Assessment of Nitrous Oxide and Carbon Dioxide Emissions and the Carbon Footprint in an Aerobic Granular Sludge Reactor Treating Domestic Wastewater

Abstract

The assessment of nitrous oxide (N₂O) and carbon dioxide (CO₂) emissions from wastewater treatment processes using aerobic granular biomass is still scarce and most of them deal with synthetic wastewater in well-controlled conditions. Also, the greenhouse gases footprint from aerobic granular process has been poorly described despite a strong impact on climate change. This work quantified and analyzed the N₂O and CO₂ emissions, as well as the treatment efficiency of a pilot-scale sequencing batch reactor with aerobic granular sludge (AGS) fed with real domestic wastewater. The operational conditions applied to the reactor allowed the development of AGS (1.7 ± 0.40 gVSS/L), with good settleability (SVI₃₀ < 50 mL/g) and removal efficiencies of 84% for BOD₅ and 65% for NH₄⁺-N. The relative amount of nitrogen denitrified to N₂O was 22.9%, which corresponded to an emission factor ( $E F_{N_{2} O}$ ) of 5.67% ± 1.68%. The contribution of N₂O emitted through the gas phase was 97.1%, with a flow-based emission factor ( $F B E F_{N_{2} O}$ ) of 3.65 × 10⁻³ g N₂O-N/L, while the contribution of N₂O released through the liquid phase was 2.9%, with a flow-based release factor ( $F B R F_{N_{2} O}$ ) of 0.11 × 10⁻³ g N₂O-N/L. Although CO₂ flow-based emission factor ( $F B E F_{C O_{2}}$ ) was higher (0.297 g CO₂/L), in terms of carbon footprint, the global warming impact of N₂O was five times higher than the impact of CO₂ emissions ( $G W I_{N_{2} O}$ of 83.6% and $G W I_{C O_{2}}$ of 16.4%).

Introduction

Over the last 200 years, the concentration of atmospheric greenhouse gases (GHGs) has increased considerably due to the intensification of several anthropogenic activities. The Kyoto Protocol has set targets based on a group of six gases: carbon dioxide (CO₂), methane (CH₄), nitrous oxide (N₂O), hydrofluorocarbons, perfluorocarbons, and sulfur hexafluoride (SF₆). N₂O is the third most important GHGs in terms of its contribution to the increase in the planet's temperature. Its global warming potential is around 300 times higher compared with CO₂ (Song et al., 2015; Capodici et al., 2018; Weißbach et al., 2018).

Wastewater treatment plants (WWTPs) emit N₂O to the atmosphere, being responsible for 3.2–10% of the total of these emissions (Law et al., 2012). The amount of GHGs emitted by WWTPs is related to the treatment process (Yan et al., 2014; Singh and Kansal, 2018). Studies have reported that high amounts of N₂O are generated during biological nutrient removal (BNR) reactions, such as nitrification and denitrification (Han et al., 2019; Liu et al., 2020; Ross et al., 2020). The IPCC (Bartram et al., 2019) guidelines propose the value for the N₂O emission factor (EF) of 1.6% for centralized aerobic treatment systems. Due to this, the control and minimization of N₂O emissions have become essential points for the proper functioning of WWTPs.

Wastewater treatment by sequencing batch reactors (SBRs) is widespread and applied worldwide. In these systems, ammonia is transformed into gaseous nitrogen during the operational cycles of the reactor, through nitrification and denitrification reactions. N₂O generation may occur as an intermediate product or as a byproduct of the biochemical reactions (Duan et al., 2017; Guimarães et al., 2017; Liu et al., 2020). Conditions that stimulate N₂O formation include insufficient oxygen supply during the aeration phase, as well as microbial inhibition by excess oxygen or carbon scarcity during denitrification in the anoxic and settling phases (Massara et al., 2017; Vieira et al., 2019; Peng et al., 2020).

Currently, SBRs with aerobic granular sludge (SBR-AGS) have been proposed for the biological treatment of domestic and industrial effluents (Winkler et al., 2018). It has been proven that AGS is a feasible option to treat different types of wastewater, and capable of remove organic pollutants such as phenol, heavy metals and nutrient as well (Foley et al., 2010; Franca et al., 2018; Wang et al., 2018a). While GHG emissions were recently assessed during conventional or modified activated sludge processes in full-scale WWTPs (Tumendelger et al., 2019; Gruber et al., 2020; Ribeiro et al., 2020), studies dealing with N₂O emission from aerobic granular biomass fed with real domestic wastewater are still scarce (Guimarães et al., 2018; Jahn et al., 2019; Thwaites et al., 2021).

Studies that address this topic generally refer to synthetic sewage-fed systems (Wei et al., 2014; Reino et al., 2017) and are not representative of real domestic wastewater, which presents a more complex composition, in addition to load variations (Wang et al., 2018a). Recent studies refer to off-gas and dissolved N₂O and other GHG gases (CH₄, CO₂) monitoring in full-scale AGS WWTP (Baeten et al., 2021; Thwaites et al., 2021; van Dijk et al., 2021). These studies evaluated how different factors affect these emissions, such as seasonal temperature and rain fluctuations, wastewater salinity, continuous or sequence batch operational modes, dissolved oxygen (DO) concentrations, and organic loads.

Besides monitoring N₂O emissions, it is also useful to quantify CO₂ emissions produced by biomass growth through oxidation of biodegradable organic matter (Law et al., 2013; Jaromin-Gleń et al., 2020). When considering CO₂ production in WWTPs, a distinction should be made between indirect and direct emissions (Massara et al., 2017). Although degradation of organic matter is sometimes considered a zero emission process because the CO₂ produced is involved in the natural carbon cycle (Bartram et al., 2019), resulting in negligible impacts on the greenhouse effect, studies show that a fraction of influent organic carbon is not of biogenic origin. Part of influent organic carbon is derived from petroleum-derived products (cosmetics, drugs, personal care products, etc.), being known as fossil organic carbon (FOC).

FOC present in wastewater is considered a source of CO₂ emission, being one of the contributors to the net carbon footprint of wastewater treatment systems (Bao et al., 2015; Mannina et al., 2016). In biological wastewater treatment systems, CO₂ is generated by the degradation of organic matter by aerobic and anaerobic processes. Direct CO₂ emissions by WWTPs are considered short-lived biogenic carbon emissions, and therefore do not contribute to total GHG emissions (Bartram et al., 2019).

Therefore, both N₂O and CO₂ emissions should be monitored during BNR and organic carbon oxidation processes in wastewater treatment, respectively. Estimates of the production potential of these gases provide information on the treatment configuration effectiveness and give a better understanding of the processes involved, and of the complete wastewater cycle (Bao et al., 2015). To compare GHG emissions from wastewater treatment systems, it is appropriate to determine the gaseous emission volume per unit, that is, the amount of GHG emitted per cubic meter of treated effluent (Pascale et al., 2017).

It is also important to consider the effects of temperature on the distribution of N₂O between the liquid phase and the gaseous phase. Due to the fact that N₂O solubility decreases with increasing temperature, it was expected that the percentage of dissolved N₂O in the system effluent would be higher in locations with colder weather, or at times of the year with lower temperatures (Reino et al., 2017; Bao et al., 2018).

In this context, the objective of this study was to evaluate and quantify the behavior of N₂O and CO₂ emissions from an SBR-AGS for the treatment of domestic wastewater. The treatment performance was evaluated through physicochemical and biological parameters and by following the granular biomass development. Quantifications of N₂O (dissolved and gaseous) and CO₂ were made using online measurements during reactor operating cycles. In addition, the global warming impact (GWI) caused by these two compounds emitted by the biological reactor during the wastewater treatment was quantified by calculating the flow-based emission factors (FBEF).

Materials and Methods

Experimental setup

The pilot reactor (SBR-AGS) had a cylindrical shape and acrylic composition, 2.18 m high, with an internal diameter of 25 cm and a working volume of 107 L. The volumetric exchange rate was 65%. Aeration was performed by a compressor (Wayne Wetzel professional-WV 15, 230L–3 hp; Wayne-Shultz, Brazil) in which the compressed air passed through the airline equipped with filters, pressure regulating valves, and a rotameter. The air flow of 32 L_air/min was supplied through a circular membrane diffuser (EPDM—340 mm diameter; B&F Dias, Brazil) positioned at the bottom of the reactor (Daudt et al., 2019).

The influent was pumped from the sewer system of Florianópolis, SC, Brazil (27°35′49″S/48°32′56″W), by means of a submerged pump (Schneider BSC—94—¾ CV 60 Hz), to a 5,000 L storage tank. The wastewater flowed by gravity to a 1,000 L buffer tank with a mechanical stirrer, to feed the biological reactor.

The SBR-AGS worked with 6-h operating cycles, with the following phases: plug-flow feeding at the reactor's bottom (1 h); anoxic period (30 min); aerobic period (3 h 56 min); settling (30 min); and discharge (4 min), all established according to Guimarães et al. (2018). The authors recommended the slow feeding and the anoxic (idle) phase to provide putative anaerobic conditions to select for phosphate-accumulating organisms and to enhance the denitrification. During the anoxic phase, aeration pulses (10 s) were applied every 15 min to keep the biomass in suspension, preventing the granules from settling. The pulses had a very brief duration to avoid causing interference in the monitoring of gases. There was no biomass inoculation during the reactor start-up.

The average sludge retention time (SRT) was 24 ± 10 days. Considering the reactor dimensions and the cycle configuration, the hydraulic retention time was 10.69 h, and the minimum settling velocity was 3.74 cm/min. Biomass developed naturally through operation. The reactor operated at room temperature (17–22°C) without pH control.

Treatment performance and biomass characterization

After a start-up period of 4 months (Schambeck et al., 2020), the reactor was monitored over a course of 106 days, when the system had already reached a stable operation. Mature granules (50% diameter ≥210 μm) (Wagner and Da Costa, 2013) were used, and there were only small fluctuations in performance, in terms of the parameters considered in this research.

Sample collections were performed weekly, resulting in 16 samples collected during the reactor monitoring period, with the following physical-chemical and biological parameters, measured according to APHA (2017): DO, pH, redox potential, and temperature measured by a multiparameter probe (YSI 6820 V2; Xylem); and total suspended solids and volatile suspended solids (VSS) measured by the gravimetric method. The physicochemical parameters were biochemical oxygen demand (BOD, model BOD-Track; Hach^®); chemical oxygen demand (COD, DR/4000; Hach); total phosphorus (TP) (molybdovanadate method, SM4500P), and nitrogen series: total and ammoniacal (TN and NH₄⁺-N, Set: 2714100 and Set: 2606945; Hach), and nitrite and nitrate (NO₂⁻-N, NO₃⁻-N, Dionex™ ICS-5000; Thermo Scientific). These were determined according to APHA (2017).

The granular biomass development was monitored by using the sieving method described by Laguna et al. (1999). The granular classification followed Liu et al. (2010), who consider biomass to be predominantly granular when the diameter of at least 50% of the particles is greater than 200 μm. The biomass settleability was measured by the sludge volumetric index (SVI), using the methodology developed by Schwarzenbeck et al. (2004), for settling times of 5, 10, and 30 min. The morphology and structure of the granules were observed by optical microscopy using an inverted microscope (Bel Photonics, Italy).

N₂O and CO₂ emission

Gaseous N₂O was measured using a gas analyzer (Guardian SP, Edinburgh, United Kingdom), based on dual wavelength nondispersive infrared technology, with a detection range from 0 to 1,000 ppm and an accuracy of 2 ppm. The N₂O concentration data were collected at intervals of 15 s. The dissolved N₂O was measured using a probe (Unisense Environment A/S, Denmark), and the values were indicated by the equipment in “mg N₂O-N/L.” The sensor presented its reading ranging from 0 to 1.5 mg N₂O-N/L, with a detection limit of 0.005 mg N₂O-N/L.

The CO₂ emitted by the reactor was measured by a carbon dioxide analyzer (model C-02; Instrutherm Measuring Instruments, Brazil). Its operation relies on the natural properties of CO₂ molecules to absorb light at a specific wavelength, which is proportional to the CO₂ concentration. The reading range of the CO₂ analyzer was from 0 to 6,000 ppm, with 1 ppm resolution and 3% accuracy. The equipment had an automatic data collection mode (data logger), enabling the recording of results once every second. The value of the atmospheric concentration of CO₂ (blank) was discounted from the value marked by the gas analyzer, so that only the CO₂ generated in the treatment process was accounted for.

A total of nine measurements were performed during the reactor monitoring period, resulting in the mean and the standard deviations result presented in the following sections.

The EF was calculated according to Equation (1). $E F (%) = \frac{m_{N_{2} O - N}}{m_{T N}} \times 100$ (1)

EF = emission factor [%];

$m_{N_{2} O - N}$ = emitted N₂O mass [g N₂O];

m_TN = total nitrogen mass [g N₂O];

GWI of N₂O and CO₂

In addition to separately quantifying the N₂O and CO₂ EFs, a comparison was made between the GWI for each of these GHGs. First, it was necessary to determine the N₂O FBEF in terms of CO₂-equivalent, that is, the N₂O EF was expressed in a standard unit, with CO₂ being the reference gas. This standardization of units allows the comparison of EFs from different gases. The calculation of the N₂O FBEF expressed in terms of CO₂-equivalent [] was made according to Equation (2). Factor 265 was used, according to data from the Fifth Assessment Report (AR5) of the IPCC (2014). This conversion factor indicates that for the same quantity of N₂O and CO₂, the amount of heat stored by N₂O is 265 times higher than the amount stored by CO₂, over a 100-year period. $F B E F_{N_{2} O (C O_{2} - e q)} = F B E F_{N_{2} O} \times 265$ (2)

$F B E F_{N_{2} O (C O_{2} - e q)}$ = flow-based emission factor of N₂O expressed in terms of CO₂-equivalent [g CO₂-eq/L];

$F B E F_{N_{2} O}$ = flow-based emission factor of N₂O [g N₂O/L];

265 = conversion factor according to IPCC (2014).

The total flow-based emission factor (FBEF_Total) was calculated according to Equation (3). $F B E F_{T o t a l} = F B E F_{N_{2} O (C O_{2} - e q)} + F B E F_{C O_{2}}$ (3)

FBEF_Total = total flow-based emission factor [g CO₂-eq/L];

$F B E F_{N_{2} O (C O_{2} - e q)}$ = flow-based emission factor of N₂O expressed in terms of CO₂-equivalent [g CO₂-eq/L];

$F B E F_{C O_{2}}$ = flow-based emission factor of CO₂, in [g CO₂/L].

The fraction of the GWI corresponding to each GHG (N₂O and CO₂) was determined from Equations (4) and (5), respectively. $G W I_{N_{2} O} (%) = \frac{F B E F_{N_{2} O (C O_{2} - e q)}}{F B E F_{T o t a l}} \times 100$ (4)

$G W I_{N_{2} O} (%)$ = global warming impact proportional to N₂O [%];

$F B E F_{N_{2} O (C O_{2} - e q)}$ = flow-based emission factor of N₂O expressed in terms of CO₂-equivalent [g CO₂-eq/L];

FEBF_Total = total flow-based emission factor [g CO₂-eq/L].

G W I_{C O_{2}} (%) = \frac{F B E F_{C O_{2}}}{F B E F_{T o t a l}} \times 100

(5)

$G W I_{C O_{2}} (%)$ = global warming impact proportional to CO₂ [%];

$F B E F_{C O_{2}}$ = flow-based emission factor of CO₂ [g CO₂/L].

FBEF_Total = total flow-based emission factor [g CO₂-eq/L].

The complete calculations to quantify gaseous and dissolved N₂O as well as CO₂ are described in the Supplementary Material S1.

Results and Discussion

Treatment efficiency and biomass characterization

Table 1 presents the characteristics of the influent and effluent wastewater, removal efficiencies, and physical characteristics of the biomass. Over the 106 days of the monitoring period, loads of 0.70 ± 0.09 kg TCOD/(m³·day) and 0.07 ± 0.01 kg NH₄⁺-N/(m³·day) were applied to the reactor.

Table 1.

Influent and Effluent Characterization, Removal Efficiencies, and Biomass Characteristics During the Experimental Period (106 Days)

Wastewater (n = 16)	Influent	Effluent	Removal efficiency (%)	Biomass (n = 16)
pH	6.8 ± 0.16	5.9 ± 0.10	—
Temp, °C	20.7 ± 2.0	21.6 ± 1.5	—	TSS, g/L	1.9 ± 0.31
TSS, mg/L	129 ± 19	33 ± 9	74 ± 9	VSS, g/L	1.7 ± 0.40
TCOD, mg/L	355 ± 47	78 ± 21	78 ± 6	SVI₅, mL/g	73.4 ± 9.3
SCOD, mg/L	170 ± 32	37 ± 9	77 ± 6	SVI₁₀, mL/g	59 ± 4.7
BOD₅, mg/L	248 ± 55	37 ± 16	84 ± 8	SVI₃₀, mL/g	47 ± 4.3
NH₄⁺-N, mg/L	36 ± 7	11 ± 4.0	65 ± 18	SVI₃₀/SVI₁₀	0.80 ± 0.07
NO₂⁻-N, mg/L	0.2 ± 0.1	1.1 ± 0.5	—
NO₃⁻, N, mg/L	0.2 ± 0.1	26 ± 8.0	—	Granule, %	56
TP, mg/L	5.0 ± 2.0	3.0 ± 1.0	40 ± 5.0	Ø > 212 μm

Ø refers to % of diameter; n refers to the analyzed samples.

BOD, biochemical oxygen demand; SCOD, soluble chemical oxygen demand; SVI, sludge volumetric index; TCOD, total chemical oxygen demand; TP, total phosphorus; TSS, total suspended solids; VSS, volatile suspended solids.

The values of the influent wastewater suggest low-strength sanitary wastewater (SCOD = 170 mg/L; BOD = 248 mg/L) according to Metcalf and Eddy (2003). The reactor was efficient in removing carbonaceous organic matter (78% total chemical oxygen demand (TCOD), 77% soluble chemical oxygen demand (SCOD) and 84% BOD₅). The TP influent was according to medium strength values for domestic wastewater (5 mg TP/L) (Metcalf and Eddy, 2003) and removal was partially achieved (40%). The average ammoniacal nitrogen removal was 65%, with an effluent concentration of 11 mg NH₄⁺-N/L. The reactor's performance attended the Brazilian standards for the wastewater discharge legislation (≤120 mg BOD/L; ≤20 mg NH₄⁺-N/L). The high nitrate values in the treated effluent (26 mg/L) show the occurrence of nitrification (Vasilaki et al., 2019).

During the reactor start-up, AGS development was observed even with conditions of low applied loads (0.70 ± 0.09 kg TCOD/[m³·day] and 0.07 ± 0.01 kg NH₄⁺-N/[m³·day]), without the addition of an external carbon source, and without biomass inoculation. The average biomass concentration in the SBR-AGS (1.7 ± 0.40 gVSS/L) was below that reported in other studies, which showed concentrations up to 6 gVSS/L (Poot et al., 2016; Pronk et al., 2017). However, this result is consistent with the system's feeding conditions, that is, real domestic wastewater, without any external source of carbon or nutrients. Granules with a diameter greater than 212 μm represented 56% of the reactor's biomass (Fig. 2), which, according to Liu et al. (2010), characterizes a predominantly granular reactor. The major fraction of the granular biomass (38%) was in the range of 212–300 μm.

Optical microscopy of the granular biomass on days 8 and 99 (Fig. 1) shows the biomass structural stability from the beginning to the end of the reactor operation. The SVI values (<80 mL/g; SVI₃₀/SVI₁₀ = 0.80 ± 0.07) (Table 1) confirm the presence of dense and compact biomass, with good settleability (Wagner and Da Costa, 2013).

FIG. 1.

Particle size distribution of the granular biomass during the experimental period and granular sludge microscopy on days 8 and 99.

During the monitored period, most of the developed granules had an irregular star shape and filamentous growth on the surface. The high fraction of particulate organic matter present in real wastewater, the applied shear stress, the available carbon and nutrients, and the reactor's configuration may have been the major factors responsible for that (Wagner et al., 2015; Franca et al., 2018; Xavier et al., 2021).

Figure 2 shows the pH and DO profiles and the nitrogen compounds (ammonia, nitrite, and nitrate) during a standard cycle of the SBR-AGS. Throughout the reactor cycle, the DO values were below 0.5 mg/L during the filling and the anoxic phases, followed by a variation between 8 and 10 mg/L during the aerobic phase. The redox potential was measured in the filling (anoxic/anaerobic) and the reaction (aerobic) phases, for each of the nine analyzed cycles. During the filling phase, the redox potential value was in the range of −70 ± 50 mV, followed by the aerated phase in the range of +200 ± 50 mV. As for the pH, there was a reduction of 7.3–5.1 during the aerobic phase, this drop being associated with the nitrification process due to the consumption of alkalinity (Hoffmann et al., 2007).

The profiles observed for the nitrogen series (Fig. 2B) show that nitrate and ammonia were reduced in the filling phase, probably due to influent dilution effects. The nitrification occurred from the beginning of the aerobic phase, simultaneous with the pH drop (Fig. 2A, B). In this phase, the production of nitrite (5 mg NO₂-N/L) and its oxidation to nitrate (26 mg NO₃-N/L) were observed, indicating nitrification, with a predominant nitrite-nitrification when ammonia concentration reached around 10 mg NH₄⁺-N/L.

FIG. 2.

Average cycle profile for pH, DO (A), and nitrogen compounds (B) during the reactor operation. DO, dissolved oxygen.

In most of the monitored period, the temperatures were between 15°C and 25°C, allowing the presence of ammonium oxidizing bacteria (AOB) in the system. According to Bao et al. (2018), when temperatures increase from 10°C to 30°C, AOB present higher ammonium oxidation activity, increasing the nitrification.

The SCOD/TN ratio was 3.14, with low-strength influent accounting for the reduced C/N ratio. A minimum stoichiometric COD/TN ratio of 3.5 is proposed by Verstraete and Philips (1998), while other proposed values range between 3 and 6 (Lahdhiri et al., 2017). According to Metcalf and Eddy (2003), higher COD/TN ratios result in reduced nitrification and increased denitrification, since more carbon is available for biological nitrogen removal. It is worthwhile to consider that the idle (anoxic) phase just after the feeding phase was used to increase denitrification, besides improving the phosphorus biological removal. However, reduced denitrification was obtained since low nitrite and nitrate concentrations were obtained due to their dilution in the influent.

Operational conditions, such as cycle duration (6 h), alkalinity (240 mg CaCO₃/L), DO (>8 mg/L), pH (∼7.5), temperature (∼18°C), and SRT (24 days), prevented the occurrence of nitrite accumulation, commonly observed in other studies with AGS (Hoffmann et al., 2007; Reino et al., 2017; Guimarães et al., 2018), which is undesirable to minimize N₂O generation (Kampschreur et al., 2008).

CO₂ and N₂O emission profile

Figure 3 shows the variation in CO₂ and N₂O (dissolved and gaseous) over a standard SBR-AGS cycle. In Fig. 3A, it is possible to observe the CO₂ emissions over an operational 6-h cycle of the SBR-AGS. Similar to that verified for N₂O emissions (Fig. 3B), CO₂ emission was not constant during the operation cycle. Occasional emissions occurred when aeration pulses were applied (during the anoxic phase), with the highest peak in the first minutes of the aeration phase (12.6 mg CO₂/s). The emissions during the pulses were insignificant compared to the total emissions, which was expected, as the pulses served only to maintain the biomass in suspension. After the peak, the emission gradually decreased, varying between 3.6 and 0.7 mg CO₂/s.

FIG. 3.

Average SBR cycle profile for emitted CO₂ (A), dissolved and emitted N₂O-N (B), and accumulated N₂O-N released (C). CO₂, carbon dioxide; N₂O, nitrous oxide; SBR, sequencing batch reactor.

The CO₂ emission during the entire aerobic phase was consistent, since most of the carbon removal occurs during this phase. Thus, the CO₂ emission is directly associated with the metabolism of aerobic microorganisms (Pascale et al., 2017). When aeration ceases, CO₂ emission stops, as there is no airflow in the system.

In Figure 3B, it can be seen that during the filling and anoxic phases (when there was no air flow in the reactor), an increasing dissolved N₂O formation occurred in the liquid medium, reaching a peak of 9.0 mg N₂O-N/L, while the N₂O emission was zero. At the beginning of the aerobic phase, due to the agitation caused by intense aeration, the dissolved N₂O concentration in the mixed liquor dropped to values below 0.5 mg N₂O-N/L and was released into the atmosphere in the form of gaseous N₂O. Values of ∼0.82 mg N₂O/s were achieved, after which the emission decreased until reaching values close to zero.

Similar profiles were also observed by Bao et al. (2018) in two SBRs, and by Vieira et al. (2019) in three full-scale biological reactors. The authors attributed the generation of dissolved N₂O to the heterotrophic denitrification process, with consequent gas emission in the moments of aeration, corroborating the results of this research.

Ding et al. (2017) also reported N₂O generation with DO concentration. According to these authors, when the DO concentration was kept at low levels, oxidation of NH₄⁺ to NO₂⁻ occurred, with the generation of N₂O by nitrifying denitrification. When the oxygen became excessive (>7.5 mg/L), due to the inhibition of denitrification, the emission of N₂O was reduced and the generation of NO₃⁻ increased. In this study, from the beginning of aeration, DO concentrations were already quite high (>7.0 mg/L), not favoring the occurrence of nitrifying denitrification. This could explain the low N₂O emission observed during the aerobic phase, just after the observed peak emission.

According to Han et al. (2019) and Yang et al. (2013), N₂O released during the aeration phase is due to the action of denitrifying microorganisms in the anoxic phase of the reactor cycle. Because there is no airflow during the anoxic phase, the generated N₂O accumulates, and is retained (dissolved) in the system. Mello et al. (2013) measured the N₂O emission by an activated sludge treatment system with intermittent aeration and found, in accordance, that less than 1% of the amount of generated N₂O was emitted to the atmosphere in the absence of aeration.

It is important to note that, in this study, small amounts of N₂O were released during the brief aeration pulses, which occurred in the anoxic phase. These punctual emissions, however, were much lower than the overall N₂O emission during the 6-h cycle. In fact, the only reason the aeration pulses were performed was to maintain the biomass in suspension, avoiding sedimentation of granules.

N₂O production in the liquid phase and emission of N₂O to the gas phase were also monitored by Reino et al. (2017). The authors investigated the effects of temperature on N₂O dynamics in an SBR with partial nitrification. When a temperature of 20°C was applied, the N₂O emission was 2.5 times higher than the emission obtained at 10°C. The authors suggested the hypothesis that the penetration of oxygen into the granules is influenced by temperature. In lower temperature conditions, there is an increase in the depth of oxygen penetration, which would cause a smaller fraction of AOB to be subjected to anoxic conditions. This would reduce the generation of N₂O from the metabolic route of nitrifying denitrification.

In addition, according to the authors, N₂O emission was higher for higher temperatures, because besides greater generation of this compound, a more intense stripping process also occurs. Similar results were obtained by Bao et al. (2018), who observed an increase in AOB activity due to temperature increase (between 10°C and 25°C) and the consequent increase in N₂O emission. However, the N₂O production was mainly due to heterotrophic denitrification, when the DO level was below 0.5 mg/L.

The accumulated N₂O released by the SBR-AGS in a standard cycle is shown in Fig. 3C. The highest level of N₂O accumulation occurred in the first minutes of the aeration phase, as previously verified in Fig. 3B. In this period, the accumulated amount reached 0.35 g N₂O. Subsequent ly, it was still possible to verify the increase in the accumulated mass of N₂O released, although with less intensity. This result corroborates what was reported by Vieira et al. (2019), where 95% of N₂O emissions measured in a WWTP occurred in the aeration tanks. In the settling phase, the accumulation of N₂O emitted was null, since there was no airflow in the system.

The last contribution to the amount of accumulated N₂O occurred at the end of the SBR-AGS cycle, marked by discharge of the effluent. This contribution refers to the dissolved N₂O contained in the treated effluent that was discharged from the system.

The SBR-AGS achieved low TN removal (29%), and most of the effluent nitrogen corresponded to nitrate (26 mg NO₃-N/L). Of the total denitrified nitrogen, the emission of N₂ exceeded the emission of N₂O by a ratio of ∼4:1. The percentage of total nitrogen converted to N₂ was 77.1% ± 4.3%, while the nitrogen denitrified to N₂O (22.9% ± 4.3%) was within the range obtained by Foley et al. (2010), who reported percentages between 0.06% and 25.3% in different effluent treatment systems. Likewise, Ge et al. (2017), when studying a biofilm SBR, found that 15.47% of the nitrogen removed by denitrification was converted to N₂O and 72.25% to N₂.

The hydroxylamine, nitrifying denitrification, and the heterotrophic denitrification are the three known N₂O production pathways. The heterotrophic denitrification seems to be the most probable N₂O production pathway observed in this study, attributed to the high DO concentration applied in the aeration phase, causing the nitrate accumulation. Also, high DO concentration stimulates AOB nitrite isomeric reductase enzyme (Nir) for NO₂-N consumption to generate N₂O during the oxidation of NH₄⁺ and NH₂OH (Sun et al., 2018).

The anaerobic feeding and the idle phase may have provided the required conditions for heterotrophic denitrification (Vasilaki et al., 2019). However, the low carbon concentration due to the diluted nature of the influent may had impacted the complete denitrification, with the generation of N₂O. Other pathways have been related to NO₂⁻ accumulation, which was not observed in this study.

Although low DO concentrations have been proposed for simultaneous nitrification/denitrification in AGS systems, several authors related oxygen-limited conditions to N₂O emissions (Baeten et al., 2021; Thwaites et al., 2021). Low DO concentrations (0.1–0.3 mg/L) combined with high ammonium can respond for high N₂O emissions through the nitrifying denitrification pathway, due to incomplete reduction of nitrate to nitrogen gas.

Dijk et al. (2021) observed that a full-scale WWTP working with a dynamic DO set point resulted in higher N₂O emission compared to a fixed set point (2.5 mg/L). Dynamic oxygen set point reduced the oxygen supply to optimize the nitrification/denitrification, but resulted in increased N₂O emissions, due to increased microbial DO consumption during organic loading peaks. Thwaites et al. (2021) suggested that increased DO concentrations reduce the N₂O emissions due to limited nitrite accumulation.

The simultaneous nitrification/denitrification is a possible advantage offered by AGS, due to its well-established granular structure and biofilm nature, which provide the required anoxic zones for denitrification even in the aerobic phase. The small granules obtained during the reactor operation (<200 μm) may not have structured anoxic layers and may not perform the simultaneous nitrification/denitrification. Preliminary studies with the same reactor system used in this work (Daudt et al., 2019) emphasized the importance of an anoxic/anaerobic phase before aeration to ensure denitrification. The results indicated that the extension of the anoxic phase was an effective way to significantly reduce N₂O emission and improve treatment efficiency.

In AGS with simultaneous nitrification and denitrification, the N₂O can be emitted by different pathways, making it difficult to distinguish which pathway is the major contribution to N₂O emissions. The combination of fluctuating influent concentrations and the variety of N₂O formation pathways results in complex solutions to avoid N₂O emissions (Vasilaki et al., 2019; Dijk et al., 2021). Excessive DO concentration may result in increased OPEX (operating expenses) and C-footprint, so dynamic control of DO concentration may provide better conditions to avoid N₂O emissions and efficiently remove nitrogen from the influent.

The EF (5.67%) was higher than the IPCC default value for aerobic treatment plants (Bartram et al., 2019), as can be seen in Table 2. The IPPC standard value is contained in an EF range from 0.016% to 4.5%, based on 30 full-scale activated sludge systems. Vasilaki et al. (2019) reported an EF range from 0.0025% to 5.6% based on 51 full-scale WWTPs analyzed during 10 years. According to the authors, SBRs generally present higher EF compared to other treatment systems, ranging between 2% and 5.6%.

Table 2.

Conversion of Influent Total Nitrogen to Nitrous Oxide Gas (Emission Factor %) (Mean ± Standard Deviation) and Comparison with the Literature

Configuration	Reactor scale	N₂O EF^a (%)	Reference
AGS-SBR	Pilot scale	5.67 ± 1.68	This study (n = 9)
Aerobic process	—	1.6	Bartram et al. (2019)
CAS	Full scale	0.054	Ribeiro et al. (2020)
CAS	Full scale	0.001	Tumendelger et al. (2019)
SBR	Full scale	2–5.6	Vasilaki et al. (2019)
SOA-SBR	Bench scale	1.51–4.32	Chen et al. (2020)
Various	Bench and full scale	0.001–14.6	Kampschreur et al. (2009)
AGS-SBR	Bench scale	3.93–8.07	Chen et al. (2011)
SBR	Bench scale	2.1	Su et al. (2017)
SBR	Bench scale	1.1	Blum et al. (2017)
Nitritation Anammox	Full scale	2.3	Kampschreur et al. (2008)

EF (%) = (kg N₂O-N/kg TN)^*100.

AGS, aerobic granular sludge; CAS, conventional activated sludge; EF, emission factor; N₂O, nitrous oxide; SBR, sequencing batch reactor; SOA, static/oxic/anoxic; TN, total nitrogen.

Higher N₂O EFs for SBRs are attributed to sudden changes in NH₄⁺ and NO₂⁻ concentrations within the cycle or the accumulation of dissolved N₂O during the anoxic phase. Ribeiro et al. (2020) observed an EF of 0.054% in a full-scale conventional activated sludge (CAS) system, whereas Chen et al. (2020) found EF values varying from 1.51% to 4.32% in a static/oxic/anoxic treatment process. Sun et al. (2013) obtained an EF of 6.52% in a full-scale SBR, and Chen et al. (2011) noted EF values varying between 8.07%, 5.37%, and 3.93% for an AGS reactor. Kampschreur et al. (2009) reported EF values ranging from 0.001% to 14.6% for different treatment systems.

According to Table 2, lower EF values are related to full-scale reactors, while higher EF is observed for pilot-scale or laboratory-scale experiments. Full-scale reactors usually work with lower DO concentration when compared to that in bench- and pilot-scale studies.

CO₂ and N₂O (emission and release) and the GWIs of SBR-AGS

Table 3 shows the CO₂ emission values. The $F B E F_{C O_{2}}$ was 0.297 g CO₂/L, and can be compared with the values reported by Flores-Alsina et al. (2011). The authors reported emissions of 0.191 g CO₂/L relative to the biomass breathing processes, and 0.219 g CO₂/L relative to the oxidation of BOD, with a total emission equal to 0.410 g CO₂/L. Singh and Kansal (2018) also evaluated GHG emissions in 10 WWTPs located in New Delhi (India) and reported emissions ranging from 0.2881 to 0.6233 g CO₂/L.

Table 3.

Values of Carbon Dioxide and Nitrous Oxide Emission and Release Parameters (Mean ± Standard Deviation) (n = 9)

CO₂ gaseous fraction	Emission/cycle (g CO₂/cycle)16.4 ± 1.56	Daily emission (g CO₂/day)65.4 ± 6.22	$P B E F_{C O_{2}}$ (g CO₂/[person·year])1.74 × 10⁴ ± 0.16 × 10⁴	$F B E F_{C O_{2}}$ (g CO₂/L)0.297 ± 0.028
N₂O gaseous fraction	Emission/cycle (g N₂O-N/cycle)0.201 ± 0.075	Daily emission (g N₂O-N/day)0.804 ± 0.317	$P B E F_{N_{2} O}$ (g N₂O-N/[person·year])213.43 ± 79.63	$F B E F_{N_{2} O}$ (g N₂O-N/L)3.65 × 10⁻³ ± 1.36 × 10⁻³
N₂O liquid fraction	Release/cycle (g N₂O-N/cycle)0.006 ± 0.002	Daily release (g N₂O-N/day)0.024 ± 0.008	$P B R F_{N_{2} O}$ (g/[person·year])6.26 ± 2.08	$F B R F_{N_{2} O} (g ∕ L)$ 0.11 × 10⁻³ ± 0.03 × 10⁻³
Total N₂O (gaseous+liquid)	Emission+release/cycle (g N₂O-N/cycle)0.207 ± 0.079	Daily emission+daily release (g N₂O-N/day)0.828 ± 0.316	$P B E F_{N_{2} O}$ + $P B R F_{N_{2} O}$ (g N₂O-N/[person·year])219.69 ± 73.23	$F B E F_{N_{2} O}$ + $F B R F_{N_{2} O}$ (g N₂O-N/L)3.76 × 10⁻³ ± 1.25 × 10⁻³

CO₂, carbon dioxide; FBEF, flow-based emission factor; FBRF, dissolved N₂O flow-based release factor; PBEF, population-based emission factor; PBRF, population-based release factor.

The average N₂O emitted as gas was 0.804 g N₂O-N/day. Considering the volume of treated effluent in each cycle, and assuming a per capita effluent generation of 160 L/(person·day) (Metcalf and Eddy, 2003), there is a $P B E F_{N_{2} O}$ of 213.43 g N₂O-N/(person·year).

Although the observed $P B E F_{N_{2} O}$ was higher than that proposed by the IPCC (2006), it is in accordance with values obtained in other studies involving N₂O emission by activated sludge systems, such as Daelman et al. (2013). They found a $P B E F_{N_{2} O}$ of 163.2 g N₂O-N/(person·year) when analyzing the N₂O emission of a municipal activated sludge WWTP, this value being 80 times higher than the $P B E F_{N_{2} O}$ proposed by the IPCC. It is important to note that the $P B E F_{N_{2} O}$ proposed by the IPCC was based on one single study, carried out by Czepiel et al. (1995), at a WWTP located in Durham (northern United States), which experiences low temperatures. The updated IPCC report (Bartram et al., 2019) adjusted the N₂O EF to 1.6%, in terms of kg N₂O-N/kg TN.

The N₂O released in the liquid fraction corresponded to 2.9%, while the N₂O emitted to the gas fraction contributed to 97.1% of the total amount of N₂O produced by the system. It is important to consider the effects of temperature on the distribution of N₂O between the liquid phase and the gaseous phase, which explains the difference between the percentages obtained. This study was carried out in a subtropical climate, where average temperatures were around 20°C. Due to the fact that N₂O solubility decreases with increasing temperature, it was expected that the percentage of dissolved N₂O in the system effluent would be higher in locations with colder weather, or at times of the year with lower temperatures.

The $F B E F_{N_{2} O}$ , expressed in terms of CO₂-equivalent, was 1.520 g CO₂-eq/L, while the $F B E F_{C O_{2}}$ presented a significantly lower value of 0.297 g CO₂-eq/L. The GWIs for N₂O and CO₂ were based on the CO₂-equivalent FBEF and are shown in Fig. 4. Of the total GWI corresponding to both N₂O and CO₂ emissions, there was a $G W I_{N_{2} O}$ of 83.6% and a $G W I_{C O_{2}}$ of 16.3%. It is clear therefore that, although the CO₂ mass emission was higher (Table 3), the N₂O impact corresponded to more than 80% of the carbon footprint emitted by the SBR-AGS (considering only N₂O and CO₂). This is largely because N₂O has a GWP around 265 times higher compared with CO₂ (IPCC, 2006).

FIG. 4.

CO₂-equivalent FBEF and GWI from the CO₂ and N₂O emissions. FBEF, flow-based emission factor; GWI, global warming impact.

The results verified in this research are in accordance with what was verified by Bao et al. (2016), in a study involving the assessment of GHG emissions (N₂O, CO_2, and CH₄) by two real-scale wastewater treatment systems. For both systems, it was found that N₂O emissions were the main sources of direct GHG emissions, being responsible for 88.5% and 90.0% of emissions from the A/O and SBR processes, respectively. Daelman et al. (2013) observed that N₂O contributed to 78.4% for the carbon footprint, while the CO₂ contribution was 8.1%. Including operational parameters as electricity, the total carbon footprint, including N₂O contributions, can increase to 97% (Vasilaki et al., 2020). The authors found that 48 kg of CO₂-equivalent are generated for the removal of 1 kg of NH₄⁺ from the wastewater, including electricity and direct N₂O emission.

This relatively higher impact of N₂O in comparison to CO₂ emissions indicates that, when aiming to minimize the GWI from a wastewater treatment system, it is especially important to achieve a BNR in which the N₂O formation is avoided or, at least, kept to a minimum.

Conclusions

The operational conditions applied to the SBR allowed the development of AGS (1.7 ± 0.40 g VSS/L) with good settleability (SVI₃₀ < 50 mL/g). The treatment performance showed the following removal percentages: 84% of BOD₅ and 65% of NH₄⁺-N, attending the Brazilian standards for the wastewater discharge legislation.

The CO₂ emissions were observed during the entire aerobic phase in the SBR cycle, when most of the carbon removal occurred due to aerobic metabolism. However, dissolved N₂O occurred mainly during the filling and anaerobic phases, due to the heterotrophic denitrification process, followed by gaseous N₂O stripping to the atmosphere during aeration. The structural singularity of AGS compared to CAS flocs results in DO gradients, providing conditions for several N₂O formation pathways.

Although CO₂ flow-based emission was higher, the N₂O carbon footprint measured in terms of CO₂-equivalent and the GWI was five times higher than the impact of CO₂ emissions. N₂O emissions were responsible for more than 80% of the total impact, which reinforces the need to maintain monitoring of N₂O emissions in AGS wastewater treatment processes, to minimize the generation and emission of N₂O. This emphasizes the need for an integrated assessment of the treatment performance and the carbon footprint evaluation for SBR-AGS.

Footnotes

Acknowledgments

The authors thank the Environmental Integrated Laboratory (LIMA) and the Laboratory of Gaseous and Liquid Effluents (LABEFLU) of the Department of Environmental Engineering, at the Federal University of Santa Catarina, for the analytical and structural support.

Authors' Contributions

G.C.D. performed the methodology, investigation, data curation and analysis, and draft writing. B.S.M. performed the data obtention, methods, and data curation draft writing. C.M.S. performed the data obtention, draft writing, and draft analysis. N.L.J. worked on the article writing, data analysis, and discussion. R.H.R.C. supervised this research, including project administration and final article writing.

Author Disclosure Statement

No competing financial interests exist.

Funding Information

This work was supported by Coordenação de Aperfeiçoamento de Pessoal de Nível Superior–Brasil (CAPES) (Finance Code 001) and Conselho Nacional de Desenvolvimento Científico e Tecnológico (CNPq) (Finance Code 404144-2016-0), and by internal funds of Federal University of Santa Catarina.

Supplementary Material

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Supplementary Material

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Assessment of Nitrous Oxide and Carbon Dioxide Emissions and the Carbon Footprint in an Aerobic Granular Sludge Reactor Treating Domestic Wastewater

Abstract

Introduction

Materials and Methods

Experimental setup

Treatment performance and biomass characterization

N2O and CO2 emission

GWI of N2O and CO2

Results and Discussion

Treatment efficiency and biomass characterization

CO2 and N2O emission profile

CO2 and N2O (emission and release) and the GWIs of SBR-AGS

Conclusions

Footnotes

Acknowledgments

Authors' Contributions

Author Disclosure Statement

Funding Information

Supplementary Material

References

Supplementary Material

N₂O and CO₂ emission

GWI of N₂O and CO₂

CO₂ and N₂O emission profile

CO₂ and N₂O (emission and release) and the GWIs of SBR-AGS