Main

Addressing the challenge of global warming requires joint efforts from all sectors to achieve carbon reduction1,2,3. While the energy and manufacturing sectors will continue to play a dominant role in the future path to decarbonization, the potential of many other sectors has also been increasingly recognized4,5. However, the wastewater sector has been relatively overlooked so far, despite the fact that wastewater treatment plants (WWTPs) are widely recognized to be intensive energy consumers6. Notably, wastewater utilities are also non-negligible sources of fugitive greenhouse gas (GHG) emissions, with considerable amounts of methane, nitrous oxide and carbon dioxide being emitted during wastewater collection, treatment, environmental discharge and sludge disposal processes7. However, previous studies have mostly concentrated on the emissions of one or several specific WWTPs8,9,10, while a national-scale detailed understanding of the carbon footprint and spatial distribution remain lacking for most countries, which impedes the execution of effective mitigation action.

In particular, the control of GHG emissions in China will have profound global impacts and deserves special attention. As the world’s largest carbon emitter, China also possesses the largest, and still growing, wastewater sector. Over the past years, this nation has been actively promoting the upgrading of wastewater utilities for enhanced water purification. This is arguably considered to aggravate GHG emissions by the wastewater sector through increased energy and chemical consumption, but detailed information regarding such impacts is still lacking. In addition, although several recent studies have attempted to explore the carbon footprint of China’s WWTPs, they either focused exclusively on energy-associated indirect emissions or neglected the fugitive emissions from sewers, effluent and sludge (Supplementary Table 1). More importantly, default emission factors (EFs), mainly in the Intergovernmental Panel on Climate Change (IPCC) framework11, have universally been adopted in previous estimations without taking into account the operational and regional discrepancies of wastewater utilities12,13,14. Such limitations not only caused inconsistent and biased estimations in previous studies (Supplementary Table 1), but also impeded a clear identification of the drivers underlying the emission heterogeneity of the wastewater utilities.

Here, we provide the results of the first detailed national-scale assessment and prediction of the spatiotemporal GHG emission patterns of the municipal wastewater sector in China. To accurately reveal the spatial diversity of wastewater GHG emissions, we assigned dynamic, plant-resolved EF values to different municipal wastewater utilities based on their case-specific operating parameters. Specifically, the operating data of 5,155 municipal WWTPs across a time span of 11 years and 45 groups of field-measured GHG emission data from the literature were adopted for the estimation. The system boundary for the estimation was expanded from the previous studies to cover the entire process of municipal wastewater management, from wastewater collection, treatment, environmental discharge to sludge disposal (Fig. 1a). The case-specific EFs for fugitive and indirect emissions at each stage of wastewater management were determined by taking into account the operational heterogeneity of wastewater facilities and the dynamic evolution of power grid emissions. Correlation analysis was performed to unravel the key factors driving the spatiotemporal variations in wastewater GHG emissions.

Fig. 1: System boundary and statistical analysis of GHG emissions generated during wastewater management.
figure 1

a, Diagram showing the processes involved in wastewater management and associated GHG emission sources. b,c, Definition (b) and statistics (c) of the EFs for fugitive GHG emissions produced at different stages of wastewater management. The EF values of emissions at each stage were further categorized according to the relevant key features of the adopted operating modes. EM, emission amount; t-DM, tonnes dry matter of sludge. A2O, anaerobic-anoxic-oxic process; SBR, sequencing batch reactor; AO, anoxic–oxic process; OD, oxidation ditch process. The WWTP influent, effluent and removed pollutants within WWTPs are denoted by the subscripts inf, eff and rem, respectively. For consistency, the unit of EF for the sludge disposal stage has been converted into emitted GHG per kg COD or TN in wastewater (that is, g GHG kg−1 CODrem or g GHG kg−1 TNrem). The points in each box are sampled from the original dataset. The box represents the middle two quartiles, the bold horizontal lines show the median and the centre square represents the average value.

Based on the above information, we next projected the future emission trends of China’s wastewater sector and evaluated its potential emission reduction under several envisaged paradigms. The future emission patterns of WWTPs may vary significantly depending on the implemented treatment technologies and the advances in decarbonization in the energy sector and chemical production industry15,16. Therefore, the impacts of key factors, including sector development planning, decarbonized energy and technology advances towards resource-oriented processes, were considered in the predictions based on the dynamic EFs determined. The GHG emission patterns and economic benefits of future wastewater treatment paradigms, assuming a gradual transition to resource-oriented processes, were predicted. Lastly, the future prospects and efforts required to enable the transition to these wastewater management paradigms in China are discussed.

EFs of fugitive GHG emission during wastewater management

Wastewater management activities generate both on-site fugitive emission from pollutant decomposition and indirect emission embodied in the energy and chemicals consumed during the wastewater collection, treatment, environmental discharge and sludge disposal processes (Fig. 1a). To estimate fugitive GHG emissions, the key challenge lies in determining the EF values, which describe the amount of GHG emissions generated on-site from the treatment of each kilogram of wastewater pollutant or sludge (Fig. 1b). The highly diverse treatment technologies and operating conditions as well as the implementation of non-standard monitoring methods in the reported case studies together have caused a huge discrepancy in the field-measured raw data of GHG emissions (Supplementary Tables 27). To acquire reliable EF values for municipal wastewater utilities, we categorized the raw data according to the different stages of wastewater management and the different operating modes adopted at each stage, for example, the sewer type for sewers, the treatment process for WWTPs and the approaches adopted for sludge disposal (Supplementary Note 1). Next, the raw data in each category were processed for unit standardization and probability distribution analysis by Monte Carlo simulation, which yielded the final EF values for subsequent calculation (Supplementary Fig. 1).

Here, the EF values for different types of GHGs were determined separately (Supplementary Note 1). In contrast to previous studies that only took the fugitive emission of CH4 and N2O into account, we also included fossil-based CO2 (fCO2), assuming that about 10% of the organic matter in municipal wastewater is of fossil origin17. The results showed substantially different EF values for the fugitive GHG emissions generated at different stages of wastewater management and under different operating modes (Fig. 1c). In particular, dumped sludge, landfilled sludge and effluent/untreated wastewater showed a relatively high intensity of CH4 fugitive emissions, likely attributable to the formation of an anaerobic local environment that favours methanogenic activity in these aspects of wastewater management.

Temporal evolution of wastewater GHG emissions in China

Multiplying the EF by the corresponding emission activity, which describes the wastewater properties or operations affecting GHG emissions, yields the volumetric GHG intensity for a specific unit of wastewater management. Due in part to the growing volume of municipal wastewater in China, the overall wastewater sector achieved 2.43-fold growth in treatment capacity during the period of 2009–2019. Meanwhile, due to the improved wastewater treatment ratio (from 75.25% to 95.5%; Supplementary Fig. 2), the volumetric intensity of GHG emissions from untreated wastewater decreased, while that from sewers increased (Fig. 2a). At the same time, the share of wastewater GHG intensity contributed by sludge disposal (normalized to per cubic metre wastewater according the sludge yield) decreased as a result of reduced sludge dumping and landfilling (Supplementary Fig. 3), the two most important sources of fugitive GHG emissions (Fig. 1c)11,18. These changes in emission activities, together with the slightly decreased wastewater pollutant concentrations (Supplementary Fig. 4)19, led to a 16.8% reduction in fugitive GHG intensity from 2009 to 2019 (Fig. 2b).

Fig. 2: Historical trend and distribution of GHG intensity in China’s municipal wastewater sector during the period of 2009–2019.
figure 2

a, Annual changes in wastewater emission intensity contributed by different wastewater management stages. b,c, Distribution of GHG emissions by emission type (b) and emission site (c) in the years 2009 and 2019.

Source data

The GHG intensity of indirect emissions by the municipal sector showed an opposite trend to that of fugitive emissions. Due to the implementation of enhanced wastewater treatment to meet the increasingly stringent WWTP effluent limits in China (Supplementary Figs. 2 and 4), the energy consumption per cubic metre wastewater increased by 31.7% during the period of 2009–2019, and the chemical consumption (mainly for enhanced nitrogen removal) also increased drastically (Supplementary Fig. 5). Accordingly, the GHG intensity of indirect emissions showed 81.5% growth during this period (Fig. 2a), despite the continuously declining grid emissions due to the greater supply of renewable energy (Supplementary Table 2). Altogether, the overall GHG intensity of China’s municipal wastewater sector increased by 17.2% from 2009 to 2019 (Fig. 2a), with the indirect emissions gradually becoming predominant (Fig. 2b). Notably, the share of wastewater GHG intensity contributed by the WWTPs, the predominant site of pollutant degradation, increased from 50.0 to 65.5% during the period of 2009–2019 (Fig. 2c), implying that the increased GHG emission might be highly associated with pollutant removal activities. This is further supported by the analysis of GHG intensity per pollutant removed from wastewater (Supplementary Fig. 6 and Supplementary Note 2).

Although the volumetric GHG intensity increased by only 17.2% from 2009 to 2019, the total amount of emissions from the municipal wastewater sector showed 2.4-fold growth (Supplementary Fig. 7), mainly driven by the drastic increase in wastewater volume (Supplementary Fig. 2a). The wastewater GHG intensity in 2019 reached 0.77 kgCO2e m3 (Fig. 2a), and the total amount of emissions increased to 53.0 MtCO2e, much lower than the values estimated in the IPCC framework (Supplementary Fig. 7b). These results suggest that the estimations of wastewater GHG emissions in many previous studies might be biased and need reconsideration.

Regional distribution of wastewater GHG emissions in China

It is noteworthy that, although the national average wastewater GHG intensity increased, the growth intensity and rate varied drastically between regions due to the huge operational heterogeneity of the various wastewater utilities (Fig. 3). Indeed, the wastewater GHG emission pattern for China clearly shows an uneven regional distribution (Fig. 3a). According to the 2019 data, higher GHG intensities were found in North China (1.14 kgCO2e m3 on average) and Northwest China (0.96 kgCO2e m3), both dominated by emissions from WWTPs and sludge disposal (87.8 and 84.7%, respectively; Supplementary Fig. 8). The GHG intensity in South China (0.58 kgCO2e m3) was only half that in North China (Supplementary Fig. 8b).

Fig. 3: Geographical distribution of WWTP operating status and GHG emissions in China’s municipal wastewater sector.
figure 3

a, City-resolved distribution of wastewater GHG intensity in different years. be, Regional distribution of WWTP emission activities, including removed COD (b), removed TN (c), electricity consumption (d) and sludge yield (e). f, Pearson correlation between key emission activities and wastewater GHG intensity. BODrem, removed BOD. g, Regional distribution of wastewater GHG intensity in different years. All the estimations were made on the basis of the average data for 2019 unless otherwise specified. The data are presented as the mean ± SD (n = 100,000). COD, chemical oxygen demand; BOD, biochemical oxygen demand; TN, total nitrogen; TP, total phosphorus.

Source data

Correlation analysis was performed to identify the key drivers for such spatial discrepancy. Our results show that the GHG intensity of WWTPs is highly correlated with the electricity consumption (correlation coefficient = 0.83), removed TN (0.74), removed COD (0.74) and sludge yield (0.64; Fig. 3f). This is consistent with the 43.4–58.5% higher values for all these activity data in North China compared with in South China (Supplementary Fig. 9), confirming that the pollutant removal activities of WWTPs critically determine their GHG emission behaviour. Notably, the regional heterogeneity in wastewater GHG intensity will be even greater by 2030, a 2.87-fold higher value in North China than in South China (vs. 1.93-fold higher value in 2019) (Fig. 3g). Therefore, the intensively emitting northern regions may deserve special attention. Nevertheless, in terms of total emission amount, the most densely populated East China stands out as the largest emitter, accounting for one-third of the total emissions by China’s wastewater sector in 2019 (Supplementary Fig. 8a).

Future wastewater GHG emissions under baseline scenario

In light of the significant impacts of the pollutant removal activities on wastewater GHG emission (Supplementary Fig. 10b) and their close relationship with the selected treatment processes, we anticipated that considerable emission reduction might be realized by adopting appropriate wastewater treatment technologies. For comparison, we first examined how future emissions will change if the currently prevailing treatment processes remain (the baseline scenario). Currently, the wastewater treatment train in WWTPs is dominated by activated sludge processes, but tertiary treatment is increasingly being added for enhanced nutrient removal to meet the stringent water quality limits for WWTP effluent (Supplementary Tables 8 and 9)20. This trend of adding tertiary treatment in WWTPs will continue into the future in China, but the implementation of biotechnology options may vary in different regions.

Specifically, constructed wetlands (CWs) as a low-carbon yet temperature-sensitive wastewater polishing technology may be preferentially deployed in the relatively warm-weather southern regions, while denitrification biofilters (DNBFs) with relatively stable nitrogen removal performance might be predominantly installed in the northern regions (Fig. 4b)21,22. In light of the predominance of indirect emission in wastewater management (Fig. 2b and Supplementary Fig. 10a), the extra energy and chemical consumption incurred in tertiary treatment may critically affect future GHG emission patterns. Generally, DNBFs require the addition of organic substrate to strengthen denitrification (Supplementary Tables 10 and 11), thus its introduction into the activated sludge-based treatment train (Baseline-a scenario) would raise the overall GHG intensity to 0.86 kgCO2e m3 (Fig. 4c). In contrast, introducing CWs as an alternative tertiary treatment (Baseline-b scenario) allows a much lower GHG intensity (0.67 kgCO2e m3), benefiting not only from savings in energy and chemicals in the treatment process, but also from the extra carbon credit achieved by using the waste plant biomass for biogas and liquid fuel production23,24,25 (Fig. 4c). Such regional-specific selection of tertiary treatment modes is expected to further increase the spatial heterogeneity of wastewater-associated GHG emissions in China.

Fig. 4: Projected wastewater GHG emissions under different future wastewater management paradigms.
figure 4

a, Schematic diagram of four envisaged wastewater treatment processes, denoted Baseline, Optimized-I, Optimized-II and Optimized-III. AS, activated sludge process; HRAS, high-rate activated sludge process; HSAD, high-solid anaerobic digestion; AnMBR, anaerobic membrane bioreactor; PN/A, partial nitritation/anaerobic ammonia oxidation; HP, heat pump. The recovered energy (electric power and heat) is used in situ in WWTPs. b, Regional-specific implementation of the different treatment processes in China. cf, Projected wastewater GHG intensities (c,d) and economic net present values (e,f) for the different treatment processes in 2030 (c,e) and 2050 (d,f). Baseline-a and Baseline-b refer to employment of DNBF and CW in tertiary treatment, respectively; Optimized-I refers to simultaneous adoption of HP and DNBF; Optimized-II a,b,c refer to adoption of DNBF, CW and DNBF+HP, respectively; Optimized-III a,b,c also adopt DNBF, CW and DNBF+HP, respectively, but differ from Optimized-II processes in secondary treatment. The data are presented as the mean ± SD (n = 100,000).

Source data

According to China’s future planning for the wastewater sector, by 2030, municipal wastewater will be fully collected and the further tightened standard for WWTP effluent (with the water quality approaching Class IV surface water) will be applied in no less than 75% of the municipal WWTPs26,27. This may motivate extended implementation of DNBF and CW tertiary treatment in the upgraded WWTPs under the baseline scenario. Notably, the contribution of tertiary treatment to indirect GHG emissions may be considerably offset by decarbonized energy. Assuming a stepwise upgrading of the WWTPs of 7.5% plants per year starting from 2019 and a gradual decline in the power grid EF in China (Supplementary Table 2), the entire municipal wastewater sector would achieve a 4.6% reduction in GHG intensity by 2030 and a 23.7% reduction by 2050 (Fig. 5a). Nevertheless, this would merely enable stabilization of the total GHG emissions in the context of continuously increasing wastewater volume (Fig. 5b). Therefore, to further reduce the wastewater GHG emissions, a fundamental transition from the baseline scenario to low-carbon wastewater management will be necessary.

Fig. 5: Future trends in GHG intensity and total GHG emissions in China’s municipal wastewater sector under different wastewater management scenarios.
figure 5

a,b, Future trends in GHG intensity (a) and total GHG emissions (b) in China’s municipal wastewater sector under different wastewater management paradigms. Mixed implementation of the baseline scenario and various optimized processes are assumed for the future paradigms. The data are presented at the 95% confidence level; the solid lines in the error bands represent the mean value.

Source data

Future routes towards carbon-neutral wastewater management

Considering the dominant contribution of WWTP operation to wastewater GHG emissions (Supplementary Fig. 10b), optimization of the wastewater treatment processes should be prioritized. In this study we envisaged three optimized treatment processes, with the optional adoption of different resource-oriented technologies to enhance energy and resource recovery (Fig. 4a), for possible application in China.

The first optimized process is a simple modification of the existing activated sludge process and involves the introduction of the heat pump (HP) to directly recover thermal energy from the secondary effluent (Optimized-I process). This technology can be readily incorporated into the treatment train of existing WWTPs without the need for reconstruction28. So far, wastewater-sourced HPs have been installed in over 500 municipal wastewater utilities worldwide, including several full-scale WWTPs in northern China29. Although this technology still faces challenges, such as device clogging and fouling as well as the current inefficient use of the recovered low-grade heat (Supplementary Note 3)28, we expect that in the near future the conditions might mature sufficiently for its large-scale application in the cold-weather northern regions30. Given a moderate temperature difference of 3–4 °C (between WWTP effluent and the ambient environment) and an average heat recovery rate of 1.18 kWh m3 (refs. 28,31), the introduction of HPs would substantially lower the wastewater GHG intensity to 0.47 kgCO2e m3 (Fig. 5a). Assuming a WWTP upgrading rate of 1.8% plants per year with HP installation starting from 2030, this would enable a 38.7% reduction in GHG intensity by 2050 relative to the 2009 level. Nevertheless, due to the drastically increasing wastewater volume, the entire municipal wastewater sector would still contribute 46.38 MtCO2e emissions by 2050 (Fig. 5b).

Apparently, to realize carbon-neutral wastewater management, more low-carbon treatment processes will have to be applied (Supplementary Note 3). One attractive option in this line is the combination of high-rate activated sludge (HRAS), high-solid anaerobic digestion (HSAD; for enhanced sidestream sludge treatment) and partial nitritation/anaerobic ammonia oxidation (PN/A; with both mainstream and sidestream operation), followed by enhanced water purification and sludge valorization downstream of the treatment train (Optimized-II process). This process can divert a considerable fraction of the chemical energy embodied in wastewater to biogas production and allows more energy-efficient nitrogen removal than conventional activated sludge processes (Supplementary Note 3). However, in spite of its successful demonstration at several full-scale WWTPs in Europe and the United States24,32 (Supplementary Note 3)15,33,34,35, the large-scale implementation of these new technologies in China is still in the very early stages with many barriers remaining. Notably, remarkable technological breakthroughs have been achieved that have substantially improved the biogas production in HSAD (for example, co-digestion of sludge with food wastes) and overcome the limitation associated with the instability of mainstream PN/A operation36. Given a reasonable upgrading rate of 5% plants per year with the implementation of these new technologies starting from 2030 (in place of the Optimized-I process), the overall GHG emissions by the municipal wastewater sector in China would decline to near zero by around 2050, achieving a near 100% reduction relative to the 2009 level.

Notably, the above process still involves intensive N2O emission from the PN/A treatment10,37,38, which commonly occurs under dynamic and suboptimal operating conditions14,15. In addition, there is still space for improving the recovery of energy and chemicals from wastewater2,39. Therefore, we envisaged another treatment process with the implementation of more ground-breaking technologies (Optimized-III process)9. Specifically, the anaerobic membrane bioreactor (AnMBR), as a mainstream treatment technology, is used to produce biogas directly from wastewater and simultaneously yield a high-quality effluent for agricultural irrigation or urban greening (Supplementary Note 3)40. Considering the spatial and seasonable restriction of such water reuse practice, as a supplementary option, nutrient recovery from the anaerobic effluent is also considered. Here, nitrogen recovery may bring extra carbon credit because the recovered ammonia may be used directly as a fertilizer to substitute synthesized ammonia (that is, chemical offset). Several emerging technologies, including membrane contactor, adsorption, struvite precipitation, microalgae treatment and their combinations, are available for such purposes, although their full-scale demonstration in WWTPs remains scarce (Supplementary Note 3)41,42,43. Impressively, with the implementation of these technologies, together with DNBFs, CW treatment or HPs optionally, the overall GHG intensity may be further lowered to −0.17 to −0.24 kgCO2e m3 (Fig. 5a). Assuming a gradual maturing and stepwise implementation of these new technologies in upgraded WWTPs from 2030 at the same rate as in the Optimized-II scenario, the wastewater GHG emission would decline more rapidly and reach net zero as early as 2044 (Fig. 5b). Therefore the Optimized-II and Optimized-III processes are both highly desirable in the upgrading of WWTPs in China towards carbon neutrality.

It should be noted that the above are just three representative treatment processes to be considered in China’s future wastewater management paradigm, and there remain many other options to be considered (Supplementary Note 4). For example, osmotic- or solar-driven treatment processes and microbial electrochemical technologies might be employed to offset part of the energy consumption44,45. In addition, microbial-nanomaterial hybrid systems may hopefully be incorporated into the wastewater treatment line to allow the greener production of high-value products46,47. Such technology innovations may bring tremendous new opportunities for carbon reduction in the municipal wastewater sector. The reduction potential may be even more remarkable if we further expand these technologies to the entire wastewater sector (including industrial, rural, livestock and aquaculture wastewaters), which is predicted to reach ~308.32 MtCO2e emissions by 2030 (Supplementary Fig. 11), or ~2.5% of the national GHG emissions48 (Supplementary Note 4).

Perspectives for promoting low-carbon wastewater management

In spite of the exciting prospects for transition from carbon-intensive to resource-oriented wastewater management in China49, considerable barriers remain. Addressing these challenges requires continued endeavours at technological, economic and policy levels. Firstly, the stable operation of many resource-oriented treatment technologies, especially those in the Optimized-III process, have been demonstrated mainly at pilot or even laboratory scales (Supplementary Note 3). Further technology advances, integration and full-scale demonstration will be essential to enable their commercialized application, and this action has already begun in China. For example, the newly built Yixing concept WWTP, an innovation centre with 1,000 m3 d−1 treatment capacity, is pioneering in demonstrating various leading-edge wastewater treatment technologies and actively promoting their integration (Supplementary Fig. 12). Several other similar WWTPs are also under construction in China. We anticipate that these pioneering plants may lead the technology innovation and upgrading of WWTPs in China, thereby ultimately reshaping the entire municipal wastewater sector.

Notably, the low-carbon technologies are unlikely to fully replace the conventional treatment processes, at least in the near future, due to insufficient technology maturity and restrictions in, for example, land availability, product quality, market demand and initial investment. In particular, putting the new processes into commercialized application requires the corresponding socio-economic transitions. Encouragingly, a cost–benefit analysis shows that the transition to resource-oriented treatment processes may also bring considerable cost reduction in WWTP operation due to the extra economic revenue from carbon credit and recovered resources (for example, heat energy, clean water and fertilizer; Supplementary Tables 12 and 13). Given a moderate carbon trading price of US$50 tonnes−1 in 203050 plus the revenues from other wastewater-derived energy and resources, which also have increasing unit cost over time (Supplementary Note 5), the net economic revenues from wastewater management (reflected by the economic net present value under 2030 conditions) may reach −US$0.05, US$0.03 and US$0.05 m3 for the best-performing processes under the Optimized-I, Optimized-II and Optimized-III scenarios, respectively, compared with a net cost of US$0.11 m3 for the Baseline-a process (Fig. 4e). The discrepancy between different processes is expected to further increase with ever-increasing costs of traded carbon (assuming US$100 tonnes−1 by 2050) and the recovered sources. Assuming a gradual, spatial-diversified implementation of the various new technologies as described above, by 2050, net economic benefits of on average US$0.02, US$0.07 and US$0.11 m−3 for municipal wastewater management in China may be realized under the Optimized-I, Optimized-II and Optimized-III paradigms, respectively. Therefore, the Optimized-II and Optimized-III paradigms have exciting economic prospects and have already attracted increasing social investment in China’s wastewater sector in recent years (Supplementary Fig. 13).

Of course, realization of such a transition would depend strongly on the rate of technology advances and market maturing, which will still take time. Although China launched its carbon market in 2021 and is also planning to initiate markets for wastewater-derived resources51, the trading mechanisms are still incomplete with limited coverage. Nevertheless, economics might not be a major barrier to application of the optimized processes, because their costs are still comparable to the baseline process even without considering the revenues from the recovered resources (Supplementary Note 6). Considering the relatively high initial investment required by the Optimized-II and Optimized-III processes, there is a high chance that they will be implemented first in economically developed regions, such as East China, which has the highest wastewater GHG emissions. Overall, a customized selection of wastewater treatment processes for different regions based on a balanced consideration of technology suitability, economic affordability as well as other environmental and social issues should be encouraged.

Considering the complexity of wastewater management and the tremendous technological and economic challenges ahead, to motivate the paradigm transition the government will need to play a decisive role in coordinating the efforts from different sides52,53. For example, the government may provide funds directly to support the research and development of new technologies and the construction of infrastructure associated with resource-oriented processes. Furthermore, policies and financial means, such as tax incentives and low-interest-rate funding, may be applied to encourage the participation of enterprise in the upgrading and operation of WWTPs. Appropriate policies may also help to incentivize the upgrading of technology and stimulate social investment in this specific field by cultivating the market for wastewater reuse and resource recycling54. In this way, a new type of public–private partnership may be established to synergistically promote the development and application of new technologies. This is already happening in China. Driven by government support and social investment, a number of WWTPs to demonstrate the new technologies are already under construction in China, and the total number is expected to reach over 100 in the coming 5–8 years. Lastly, while current efforts to reduce emissions are mainly driven by top-down policies, measures that can incentivize bottom-up efforts, such as mobilizing the society to participate in wastewater resource recycling activities and supervising wastewater utilities should also be encouraged.

In summary, the goal of realizing carbon neutrality in China will trigger profound top-down reforms in every aspect of society, including the wastewater sector. Here, we have shown that the wastewater sector in China could make valuable contributions to the national carbon reduction targets by revolutionizing treatment processes52,55,56. Importantly, China is currently in the right time and right position to promote such a fundamental transition in the wastewater sector, benefiting from its powerful central administration, rapidly rising research capability and role as global leader in infrastructure construction. In addition, the findings presented here may also have important global implications for controlling wastewater GHG emission and inspire exploration towards carbon-neutral wastewater management in other countries (Supplementary Note 7).

Methods

To obtain the detailed emission activity information of China’s municipal wastewater sector, we established a nationwide inventory covering the operating data of 5,155 municipal WWTPs and the associated sewers throughout the country during the period of 2009–2019. The dataset includes information on geographical distribution, treatment capacity, adopted treatment process, electricity consumption, influent and effluent properties, and sludge yield for each WWTP, as well as information on the wastewater collection ratio, sewer types and sludge disposal approaches in each province (Supplementary Figs. 1417)57. The data on GHG emissions generated in different stages of wastewater management were collected from case studies reported in the literature (Fig. 1a). To improve data validity, the emission data were screened in the following order of decreasing priority: (1) field-measured data from full-scale wastewater facilities in China, (2) field-measured data from full-scale wastewater facilities in other countries, (3) data measured in pilot projects, (4) data measured in laboratory studies and (5) model simulation results. Selected EF data for fugitive GHG emissions at different stages of wastewater management, after brief processing to unify the units, are listed in Supplementary Tables 37.

To estimate the wastewater GHG intensity, we used the activity data for each WWTP and recalibrated the EF values taken from the literature (Supplementary Note 1 and Supplementary Figs. 1417). As fugitive GHG emissions are mainly generated from the decomposition of organic and nitrogen pollutants, the removed COD was adopted as the activity data for estimating fugitive emissions of CH4 and fCO2. Similarly, N2O emissions were estimated by using removed TN as the activity data. For the emissions from sewers, WWTP effluent and untreated wastewater, the total COD or TN was used as the activity data as no data on removed pollutants were available. To estimate the sludge disposal-associated emissions, the sludge yield (that is, the excess sludge produced from treating each cubic metre of wastewater) was taken as the activity data. Details of activity data collection and the procedures for determining the EFs for both fugitive and indirect GHG emissions, which vary for different emission activities and for different wastewater management stages, are given in Supplementary Note 1.

The intensity of fugitive emission for a specific type of GHG was obtained by multiplying the case-specific EF by the corresponding plant-resolved activity data (Supplementary Tables 36). Lastly, the estimated GHG intensities for all types of GHGs were added together to yield the overall wastewater GHG intensity. To facilitate the calculation, the emissions of different GHG types were all converted into CO2 equivalents of emission according to the global warming potential values of 25 CO2e for CH4 and 298 CO2e for N2O over a 100-year global warming potential58,59.

Cost analysis was performed to evaluate the economic competitivity of the wastewater treatment processes under different future paradigms. Details of the calculations are given in Supplementary Note 2.

Reporting summary

Further information on research design is available in the Nature Portfolio Reporting Summary linked to this article.