Land application, composting, incineration, anaerobic digestion, pyrolysis and landfilling are the most commonly used sludge management methods, to recover energy and nutrients from sludge (Liu et al., 2017). All sludge treatment and disposal methods have their advantages and disadvantages (Table 1).

Life cycle assessment (LCA) is a globally recognized analysis method for assessing the environmental impacts of systems and is standardized by the International Standardization Organization (ISO). LCA is typically conducted through four phases: goal and scope definition, inventory analysis, impact assessment, and interpretation (Godlee, 2006).

The goal of LCA indicates the intended application and the reason for conducting the study, as well as the audience and comparison assertions. The scope definition is sufficient to ensure the breadth, depth and detail of the research to meet the stated goals. The functional unit (FU), as a measure of the output function of the product system, is to supply a reference benchmark for related inputs and outputs. FU is a crucial basis for the simultaneous comparison and analysis of optional scenarios (Hertwich et al., 2001; Teoh and Li, 2020). In previous LCA studies of sewage sludge treatment units, the specified amounts of sewage or sludge mass is a common FU (Hospido et al., 2004; Yoshida et al., 2013; Ding et al., 2021).

System boundary defines the boundary of the basic flow unit process of inputs and outputs, which have a significant impact on LCA (Finnveden et al., 2009; Pradel et al., 2016). Life cycle inventory analysis covers energy and material requirements, pollutant emissions and environmental hazards resulting from raw materials mining, refining, product manufacturing, transportation, sales, consumption and disposal, and is referred to as foreground data, followed by raw material data form processing technologies (Pradel et al., 2016). Data for treatment techniques are commonly sourced from literature, field test data and LCI database, such as Econinvent (Yoshida et al., 2013; Ding et al., 2021). The construction and decommissioning of treatment plants are generally excluded, previous studies have shown their contribution to the impact to be negligible (Fransson et al., 2008; Yoshida et al., 2013). In general, LCI plays an essential role in LCA analysis, and LCI data can affect the quality and uncertainty of LCA results.

Sludge is a by-product of sewage treatment in wastewater treatment plants (WWTPs) that receive wastewater from several municipalities, including domestic, industrial, commercial, hospital, road and agricultural water, human excreta, personal care products, detergents, disinfectants, pharmaceuticals, sediments, etc (Cantinho et al., 2016). Most conventional WWTPs normally have the following steps: preliminary treatment, primary treatment and secondary treatment, the layout of which is shown in Figure 1.

In preliminary processing, large objects in raw wastewater are filtered through a grating screen. Due to the difference in gravity, sand and grit settle in the gravel channel for treatment. In the primary treatment, grease and oil rise to the surface and are skimmed off in the primary clarifier/sedimentation tank, while sludge (in the form of semi-solids/slurry) settles at the bottom of the tank. The primary stage removes 50%–70% of the total suspended solids from the raw wastewater and the sludge produced in this stage is referred to as primary sludge (Pathak et al., 2009).

Secondary treatment is a biological process that uses microorganisms to stabilize organic constituents and remove non-colloidal solids. Activated sludge processes and settling tanks are commonly used for secondary treatment, producing secondary sludge. Additional treatment (tertiary treatment for nutrient removal) can be added to improve the sewage quality (Syed-Hassan et al., 2017). Sludge from WWTPs is generally treated in a sequence of thickening, stabilization/conditioning and dewatering to reduce the total weight and volume for transportation and additional disposal (Yoshida et al., 2013; Zhang et al., 2014).

Proximate analysis includes moisture content (M), volatile content (V), fixed carbon content (FC) and ash content (A), and the moisture content of sewage sludge is extremely variable and shows large fluctuation. After thickening, the sludge mass is generally reduced by 70% and the moisture content of sludge is decreased to 97%–98% (Stefanakis et al., 2014). The moisture is significantly reduced to about 80% by dewatering (belt filter press or centrifuge) (Kumar and Kumar, 2018). Figure 2 showed the results of the dry basis proximate analysis (d in form of subscripts) obtained from previous literature (Domínguez et al., 2008; Zhai et al., 2012; Xiong et al., 2013; Zhang L. et al., 2015; Zhang Y. et al., 2015; Kan et al., 2016; Lishan et al., 2018). Sludge possessed a relatively stable A_d content of approximately 31.2–57.2 wt%, which is domainated by inorganic components left over from the wastewater treatment process. The volatile content of the sludge is extremely variable, i.e. 38.6–60 wt%. The fixed carbon content in the sludge is roughly below 11.2%. The proximate analysis of sludge varies widely compared to alternative fuels (biomass or coal), and sludge is considered a solid fuel with high ash and volatility (Zhu et al., 2017; Usman et al., 2020).

The ultimate analysis determines the content of C, H, O, N and S, Figure 3 shows the ultimate analysis data for the dried ash-free base of sludge reported in the literature (Domínguez et al., 2008; Zhai et al., 2012; Xiong et al., 2013; Zhang L. et al., 2015; Zhang Y. et al., 2015; Kan et al., 2016; Lishan et al., 2018). The C contents in sludge is dominant with about 40.6–57.4 wt%. The content of H in sludge is 3.7–10.6 wt% (average content of 7.4 wt%). The total content of O, N and S in sludge ranged from 34.6 to 54 wt%, with average levels of 31.3 wt%, 7.1 wt% and 2.6 wt% for O, N and S, respectively.

Figure 4 shows several typical heavy metal concentrations for different sewage sludge. The hollow symbols indicate the heavy content of different sludge from previous literature (Dai et al., 2007). And the solid symbols and error bars represent the average and maximum, minimum contents of heavy metals (Cheng et al., 2022). These data were obtained from the database of sewage sludge data containing heavy metals concentrations established by analyzing 104 data sets selected from 420 references on the China National Knowledge Infrastructure (CNKI). The heavy metal thresholds of sludge for agricultural use (Leval A) in “Control standards of pollutants in sludge for agricultural use (GB 4282–2018)" are listed as dotted line for reference. The avarage values of the eight typical heavy metals in the #1∼#6 sludge and the average content in the above sludge database are below the threshold, while the contents of As, Ni, Zn and Cu have the potential to exceed the threshold based on the previous statistical data.

Previous literature indicated that about 50%–80% of heavy metals (i.e., Cd, Hg, Pb, Cr, As, Ni, Zn and Cu) in wastewater enter suspended solids through physicochemical and biological interactions (passive cellular adsorption, i.e., complex and inorganic microprecipitation reactions, and active cellular uptake, i.e., bioaccumulation of heavy metals by biochemical reactions with intracellular compounds) (Pagnanelli et al., 2009; Choińska-Pulit et al., 2018; Huang et al., 2018; Qin et al., 2019; Wang et al., 2020; Yang et al., 2020). In particular, suspended sediments containing large amounts of organic matter, such as proteins, polysaccharides and humic substances, readily interact with heavy metals (Chojnacka, 2010; Qin et al., 2019). Cr, Cd and Cu originate mainly from pipeline sediments, Zn and As are mainly from agriculture and domestic sources, Ni is mainly industry and Pb originates from fuel combustion and traffic (Cheng et al., 2022).

Composting is an aerobic process in which most of the degradable organic carbon in waste is converted into CO₂. CH₄ is formed in the anaerobic processes, and is mostly oxidized in aerobic. CH₄ released into the atmosphere accounts for between one and several percent of the initial carbon content of the sludge. N₂O is also emitted during composting, ranging from 0.5% to 5% of the initial nitrogen content. Operating electricity for auxiliary machinery such as mixers and conveyors is the main energy consumption in the composting process. The electricity consumption of the compost mixing process was 33.2 kW h/t DS and that of the fermentation and maturation process was 501 kW h/t DS (Tarpani et al., 2020).

The environmental impact of the agricultural application of composted sludge was investigated and the net GHG emissions were 197 kg CO₂-eq/t DS. The main contribution of different life cycle stages to GHG emissions is sludge treatment (according to 75.25% of the total emissions), that is, CH₄ emissions from sludge composting and CO₂ emissions corresponding to electricity consumption. The credits for avoiding CO₂ emissions accounted for −19.54%, mainly due to compost being used as a fertilizer substitute. Piippo et al. (Piippo et al., 2017) assessed the GHG emissions from composting in the form of “with CO₂” and “without CO_2,” which depends on whether the biological CO₂ is considered. The net GHG emission from composting is 2047.1 kg CO₂-eq/t DS, or 1113.8 kg CO₂-eq/t DS excluding biological CO₂. Biological CO₂ is mainly expelled during composting, and CH₄ and N₂O are mainly produced during the composting process. The avoided CO₂ emission is 288 kg CO₂-eq/t DS. By using compost as fertilizer, it can replace some chemical fertilizers and reduce the emissions in their production. Lishan et al. (Lishan et al., 2018) assessed the CO₂-eq emissions of composting. The amount of CO₂-eq produced in the composting process is 1097.9 kgCO2-eq/t DS, and the disposal process is a major contributor to CO₂ emissions compared to transport and dewatering.

The GHG emission during composting is about 180–1100 kg CO₂-eq/t DS (Hong et al., 2009; Sablayrolles et al., 2010; Lombardi et al., 2017), which is mainly affected by different profiles of grid electricity and alternative schemes of compost.

Anaerobic digestion, a low-energy biochemical process, is recognized as a cost-effective method for solid waste disposal and resource recovery. During anaerobic digestion of sludge, biodegradable waste is decomposed into biogas, which is a mixture of CO₂, CH₄ and a slight amount of addition gases (Yan et al., 2021).

Anaerobic digestion process, belt filter dewatering after anaerobic digestion, storage and agricultural application consume 88.6 kW h, 49.1 kW h and 58.5 kW h electricity respectively (Tarpani et al., 2020). Digestion can replace 50–100 kg of equivalent synthetic NPK 15-15-15 fertilizer, depending on the phosphorus content. Biogas is used to maintain the digester at 35°C (mesophilic conditions), and the excess is used for electricity generation of 397–794 kW h. Li and Feng (Li and Feng, 2018) compared the integrated approach with anaerobic digestion and pyrolysis in terms of life cycle environmental impact and energy efficiency. The energy consumption of the anaerobic digestion process is 2.8 GJ/t DS of heat and 50 kW h/t DS of electricity. The energy consumption of thermal drying is 2940 MJ/t H₂O of heat and 40 kW h/t H₂O of electricity.

Tarpani et al. (Tarpani et al., 2020) assessed the GHG emissions from anaerobic digestion, and the results showed that the main contributors were CH₄ emissions from agricultural application (80 kg CO₂-eq/t DS) and fossil CO₂ from emissions from grid electricity generation (59 kg CO₂-eq/t DS). The credits GHG emissions mainly include digestion as substitute fertilizer (−127 kg CO₂-eq/t DS) and biogas for electricity generation (−239 kg CO₂-eq/t DS). Piippo et al. (Piippo et al., 2017) assessed the GHG emissions from anaerobic digestion. The net GHG emission is 141.15 kg CO₂-eq/t DS, which is −341.3 kg CO₂-eq/t DS without accounting for biological CO₂. Biolohical CO₂ is mainly expelled during the utilization process of biogas in the CHP process and abiotic CO₂ is mainly produced during the utilization process of digestion. The avoided CO₂ emission is −565.7 kg CO₂-eq/t DS by replacing some chemical fertilizers with compost as fertilizer and reducing CO₂ emissions. The GHG emission of the anaerobic digestion and agricultural application route is −280–650 kg CO₂eq/t DS (Murray et al., 2008; Peters and Rowley, 2009; Gourdet et al., 2016). Variation is mainly due to different sludge treatment conditions and the amount of electricity generated and fertilizer avoided.

Pyrolysis of dried sludge produces syngas, tar and/or biochar, and the drying process and pyrolysis require energy consumption. The syngas generated is used to aid the pyrolytic process and the excess can be exported. The production of tar and char is also credited with producing comparable amounts of fossil fuels such as heavy fuel oil and coal.

The filter press process, the thermal drying process and the pyrolysis process of the pyrolysis route consume 40, 118 and 244 kW h of electricity, respectively. During the thermal drying process, 1638 kW h of heat is supplied from a natural gas industrial furnace. In the pyrolysis process, 115–230 kg of biochar and 20–40 kg of bio-oil can be recovered, and 2.1–4.2 kW h of heat can also be recovered by pyrolysis gas (Tarpani et al., 2020). In the pyrolysis pathway, electricity of 77.78 kW h/t DS is used for the operation of reactors, conveyors and accessory devices and natural gas is used as the complementary energy of 800–1,320 MJ/t DS for dewatered sludge (60%wt moisture content) pyrolysis, which is depended on the organic content and pyrolysis condition, e.g., 1,180 MJ/t DS at 550°C (Hossain et al., 2010), 1,147 MJ/t DS at 500°C (Agarwal et al., 2015) and 950 MJ/t DS at 450°C (Tomasi Morgano et al., 2018).

The GHG emission from the pyrolysis route is 315 kg CO₂-eq/t DS, of which 422 kg CO₂-eq/t DS can be produced from the gas consumed in the thermal drying process. The utilization of pyrolysis gas, biochar and bio-oil can reduce GHG emissions (−251 kg CO₂-eq/t DS) by the avoidance of heat and fuels. Barry et al. (Barry et al., 2019) assessed the GHG emissions in different application scenarios of biochar from sludge pyrolysis. The GHG emissions from incineration without energy recovery and incineration with organic Rankine cycle energy recovery ash were 646.98 and 238.28 kg CO2-eq/t DS, respectively. Agricultural applications and fuel utilization of biochar showed reduced global warming potential, with credit GHG emissions of −87.14 and −495.46 kg CO2-eq/t DS respectively. Agricultural land applications of biochar show a reduced potential for global warming. This is accomplished by recycling the energy from the pyrolysis process, reducing the need for fossil fuels, displacing mineral fertilizers, and carbon sequestration in the biochar. The use of biochar in cement kilns shows a significant reduction in global warming potential. This is partly due to the re-use of energy from the pyrolysis process, but primarily due to the replacement of lignite coal as a fuel. The GHG emission is about −495–500 kg CO₂-eq/t DS (Hospido et al., 2005; Houdkova et al., 2008). This is achieved by recovering energy during pyrolysis, reducing the need for fossil fuels, replacing mineral fertilizers and sequestering carbon in biochar.

In the sludge incineration process, the organic compounds of sludge are oxidated at elevated temperatures, and biomass energy is converted into heat and can be further converted into electricity, biosolids are burned to form mainly biological CO₂ and water, leaving only inert material (ash). The calorific value of dried sewage sludge is the same as that of lignite, and it can be used for sludge drying, heat generation and power generation through energy recovery.

In the sludge incineration route, 52.5 kW h of electricity is consumed during the centrifuge dewatering of the sludge to 75% moisture content, and 31 kg/t DS diesel heavy oil is burned in the furnace to be used as an auxiliary fuel. 227–454 kW h of electricity and 12–24 kW h of heat can be recovered by sludge incineration (Tarpani et al., 2020).

The dewatered sludge is incinerated in fluidized bed boiler at 850°C, with the addition of fuel and lime to improve the combustion efficiency and to control acid gases, respectively. The main contributors of GHG emissions (39.8 kg CO₂-eq/t DS) from sludge incineration are the CO₂ emissions of heavy oil (108 kg CO₂-eq/t DS) and credit emission from energy recovery (94% electricity and 6% heat) reduce the emissions (−147 kg CO₂-eq/t DS) (Tarpani et al., 2020).

The GHG emission from thermal drying-incineration is 103 kg CO2-eq/t DS. Most of the heat generated by dried sludge incineration is used for thermal drying, and the main contribution of CO₂ emission is transportation, as the moisture content of sludge after the belt filter press is 81.8%, the sludge reduction is not sufficient (Zhang et al., 2019). Compared to thermal, electro-dewatering consumes 223 kg CO2-eq/t DS, with a credit emission of −312 kg CO₂-eq by displacing the fossil fuel. Therefore, the GHG emission from electro-dewatering-incineration is −32 kg CO2-eq/t DS.

Piippo et al. (Piippo et al., 2017) assessed the GHG emissions from incineration with mechanical drying and with thermal drying. The net GHG emission from incineration with mechanical drying is 1002.7 kg CO₂-eq/t DS, and that from incineration with thermal drying is 924.01 kg CO₂-eq/t DS. Regardless of biogenic CO₂, the net GHG emissions from incineration with mechanical drying and thermal drying are −617.7 kg CO₂-eq/t DS and −874.7 kg CO₂-eq/t DS, respectively. Biological CO₂ is mainly expelled during the combustion of sludge. Electricity consumption in the drying process and supplement fuel consumption are major contributors to abiotic CO₂ GHG emissions from incineration with mechanical drying, while electricity and steam consumption are major contributors to abiotic CO₂ GHG emissions from thermal drying. The avoided CO₂ emissions from mechanical drying and thermal drying are −942.5 kg CO₂-eq/t DS and −1,047 kg CO₂-eq/t DS respectively. CO₂ emissions from fossil fuels can be reduced by using dry sludge as fuel in power plants.

The GHG emission of sludge incineration is −617–670 kg CO₂-eq/t DS (Houdkova et al., 2008; Hong et al., 2009; Lombardi et al., 2017), which is mainly affected by the sludge drying method and/or the addition amount of auxiliary fuel and energy recovery.