculated from triplicate samples.
563
Fig. 2. The number of narG, napA, nirS, norB, nirB and nosZ mRNA transcripts
564
normalized against the number of proC mRNA transcripts detected in the
565
nosZ-deficient P. aeruginosa cells exposed to different FNA concentrations. Error
566
bars were calculated from triplicate samples.
567 568 569 570
Fig. 3. Performance of the MBBR supplemented with the nosZ-deficient P. aeruginosa strain. Fig. 4. Proportion of N2O in biogas generated by the MBBR supplemented with the nosZ-deficient P. aeruginosa strain.
571
Fig. 5. Energy recovered (calculated by nitrous oxide combustion with methane)
572
from nitrous oxide generated by the nosZ-deficient P. aeruginosa-supplemented
573
MBBR system.
574 575
Fig. 6. Proportion of various genera detected in the MBBR system supplemented with nosZ-deficient P. aeruginosa after cycles 1, 55, and 85.
576 577
Table captions
578
Table 1 Characteristics of experimental wastewater used in this study
579
Table 2 Nitrogen removal efficiencies, N2O production rates, and N2O
580
conversion efficiencies obtained with the method used in this study compared to other
581
approaches used for N2O recovery.
582
Table 1. Characteristics of experimental wastewater used in this study COD
BOD5
NH4+-N
NO3--N
NO2--N
(mg/L)
(mg/L)
(mg/L)
(mg/L)
(mg/L)
1980-3660
850-2380
1140-1420
1124-1545
125-342
202-310
Item
pH
AnaerobicallyN/A
N/A
8.05-8.55
1006-1090
7.12-7.53
treated leachate Partial nitrification treated leachate 583
18.28-24.65
584
Table 2. Nitrogen removal efficiencies, N2O production rates, and N2O conversion
585
efficiencies obtained with the method used in this study compared to other approaches
586
used for N2O recovery. Nitrogen Study
N2O
N2O
removal
production rate
conversion
efficiency
(mgN/(gVSS·h))
efficiency
nosZ-deficient
This study
＞97%
119.7±5.7
＞95%
Pseudomonas
(Lin et al., 2018)
＞97%
106.1±4.9
70-80%
(Scherson et al., 2013)
＞98%
5.6
60-65%
(Scherson et al., 2014)
＞98%
25.2
75-80%
72%
2.1±0.4
65-75%
(Gao et al., 2017)
＞98%
5.1±1.6
70-80%
(Weißbach et al., 2018b)
＞95%
1.05 ± 0.20
53-63%
80%
2.38±0.26
50-65%
＞98%
N/A
71-73%
CANDO
NO inhibit Light-driven
(Myung et al., 2015)
(Yu et al., 2019) (Chen et al., 2019)
587 588 589
Highlights
590
Optimal FNA concentrations for the nosZ-deficient Pseudomonas were <0.02 mg/L
591
FNA inhibited transcription of genes from the denitrification pathway
592
N2O conversion efficiencies of MBBR supplemented with nosZ-deficient
593 594
strain were >95%
595 596 597 598 599
Drainage ratio adjustments increased the proportion of N2O in biogas to 80%
The nosZ-deficient strain remained the dominant denitrifier throughout MBBR operation