Methane and nitrous oxide emission from different treatment units of municipal wastewater treatment plants in Southwest Germany

We measured the atmospheric emission rates of methane (CH4) and nitrous oxide (N2O) in two wastewater treatment plants in Southwest Germany, which apply different treatment technologies. Dissolved gas concentrations and fluxes were measured during all processing steps as well as in the discharge receiving streams. N2O isotopocule analysis revealed that NH2OH oxidation during nitrification contributed 86–96% of the N2O production in the nitrification tank, whereas microbial denitrification was the main production pathway in the denitrification tank in a conventional activated sludge (CAS) system. During wastewater treatment using a modified Ludzack-Ettinger system (MLE) with energy recovery, N2O was predominantly produced by the NO2- reduction by nitrifier-denitrification process. For both systems, N2O emissions were low, with emission factors of 0.008% and 0.001% for the MLE and the CAS system, respectively. In the effluent-receiving streams, bacterial denitrification and nitrification contributed nearly equally to N2O production. The CH4 emission from the MLE system was estimated as 118.1 g-C d-1, which corresponds to an emission factor of 0.004%, and was three times lower than the emission from the CAS system with 0.01%.


Introduction
Sewage treatment is an important source of anthropogenic greenhouse gases with significant amounts of methane (CH 4 ), nitrous oxide (N 2 O) and carbon dioxide (CO 2 ) being released during wastewater treatment [1]. CO 2 is produced both indirectly as a result of fossil fuel combustion for energy generation that is required for the operation of waste water treatment plants (WWTPs), and it is produced during the degradation of organic matter during the treatment process. While the latter emissions are considered as short-cycle CO 2 that does not contribute to increasing atmospheric CO 2 concentrations [1], the emissions of CH 4  Over a 100-year time span, CH 4 has 34-fold a global warming potential compared to CO 2 [2]. It is mainly generated in the sewer system [3], in the anaerobic treatment zone and during sludge treatment [4]. N 2 O is produced during the biological nitrogen removal processes via nitrification and denitrification. It is the most powerful gas that destroys the ozone layer [5,6] and it has global warming potential which is 265 times greater than that of CO 2 [2]. CH 4 and N 2 O contribute for 16% and 6.2% to the global anthropogenic greenhouse gas emission, respectively, with the remainder consisting of CO 2 (76%) and fluorinated gases (2%) [2]. The IPCC classifies the global anthropogenic emission in seven sectors, one of them being waste and wastewater. The wastewater treatment sector is assumed to be responsible for 3.2% (CH 4 ) and 4-5% (N 2 O) of the total anthropogenic emissions [5,7]. However, these assessments are associated with a high degree of uncertainty because microbial activities are sensitive to numerous variables of the actual treatment processes, such as dissolved oxygen (DO) concentration, pH, temperature, and substrate availability. Very few direct measurements of emission rates exist, which showed relatively large variations [1,4,8,9,10,11,12]. Daelman et al. [2012] found that about 1.13% of the incoming chemical oxygen demand (COD) in the wastewater is emitted as CH 4 from a WWTP located in the Netherlands, while 1.7% of the nitrogen loading rate (NLR) from nitrifying reactors, 15% of the NLR in full-scale reactors, and 95% of the NLR in lab-scale bioreactors were converted to N 2 O.
In addition to in-plant emissions during the treatment process, WWTP export CH 4 and N 2 O in dissolved form with effluent discharge. In a regional-scale study, Alshboul et al. [2016] observed enhanced CH 4 concentrations and fluxes downstream of WWTP in small streams in Central Europe [13]. The annual mean CH 4 concentration in the treatment plant effluents was correlated to the corresponding mean COD of the treated wastewater. Enhanced concentrations and fluxes of N 2 O from rivers near urban areas and downstream of WWTP have also been attributed to WWTP effluents [14,15,16].
To assess the anthropogenic influence on emission rates and to obtain reliable climate change projections of anthropogenic N 2 O and CH 4 emissions, improved process-based understanding of the relevance of varying environmental conditions are required. Prior studies revealed that N 2 O production rates are predominantly affected by varying temperature and DO concentrations [15;17]. The relative importance of the different microbial production pathways to N 2 O was based on correlation analysis with DO, dissolved organic carbon (DOC), nitrate (NO 3 -), and dissolved nitrous oxide (DN 2 O). Stable isotopocule ratios of N 2 O ( 15 N and 18 O) have been explored as a noninvasive method to assess N 2 O production and consumption mechanisms in WWTPs, as well as in rivers and streams which receive effluents [18]. The interpretation of bulk δ 15 N and δ 18 O of N 2 O, however, is challenging because of the dependency of δ 15 N and δ 18 O on the isotopic composition of the precursors (ammonium (NH 4 + ) and NO 3 -) and the uncertainty of isotopic fractionation during various processes. It has been suggested that the intramolecular distribution of δ 15 N in the asymmetric N 2 O molecule could serve as a tool to discern various N 2 O production and consumption processes and can help to constrain the global N 2 O budget [19; 20]. The sitepreference (SP) is defined as the difference between the N isotopic ratios of the central and the terminal N atom (δ 15 N α and δ 15 N β values). An advantage of using SP is that it is assumed to be independent on δ 15 N of the precursors [19,20]. The SP of N 2 O differs for hydroxylamine (NH 2 OH) oxidation during nitrification and nitrite (NO 2 -) reduction during denitrification. This selective approach has been applied in several studies including full scale wastewater treatment and in-situ measurements of different N 2 O production pathways during wastewater treatment processes [21,22,23], however, to a very limited extent to inland waters [24].
This study was conducted (1) to quantify CH 4 and N 2 O emissions from two different WWTPs located in southern Germany, (2) to analyze the N 2 O production pathways based on isotopocule ratios of N 2 O in these systems to parameterize isotopocule ratios of N 2 O emitted from WWTP, and (3) to determine the influence of N 2 O and CH 4 discharge with WWTP effluent on concentrations and fluxes in the receiving streams.

Study site
Two municipal WWTPs located in Southwest Germany were studied in this investigation: Bellheim and Ruelzheim. The Bellheim WWTP, which receives wastewater from a population equivalent (PE) of 14.3 thousand at an average flow rate of approximately 5×10 3 m 3 day -1 , applies a conventional activated sludge (CAS) treatment system. The system comprises two clarifiers and a series of biological reaction tanks. Heavy solids are removed from wastewater in the primary clarifier. The water then undergoes biological treatment to decompose organic matter by activated sludge under aerobic and anaerobic conditions. The respective switching time between denitrification and nitrification process were 45 and 80 min, respectively. Subsequently, the microbe-rich liquid flows into the secondary clarifier where activated sludge is separated from treated wastewater by gravity. Parts of the settled sludge is continuously recycled back to the aeration tanks to maintain the microbial community and the treated effluent is finally discharged into a neighboring stream ( Fig 1A).
The treatment capacity of Ruelzheim WWTP is three times greater than Bellheim WWTP, and it applies a modified Ludzack-Ettinger (MLE) system with energy recovery. It treats wastewater from 41.5 thousand PE with a flow rate of approximately 14.7×10 3 m 3 day -1 . The surplus sludge from the plant is treated in an anaerobic digester for biogas production. The gas is used for digester heating and electricity generation and provides about 60.8% of the total in-plant energy requirements. Wastewater treated by the primary clarifier enters an anaerobic tank for denitrification and then an aerobic tank for nitrification by activated sludge. From the aeration tank, the mixed liquor flows to the secondary clarifiers to separate treated wastewater from the sludge. Parts of the sludge are recycled back into the biological tank, while the remaining sludge is fed to the anaerobic digester for energy generation. The last step of this treatment system is identical to the CAS system ( Fig 1B).

Sample collection
Water samples were collected as duplicates at eleven sites in each treatment system ( Dissolved CH 4 (DCH 4 ) and CO 2 (DCO 2 ) concentrations in water samples were measured on-site using the headspace method. The headspace was created in a borosilicate glass bottle and the gas partial pressure was measured in a closed gas loop with an ultraportable greenhouse gas analyzer (UGGA, Los Gatos Research Inc.). More detailed information about the concentration measurements can be found in Alshboul et al. [2016]. DN 2 O concentrations were measured using a gas chromatograph equipped with an electron capture detector (ECD). The concentrations of dissolved NH 4 + , NO 3 and NO 2 were measured using a portable  Water temperature, DO, pH, specific conductivity and redox potential were measured on site using a pH-temperature electrode with gel electrolyte (SenTix 21; 0-14 pH; 0-80˚C; ±0.2˚C), a DO sensor (FDO 925; 0-20 mg L -1 ± 0.5%; 0-50 ± 0.2˚C), a conductivity cell (Tetra-Con 925, 10-2000 ± 0.5% mS cm -1 ; 0-100 ± 0.2˚C) and an oxidation/reduction potential electrode (SenTex ORP 900; ± 1200.0 ± 0.2 mV; platinum) connected to a portable three channel multi meter (3430 IDS; WTW GmbH).

Flux measurements
Fluxes of N 2 O and CH 4 at the air-water interface in the treatment ponds and the effluent receiving streams were measured using floating chambers (surface area: 0.078 m 2 , chamber volume: 7.66 l), covered with aluminum foil to reduce the internal heating. The flux measurements were performed using triplicate floating chamber deployments with 35 min duration each. The slope describing the rate of change of the measured headspace concentration in the chamber over time was used to calculate the CH 4 and N 2 O fluxes (fN 2 O and fCH 4 ) by applying the following equation [25]: where S is the slope (ppm s -1 ), V is the chamber volume (m 3 ), A is the chamber area (m 2 ), a 1 (86400 s d -1 ) is a conversion factor from seconds to days, and a 2 (0.6788 mg m -3 and 1.8625 mg m -3 for CH 4 and N 2 O, respectively) is a conversion factor from ppm to mg m -3 at in-situ measured temperature (T in K) and standard pressure (p in Pa): where M is the molar mass of CH 4 and N 2 O (g mol -1 ). Fluxes of N 2 O and CH 4 from each tank in both plants were used to estimate emission factors (EF) which expressed by kg N and C emitted as N 2 O and CH 4 per kg DIN and COD in influent wastewater. These factors provide information about the influence of the operational procedures and process design on the mass balance approach over each treatment stage.

Analysis of isotopocule ratios
For analysis of isotopocule ratios of DN 2 O, samples were prepared by injecting a headspace of 120 mL of ultrapure helium (He) and subsequent equilibrating of liquid and gas phases at constant temperature (20˚C). The headspace was transferred into 115 mL glass bottles, which had been flushed with N 2 gas. The analyses were performed using a Delta XP isotope ratio mass spectrophotometer (IRMS, MAT 251, Thermo-Finnigan, Bremen, Germany), which allows simultaneous detection of m/z 30, 31, 44, 45 and 46. The notation of the isotopocule ratios is the following: where 15 R α and 15 R β represent the 15 N/ 14 N ratios of α and β N atoms, respectively. 15 R bulk and 18 R denote average isotope ratios for 15 N/ 14 N and 18 O/ 16 O, respectively. Subscripts "sample" and "std", respectively, signify isotope ratios for the sample and the standard, atmospheric N 2 for N and Vienna Standard Mean Ocean Water (VSMOW) for O.
The 15 N site preference (hereinafter, SP) was also defined as an illustrative parameter of the intramolecular distribution of 15 N. Site-specific N isotope analysis in NO was conducted using ion detectors modified for mass analysis of the N 2 O fragment ions (NO + ), which contained N atoms in the α position of the N 2 O molecules, whereas bulk (average) N and O isotope ratios were determined from molecular ions (N 2 O + ) [19]: Therein, SP NO2-reduction and SP NH2OH oxidation , respectively, denote the SP values when N 2 O is produced only by NO 2 reduction and when N 2 O is produced only by NH 2 OH oxidation.
The δ 15 N of NH 4 + was measured using the diffusion method [32] where about 10 μmol of NH 4 + in the sample was concentrated onto a GF/D glass fiber filter containing H 2 SO 4 and analyzed using an EA1110 elemental analyzer (Thermo Fisher Scientific K.K.) coupled with the IRMS. DN 2 O concentrations in the nitrification basin in Bellheim were almost constant and ranged between 13.7 and 20.1 nmol L -1 at the points from B-N1 to B-N3, where the DO concentrations were between 0.5 (31.3 μmol L -1 ) and 3.8 mg L -1 (237.5 μmol L -1 ). N 2 O production is closely linked to oxygen concentration, which plays a critical role in influencing N 2 O emission. This is proved by the obtained results in the nitrification basins in Ruelzheim (R-N1-R-N4), where N 2 O concentrations were relatively high (15.3-92.6 nmol L -1 ), while DO ranged from 0.1 (8.1 μmol L -1 ) to 1.4 mg L -1 (87.5 μmol L -1 ). Our results agree with those of Tumendelger et al., [2014] where lower DO was found to increase N 2 O production in the aerobic treatment tank due to local oxygen limitation. In the denitrification basins in Bellheim, N 2 O concentrations increased gradually up to 103.5 nmol L -1 in B-D3 at a DO concentration of 0.03 mg L -1 (1.9 μmol L -1 ) while DN 2 O increased to 69.9 nmol L -1 at R-D2 in Ruelzheim at a comparable oxygen concentration. However, oxygen can inhibit both synthesis and activity of denitrifying enzymes. N 2 O reductase is more sensitive to oxygen than the other enzymes, leading to N 2 O emission during denitrification when oxygen is present even in low amounts [Kampschreur et al., 2009]. Aside from DO, N 2 O production can depend on the carbon to nitrogen (C/N) ratio as electron donor is considered to be an important parameter for N 2 O accumulation, especially at low C/N ratios [23]. The highest DN 2 O concentrations of 885-970 nmol L -1 were observed in the secondary settling tank (B-SS) and in the effluent water (B-E) at Bellheim, which exhibited a three times greater increase than the DN 2 O concentration at same sampling locations in Ruelzheim. DN 2 O concentrations at all stations in both plants were higher than the atmospheric equilibrium concentration of about 12 nmol L -1 at 12˚C and 13 nmol L -1 at 8˚C [35]. The water in the biological reaction basins (nitrification and denitrification) was supersaturated with N 2 O by about 800% (Figs 2B and 3B), indicating that the wastewater treatment system is a source of N 2 O to the atmosphere.

Results and discussion
In the receiving streams, the effluent discharge had a significant effect on most of the measured physico-chemical parameters and also on the concentration of DIN, DO, DN 2 O (Table 1). In particular, the upstream water had lower DN 2 O concentration (48.6 nmol L -1 at B-US), than the water downstream of the effluent discharge location. However, upstream water was saturated with DN 2 O in comparison to atmospheric equilibrium. The extreme changes in DN 2 O concentration of up to a factor of three suggest that effluent-DN 2 O significantly affected downstream concentrations and that additional amounts of N 2 O produced  Fig 2D). The values observed at B-N1-B-N3 were comparable to the value of -13.5‰ measured in the oxic tank of a Japanese WWTP by Toyoda et at. [2011]. The δ18O showed a wide range of variation between +41.3‰ and +82.8‰ throughout the treatment processes, however, the general trend was similar to that of the δ15N. The observed decrease in δ15N bulk and δ 18 O could be interpreted as an isotope effect in microbial N 2 O production during the treatment process. We found high SP values (+28.5‰-+31.6‰) at B-N1-B-N3 whereas it was slightly negative (-1.8‰--1.9‰) in B-D1-B-D3 (Fig 2D) which coordination of NO molecules to Fe center is followed by invasion of another NO [36]. Therefore, SP would be negative because of (1) 15 NO bounds to Fe more preferentially than 14 NO due to intermolecular isotope fractionation and (2) (Fig 4). The range of δ 15 N bulk of N 2 O associated with each experiment was estimated from δ 15 N of NH 4 + (in case of nitrification) and the isotopic enrichment factor (ε( 15 N) NH4 + -!N2O = -60 to -48‰, [39]) was estimated from the   (Fig 2A). The contribution of the NH 2 OH oxidation pathway to N 2 O production in this basin was about 86.3-96.1% (Eq 6). In contrast, the data at B-D1-B-D3 was falling in the range of NO 2 reduction, suggesting that the dominant source of the produced N 2 O in the denitrification basin was NO 2 reduction by bacterial denitrification (Fig 4). Moreover, the lower SP values observed at B-I and B-SS are within the range of the NO 2 reduction source, indicating that the NO 2 reduction pathway was dominant (>90%).
In Ruelzheim, N 2 O sampled in the denitrification basin had a greater δ 15 N bulk (9.9‰) than N 2 O produced in the nitrification basin where SP was the lowest and ranged between -12.5‰ and -10.7‰ (Fig 3D). This negative SP values can be caused by the inorganic N 2 O production via NO 2 reduction (nitrifier denitrification) by Fe 2+ where SP ranged from -13.3 to +22.6‰ in the measurements of Samarkin et al.
[2010] [40]. Relatively few measurements of isotopocule ratios of DN 2 O in fresh water including river and streams have been reported to date [14,24]. In our samples, stream-emitted N 2 O has lower δ 15 N bulk values (+0.35‰-+1.95‰ at Bellheim and -6.1‰--2.8‰ at Ruelzheim) than tropospheric N 2 O (δ 15 N: +6.72%), indicating that biological N 2 O production is from additional "light" N 2 O in sewage plants. In contrast, the δ 18 O was relatively high about 70‰ in both streams. High δ 18 O values (> 30‰) in rivers or streams are likely produced from denitrification and/or N 2 O consumption which is expected to dominate riverine N 2 O production [15]. In case of upstream (B-U), this may also receive significant N 2 O inputs from ground water that can have high δ 18 O values (see Table 1). For the first time, we report measurements of the SP-N 2 O in small streams nearby WWTPs. The SPs observed at all in-stream sampling sites (+24.3 -+28.3‰) near to Bellheim were comparable to that of the N2O produced via NH 2 OH oxidation by AOB. This production pathway was also suggested by the NO 3 concentration at the sampling sites (Table 1). At Ruelzheim, in contrast, NO 2 reduction pathway dominantly contributed to the production of the N 2 O sampled in the streams, with SP values ranging between +12.9 and +13.2‰. Moreover, N 2 O reduction at sampling these sites might be the cause of increased SP (Table 1). Atmospheric fluxes of N 2 O. The amounts of N 2 O emitted to the atmosphere from both WWTPs are presented in Table 2. The total flux from Ruelzheim, which has separate basins for nitrification and denitrification, was 12.1 g-N d -1 that converts into an emission factor of 0.008% of N removed from influent DIN and released as N 2 O. Approximately 84% of the total N 2 O emission (10.2 g-N d -1 ) was from the nitrification basin where relatively low DO concentration existed. This high emission could be explained by the oxygen concentration, which is required for the oxidation processes. Because AOB have a higher oxygen affinity than NOB, low oxygen concentration resulted in NO 2 accumulation. For instance, the combination of low oxygen with elevated NO 2 accumulation can induce N 2 O production by the nitrifier-denitrification pathway [41]. The end product of this pathway is N 2 O since the genes encoding N 2 O reduction to N 2 are not yet found. The total flux of N 2 O from Bellheim (3.67 g-N d -1 , emission factor of 0.001%) was three times lower than that of Ruelzheim. About 76.8% of total N 2 O flux was emitted from the nitrification basin and this high emission rate could be caused by stripping and does not reflect the local production rate of N 2 O. Assuming that if emission  [43], corresponding to approximately 0.035% of the nitrogen load of a WWTP based on first measurement by Czepiel et al. [1995]. Our estimated data at both plants were lower than the IPCC base, and the emission factor observed at Bellheim was similar to Suemer et al. [1995] who obtained a value of 0.001% for activated sludge plant (Table 3) [44]. We found that the emission factors among different WWTPs are highly variable, which could be caused by (1) different treatment capacity and/or (2) different operational parameters including DO concentration, switching time between nitrification and denitrification and so on. Therefore, more research is required to reveal the dependence of the emission factor on operational parameters. It can be assumed that all N 2 O dissolved in the effluent water will be released to the atmosphere from the stream. In the stream nearby Ruelzheim, an unexpected negative flux (-0.83 mg N m -2 d -1 ), i.e. N 2 O uptake by the stream, was measured at the upstream sampling site (R-US), which can probably be (I) an artifact resulting from relatively short chamber deployment duration or (II), there was some residual organic carbon that may contributed to denitrification or N 2 O consumption that has not been discussed in this study [45]. However, a positive flux of 0.02 mg N m -2 d -1 was observed at mixing point (R-DS-M). The most likely explanation for the flux observed at this point relates to higher contribution of effluent-N 2 O from the plant (Table 3).

CH 4 emission
Characteristic of DCH 4 . All water samples were supersaturated with CH 4 in respect to atmospheric equilibrium and dissolved gas concentrations varied widely among sampling sites  (Figs 2 and 3). DCH 4 ranged between 0.14 and 12 μmol L -1 (average = 3.2±4 μmol L -1 ) and 0.01 and 9.9 μmol L -1 (average = 2.2±3 μmol L -1 ) at Bellheim and Ruelzheim WWTP, respectively. We did not observe a significant correlation between DO and DCH 4 at the sampling sites at Bellheim WWTP, suggesting that large portions DCH 4 were not produced at the sampling sites, but delivered from the preceding treatment steps. In contrast, we observed a correlation between DO and CH 4 for treatment processes at Ruelzheim indicating that CH 4 existence at these sites may be linked to local production. DCH 4 , and DCO 2 did not differ significantly between the nitrification and denitrification basins at both WWTPs (ANOVA, p < 0.05). High concentrations of dissolved CH 4 were observed in the inflows of both WWTPs (Fig 2 and Fig 3), indicating that considerable amount of CH 4 is delivered to the plants prior the treatment. High concentrations declined rapidly during the further treatment process, suggesting that a high percentage of the dissolved CH 4 is emitted during the mechanical motion of the wastewater at primary settling systems. Approximately 60% of the dissolved CH 4 in the inflowing wastewater was released at the primary settling at both WWTPs. However, there were no flux measurements at these points. Consistent and relatively low DCH 4 concentrations were observed during the nitrification processes (N1, N2 and N3) at both plants. High DCH 4 values were observed in the denitrification basins of Bellheim WWTP, which could be caused by the low oxygen concentration. DCH 4 increased in the secondary clarifier at both plants, where it was potentially produced under anaerobic conditions in the settled sludge. The discharge of effluent water caused a 13-fold increase of DCH 4 in the receiving streams at Bellheim, while the concentration decreased in Ruelzheim (Tabel 1). WWTPs have been shown to export additional amount of DCH 4 and this study confirmed the potential importance of WWTP effluents on the concentration of CH 4 in inland water [13]. More studies are required for investigation the factors controlling the magnitude and variability of DCH 4 exported with effluent water from WWTPs.
Atmospheric fluxes of CH 4 . Most sampling sites (16 out of 18 flux measurements), were net sources of atmospheric CH 4 ( Table 2). In accordance with the high concentrations of DCH 4 , high fluxes were observed at the denitrification basins and secondary settlers at both plants. At the secondary settlers, fluxes were additionally enhanced by the high gas exchange rate caused by the mechanical motion of skimmer arm and associated turbulence at the water surface. Based on our measurements, the nitrification basins are hot spots of CH 4 emission. In contrast to the denitrification process, the nitrification process is expected to have a high gas exchange velocity caused by air injection. The flux measurements are subject to uncertainty caused by spatial and temporal variability of flux rates, which were not resolved in the single point measurements. Additional uncertainty of the flux measurements at the nitrification/ denitrification basins are related to the chamber effect on the surface turbulence [46], as the chambers were anchored with the frame of the basins, while the chambers were freely floating at the other different sites.
Our measurements in Ruelzheim resulted in a total flux of 118.1 g-C d -1 as CH 4 or an emission factor of 0.004% of influent oxidized COD. Approximately 60.9% of the total CH 4 emission (71.9 g-C d -1 ) occurred at the secondary settler while 21.1% (25.0 g-C d -1 ) were emitted from the nitrification basin and 7% from the denitrification basin. At Bellheim, the total flux of CH 4 was 373.3 g-C d -1 , corresponding to an emission factor of 0.01%, which is about three times greater than the emission factor at Ruelzheim. Most CH 4 was emitted from the denitrification and secondary settling tanks.
Only few estimates of CH 4 emission factors exist. Czepiel [1,8,47]. The Dutch Ministry of Housing, Spatial Planning and the Environment in the Netherlands presents estimation results in a CH 4 emission factor of 0.85% for WWTPs with anaerobic sludge treatment while it was 0.70% for plants without anaerobic sludge treatment [1]. Our estimated data at both plants were lower than the emission factors reported in the above-mentioned studies. We expect that considerable amounts of DCH 4 are oxidized prior to discharge. However, high spatial and temporal measurement resolution is an urgent need for investigating production, emission and oxidation process of CH 4 during wastewater treatment process.

Conclusions
A flux of 12.1g-N d -1 was emitted as gaseous N 2 O from the MLE system, while it was 3.67g-N d -1 from CAS system. In contrast, the CAS system had a higher CH 4 emission factor than the MLE system. SP-N 2 O suggested that 86.3-96.1% of N 2 O was produced by NH 2 OH oxidation in the CAS system, whereas nitrifier-denitrification was the major N 2 O production pathway in the same basins in the MLE system. Effluent from the WWTPs increased dissolved gas concentrations and fluxes in the effluent-receiving streams. Microbial denitrification and NH 2 OH oxidation mainly produced N 2 O in the receiving streams nearby WWTPs as revealed by SP.