Thermal treatment options for biosolids management: a critical review

Abstract

Thermal treatment of biosolids is receiving significant attention in the water industry as an alternative management option to land application. Traditional thermal treatment processes for biosolids management include drying and incineration, whereas emerging thermal technologies comprise dry thermal processes, such as pyrolysis and gasification, and wet thermal processes, such as hydrothermal carbonisation/liquefaction and supercritical water gasification. Thermal treatment is considered an efficient approach for the volume reduction of biosolids, contaminant destruction, and valuable product generation. However, there is a clear gap in the literature in benchmarking the range of available technologies, considering their techno-economic viability, emission potential, resource (energy and nutrient) recovery, and contaminant reduction. This knowledge is crucial for understanding the techno-commercial readiness, integration flexibility, and potential adoption of the thermal treatment technologies for biosolids management in wastewater treatment facilities. This critical review provides a comprehensive comparison of the various thermal treatment processes based on the parameters such as fate of nutrients and emerging contaminants, emissions, energy requirement, capital and operating expenditures, and scale-up maturity. It was found that dry thermal processes have substantial benefits over traditional incineration technologies, with pyrolysis and gasification being more energy-efficient and providing opportunities to generate valuable products (biochar and bioenergy). Hydrothermal liquefaction offers further benefits with high bio-oil and nutrient recovery and strong synergies with the existing water treatment infrastructures. Gasification and pyrolysis have high technology- and commercial-readiness level for biosolids treatment, making them suitable for the wastewater treatment industry. However, to ensure efficient and sustainable management of biosolids through thermal processes, there are some techno-commercial challenges, which are highlighted as future research perspectives.

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Savankumar Patel

Dr Savankumar Patel received his PhD in Chemical Engineering from RMIT University in 2020 and currently works as a Research Fellow at RMIT. With nine years of academic experience, he specializes in product development and process engineering, focusing on optimizing and improving thermochemical conversion processes. He holds two patents and has co-authored over 35 publications. His expertise spans in fluidised bed reactor design, pyrolysis–gasification systems, techno-economic analysis, and commercialisation pathways for clean energy innovation.

Ibrahim Gbolahan Hakeem

Dr Ibrahim Hakeem is a Research Translation Fellow in the School of Engineering at RMIT University, specialising in sustainable resource recovery, advanced thermal conversion, and circular economy technologies. He completed his PhD in Chemical Engineering at RMIT in 2023, focusing on heavy metal recovery from biosolids prior to pyrolysis conversion to biochar. His research develops innovative processes to transform waste plastics and biosolids into high-value carbon materials and clean energy while safely managing contaminants. Dr Hakeem has contributed to multiple industry-collaborative projects with major Australian water utilities and has published over 40 journal papers.

Aravind Surapaneni

Dr Aravind Surapaneni is a Principal Scientist at South East Water, Australia, with over two decades of experience in wastewater management, biosolids treatment, and resource recovery. He manages the Intelligent Water Networks Biosolids and Resource Recovery Program and serves as the Deputy Director of the ARC Training Centre for the Transformation of Australia's Biosolids Resource at RMIT University. Dr Surapaneni leads industry–research collaborations, advancing circular economy solutions through thermal, chemical, and biological processes. He has authored numerous peer-reviewed publications and technical reports that have shaped best practices and policies in sustainable water and waste management.

Damien J. Batstone

Damien Batstone is a Professor and Director of the Australian Centre of Water and Environmental Biotechnology at The University of Queensland and has over 20 years of experience in the design and technology development for biosolids management, including extensive bioprocess modelling and integrated process modelling. His recent work focuses on alternative valorisation pathways for biosolids reuse via the integration of biological, chemical, and thermal treatment technologies.

Kalpit Shah

Kalpit Shah is a Professor and Deputy Head of Chemical and Environmental Engineering Department at RMIT University. He is also the Deputy Director of the ARC Training Centre for the Transformation of Australia's Biosolids Resource at RMIT. He is internationally recognised for his expertise in thermochemical conversion, waste valorisation, and process intensification for clean energy and circular economy applications. Professor Shah leads multidisciplinary research, developing sustainable technologies for converting waste biomass, plastics, and biosolids into high-value fuels and carbon materials. With over 200 peer-reviewed publications and strong industry partnerships, he has significantly contributed to advancing low-emission technologies and promoting resource-efficient solutions.

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Water impact

Thermal treatment offers a sustainable alternative for biosolids (stabilised sewage sludge) management, enabling volume reduction, contaminant destruction, and resource recovery. This critical review benchmarks key thermal technologies across different technical and commercial indicators. The findings highlight pyrolysis, gasification, and hydrothermal processes as promising solutions for wastewater treatment facilities, seeking energy-efficient, low-emission, and commercially viable strategies for biosolids valorisation and circular resource integration.

1. Introduction

Biosolids are stabilised sewage sludge produced in wastewater treatment plants (WWTPs) from wastewater treatment. The growth in global population and urbanisation has led to the rapid expansion of WWTPs, leading to increased production of biosolids. Globally, biosolids production is around 100 million dry tonnes, which is projected to reach 175 million dry tonnes by 2050,¹ with a daily per capita generation of 35–85 g dry matter.^2,3 In Australia, 372 000 dry tonnes of biosolids were generated in 2023 and New Zealand produced 75 000 tonnes of dry biosolids in 2023.⁴ In 2024, over 2300 WWTPs in the US generated around 4 million dry tonnes of biosolids,⁵ while 15 million tonnes of dry biosolids were generated in EU-28 in 2021.⁶ In a typical WWTP, the management of biosolids accounts for more than 50% of the operating cost and about 40% of greenhouse gas emissions.⁷ The increasing production volume and stricter regulations pose challenges for the sustainable management of biosolids.

Biosolids consist of organic matter (polysaccharides, proteins, lipids, and non-degradable organics such as humic substances, lignin, and synthetic organics), inorganic materials (such as clay, sediments and minerals), macronutrients (such as potassium, calcium, magnesium, nitrogen, phosphorus, and sulfur) and micronutrients (such as zinc, iron, copper, boron, manganese and molybdenum).⁸ The composition of biosolids, particularly their macro- and micro-nutrient constituents, makes them suitable materials for land application in agricultural soils. Land application is one of the major traditional routes for biosolids management. For example, the US land applied about 60% of its biosolids produced in 2024.⁴ In Australia, 80% of biosolids were used for land application in agricultural soils in 2023.⁹ However, growing concerns regarding the presence of contaminants such as microplastics, per- and poly-fluoroalkyl substances (PFAS), pharmaceuticals, and pesticides, in biosolids and their associated impacts on soil and groundwater may restrain their direct land application in the future.¹⁰ In addition, heavy metals and microbial pathogens are present in biosolids, often at levels that raise concerns. The uptake of these contaminants by crops may result in their entry into the food chain and pose severe risks to the human health if not managed effectively. While pathogens can be effectively controlled by biological treatment or eliminated (by moderate thermal treatment), heavy metals are difficult to remove by existing treatment approaches and should be source controlled. Emerging contaminants such as microplastics and PFASs are potentially subject to regulations as they present a risk to the agricultural use of biosolids.¹¹

Other conventional disposal methods of biosolids, such as landfilling and stockpiling, are not considered environmentally sustainable. Globally, about 30% of biosolids produced yearly are stockpiled or landfilled. In Australia, the US, and China, about 15%, 28%, and 44% of biosolids are either stockpiled or landfilled, respectively.¹² While stockpiling and landfilling have been conventional biosolids management options, they offer little environmental benefits and face several issues including health risks, greenhouse gas emissions, groundwater contamination, and increasingly stringent legislations and negative public perceptions.¹³ Composting is another popular alternative option for biosolids management. However, it is a labour-intensive and costly approach, which effectively increases the biosolids volume, dilutes rather than eliminates contaminants, and requires a market for the composted biosolids.¹⁴ Therefore, the increasing biosolids production and stringent environmental regulations are pushing the wastewater industries to look for alternatives that could massively reduce biosolids volume, destruct/eliminate contaminants, and recover resources.

Thermal treatment of biosolids, having undergone numerous technological iterations in the last 100 years, has emerged as an attractive approach focusing on emission management and energy efficiency. Established technologies include thermal drying, thermal hydrolysis, and incineration. Emerging technologies include dry thermal processes (torrefaction, pyrolysis and gasification), and wet thermal processes (hydrothermal carbonisation, hydrothermal liquefaction, and hydrothermal gasification). The traditional thermal treatment methods, such as drying and incineration, can destroy organic contaminants and pathogens and facilitate energy recovery, but have a few shortcomings.¹⁵ For instance, incineration produces CO₂, NOx, dioxins, and particulate matter, which must be controlled. The emerging thermal techniques have substantial benefits over conventional ones in generating valuable products, such as char and bio-oil and non-condensable gas for usable heat and power generation.¹⁶

Recently, there has been significant interest by water utilities in adopting thermal technologies for biosolids management. Plausible benefits include reducing biosolids volume by up to 70%, greatly reducing or eliminating odour, pathogens and persistent organic contaminants, and recovery of valuable resources (carbon, energy, and nutrients).¹⁵ However, a lack of understanding about the environmental performance and suitability of the diverse thermal treatment technologies in recovering valuable products limits their adoption at a commercial scale in the water industries. Therefore, benchmarking thermal and hydrothermal technologies is essential to evaluating their suitability for water industries, which is missing in the current literature.

Numerous reviews are available in the literature and have focused on the operating principles of different thermal/hydrothermal technologies and effects of process conditions on product distribution and properties.^17–22 For example, Gao et al. and Sharma et al. reviewed the effects of critical process conditions on product yields during thermochemical treatment of biosolids and evaluated the energy recovery potential of the various processes.^3,18 Patel et al. reviewed the relationships of pyrolysis conditions with char, oil and syngas product fractions and overviewed the numerous environmental and energy applications of the biochar product.²³ Zhang et al. reported heavy metals and nutrient migration behaviour during the hydrothermal treatment of biosolids as a function of process conditions.²² The integration of thermal drying process with pyrolysis and gasification for energy-efficient biosolids conversion and contaminant reduction was reviewed by Nylen and Sheeban.²⁴ In recent studies, Marzbali et al. demonstrated a range of emerging industrial applications of char product from the thermal treatment of biosolids.²⁵ Vuppaladadiyam et al. provided a review of application of biosolids-derived biochar and hydrochar for H₂S removal from biogas with a demonstration of how WWTPs can benefit from this circular model of utilising biosolids char for the upgrading of biogas generated from the anaerobic digestion of their sewage sludge.²⁶

In all the existing reviews, there has been limited focus on the technological challenges, energy and nutrient recovery potential, emission characteristics, and techno-commercial feasibility of the different thermal treatment processes ranging from the traditional thermal process (drying and incineration) to emerging thermal process (pyrolysis, gasification and hydrothermal). The understanding of these various aspects of the diverse thermal treatment technologies is essential for fair comparison and benchmarking of the available technologies. The knowledge will guide the WWTPs on the identification and adoption of the most efficient, cost-effective, and commercially matured thermal technologies for their biosolids treatment.

Therefore, the primary aim of this review is to critically discuss the technological level assessment of different thermal and hydrothermal processes, considering the fate of nutrients and contaminants, emission characteristics, range of generated product streams, scale-up plants and technology maturity level, as well as costs and energy requirements. As such, we compare the conventional and emerging thermal technologies by considering their energy requirements, transformation and fate of nutrients and contaminants during treatment, and economic analysis of each process. Furthermore, this study benchmarked the thermal treatment technology and assessed its integration prospects with the existing sludge treatment process, particularly anaerobic digestion. Lastly, the opportunities and barriers to implementing the thermal treatment technologies in the water industry were highlighted.

2. Thermal technologies for biosolids transformation

Thermal technologies for biosolids transformation can be grouped into two main categories: (i) wet processes and (ii) dry processes. Each of the processes can further be classified as oxidative and non-oxidative processes depending on whether oxygen or air is added for the process operation, as summarised in Fig. 1, where the added oxygen concentration is at or above the stoichiometric oxygen requirement for the complete conversion of the organic fuels to CO₂ and H₂O, and such process is referred to as combustion. In this case, incineration is an example of an oxidative process where the added oxygen is at or above the stoichiometric concentration. However, where the added oxygen concentration is below the stoichiometric requirement for the complete combustion of the hydrocarbon fuel, such oxidative process is considered partial/incomplete combustion. In this case, gasification process is a typical example. However, non-oxidative processes operate in an inert environment with no external oxygen or air added, thus preventing organic matter oxidation (as in pyrolysis). Therefore, incineration is a dry and oxidative thermal process, while hydrothermal oxidation and supercritical wet air oxidation are wet oxidative thermal processes. Similarly, pyrolysis is a dry and non-oxidative thermal process, while hydrothermal carbonisation/liquefaction is a wet and non-oxidative thermal process.

Fig. 1 Spectrum of thermal treatment technologies for biosolids transformation.

Thermal processes for biosolids transformation operate under a spectrum of temperature and oxygen concentration with different energy requirements, conversion efficiencies, and product streams. In addition, each of the thermal processes highlighted in Fig. 1 has different resource (energy and nutrient) recovery potentials, emission profile, product distributions, contaminant destruction efficiency, and product quality. Furthermore, some of the thermal treatment techniques such as drying, thermal hydrolysis, and incineration are well established and already adopted by the water industry for treating sludge and biosolids. While other treatment techniques such as pyrolysis, gasification, and hydrothermal processes are yet to be widely adopted and are considered rapidly emerging treatment technologies. In this section, we discuss a range of thermal treatment technologies for biosolids transformation across different metrics, such as energy requirements, fate of nutrients, fate of organic and metal contaminants, capital and operating costs and scale-up plants.

2.1 Thermal drying

Drying is employed as a pre-treatment step prior to incineration or other thermal treatment processes, as the low feedstock moisture content needed for these processes cannot be achieved by mechanical dewatering means alone. Drying helps to reduce volume and serves as a stabilisation method for biosolids use in agricultural soils. For instance, thermally stabilised (dried) biosolids in the US received substantial brand recognition and the final product has a higher sale price than the raw nutrient value (e.g., milorganite, bloom, and “bay state fertilizer”).

The thermal drying of biosolids usually begins with a constant drying phase, during which unbound moisture or free water is removed from the surface. This is followed by two falling rate phases, where bound moisture is gradually eliminated.⁸ Biosolids are dried using different types of dryer technologies, which can be operated in direct or indirect mode. Direct dryers operate by direct contact of biosolids with a hot stream (generally air, steam, or combustion flue gas). Indirect dryers operate by heat transfer from an internal or an external heating element (normally by steam, hot air, or hot oil).^17,27 A comparison of the different types of dryers is shown in Table 1. Direct dryers are generally simpler and may have a smaller footprint due to higher operating temperatures and heat transfer rates. However, they generate contaminated off-gases (even when the contact gas is recycled). Indirect dryers are more complex and larger but allow better heat usage efficiency, as heat is not lost with the exhaust air.²⁸ Since most of the water vapour generated is condensed, it enhances energy efficiency. Both dryer types can be configured to produce a pelletised product, which is formed through agglomeration of particles (particle–particle contact within the dryer). Indirect dryers have better control over pellet size and can produce large pellets. Both approaches allow for the classification of dried biosolids products allowing for the recycling of undersized particle sizes and fines.²⁹ Pellets are generally preferred when the agricultural use of the dried product is desired or when the dried sludge is to be incinerated. Alternative non-pelletising drying processes such as flash, fluidised bed (and spray) are used with off-gases fed with entrained sludge to the incinerator.²⁹ While both types of dryers can be designed to operate without solids recycling, direct dryers are generally dependent on recycling undersize particles due to classification issues, while indirect dryers benefit from recycling to enhance heat transfer.³⁰ Recycling avoids the “sticky zone” at <55–70% solids and allows for the recycling of undersized particles.^30,31

Table 1 Comparison of direct and indirect dryers³¹

Direct dryer	Indirect dryer
• Types: rotary drum, belt, flash, fluidized bed, spray, toroidal	• Types: tray (including multistage and rotating tray), paddle, disc, hollow flight
• Faster heat transfer (smaller dryer)	• More energy efficient (no convective losses, may allow condensate energy recovery)
• Higher thermal energy consumption (3.5–4 MJ kg^−1)	• Higher electrical energy consumption (50 Wh kg^−1)
• Contact gases must be treated (may be reused depending on configuration)	• Reduced off-gases (mainly evaporate, CO2etc.)
• Simpler, available at smaller scale	• Only at larger scale (>10 tpd)
• More market penetration (in USA)	• Mature tech, well established in Europe
• Generally requires product recycle	• Products recycle reduced
• Lower CAPEX	• Lower OPEX
• Can operate at higher contact temperature	• Higher temperature reduces energy efficiency, dependent on HX
• Generates dust in off-gas	• Generates dustier product
• Process explosive risk due to off-gas/dust	• Product explosive risk due to dust
• Smaller pellets (less controllable, configured by design)	• Larger pellets (more controllable)

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2.1.1 Energy requirements. Drying of biosolids is an energy-intensive process, which theoretically requires around 0.63–0.69 kWh_th (thermal energy) per kilogram water evaporated, depending on the initial and final water contents of the biosolids, and whether evaporated latent heat is recovered.^32–34 However, practically, dryers consume about 1.8–2.8 kWh_th per kilogram of dry biosolids, depending on the initial and final water contents of the biosolids. The key factor that determines the overall energy consumption is the state of the input of dry solids (Fig. 2). For instance, a wet cake feed of 10% dry solids requires evaporation of almost 8.95 tonne water, while a wet cake feed of 20% dry solids requires evaporation of 3.95 tonne water (at 3 GJ per tonne) to achieve 95% dry solids. This emphasises the critical role of mechanical dewatering prior to drying.

Fig. 2 Thermal energy required by the dryer to dry biosolids cake to 95% solids using 3 GJ per tonne water. Energy costs $10 per GJ.

The energy requirements of biosolids drying technologies vary with the dryer type.²⁷ For instance, the thermal energy requirement for a drum-type dryer is estimated between 0.85 and 1.07 kWh_th kg⁻¹ water, and for fluidised bed-type and disc-type dryers. it is 0.8–1.1 kWh_th kg⁻¹ water. The belt and paddle dryer requires 0.95 kWh_th kg⁻¹ water while the flash and rotary dryer consumes around 1.05 kWh_th kg⁻¹ water removed. The solar greenhouse drying technology exhibits a significantly lower energy consumption than that of conventional thermal drying. For instance, the energy consumption of solar greenhouse drying to dry flocculated and unflocculated sludge is estimated as 0.22 and 0.28 kWh_th kg⁻¹ water, respectively, which can be further reduced using open airflow greenhouses than the climate-controlled ones.^32,35 It is worth mentioning that Germany and Poland treat around 10 000 and 9000 tonnes of wet biosolids per year, respectively, using a solar greenhouse drying technology.³² Although solar dryers do not require any additional thermal energy, they need prolonged exposure time and rely on weather conditions.^17,18

Arjona and Cisneros performed a feasibility assessment of commercially available dryers in different scenarios in a typical WWTP.³⁴ The authors observed that the net energy from drying anaerobically digested sludge was negative, suggesting that conventional drying is not feasible to produce dried digested sludge for energy applications. However, drying undigested sludge results in a net positive energy of about 1185.98 kWh_th per tonne dry sludge. Đurđević et al. also studied the drying energy balance of sludge, which is the difference between combustion energy recovery and energy required for the drying process.³³ Considering the specific energy consumption as 0.9 kWh_th kg⁻¹ water, energy recovery efficiency as 0.8, and initial sludge solid content as 30%, the thermal energy balance (Q_EB) in kWh_th kg⁻¹ DM can be expressed using eqn (1):

where ODM and DM represent the organic dry matter and dry matter (solid content), respectively, in the sludge after the drying process.

Đurđević et al. reported a negative energy balance for sludge with ODM < 0.44, irrespective of the value of the DM, indicating the case of drying digested sludge.³³ On the contrary, the energy balance was positive for ODM > 0.66 over the entire range of DM tested, indicating the case of drying non-digested sludge. The positive thermal energy balance in the case of drying undigested sludge is most likely due to the high calorific value associated with the high organic matter in the undigested sludge. Fig. 3 shows the thermal energy balance profile as a function of sludge ODM and DM.

Fig. 3 Energy balance of sludge drying with respect to the ODM and DM content (adapted with permission³³).

2.1.2 Fate of nutrients. During thermal drying, small quantity of carbonates can be volatilised; however, organic carbon mostly remains with the sludge. Similarly, phosphorus and other macronutrients remain entirely within the solids as they are generally thermally stable.³⁶ A fraction of ammonia will be volatilised with either the hot air stream (for direct dryers), or in the condensate (for indirect dryers), particularly if the pH is high (>8.0). In both cases, the ammonia can be removed by wet scrubbing and returned to the head of the plant. It is estimated that thermal drying of activated sewage sludge at 160 °C using a continuous paddle drier can release 0.73–1.03 g NH₃-N per kilogram of total sludge solids.³⁷ Approximately 9.8–11.2% of nitrogen (4.9–5.5 g N per kilogram of total solids) was released from dewatered sludge during drying with a hot oil drum dryer at 240–260 °C.³⁸ The amount of ammonia to be volatilised depends on whether it is digested sludge (∼800–1000 mg_N L⁻¹), waste-activated sludge (<100 mg_N L⁻¹), or primary sludge (which can vary). The solar drying process will stabilise the organic nitrogen content in the dried biosolids but reduce the ammonium-nitrogen content due to ammonia volatilisation, which is proportional to moisture loss.³⁹ Intensive volatilisation of organic nitrogen (from biosolids protein content) to odorous gases such as NH₃ and H₂S is typical if the drying temperature is too high (>150 °C).⁸

2.1.3 Fate of microbial, organic, and metal contaminants. Thermal drying processes such as drum drying, agitated conductive drying, solar drying and fry-drying can reduce harmful bacteria and pathogens in biosolids to below detectable levels only when time and temperature requirements are met.⁴⁰ The thermal drying process can reduce organic load in sludge via volatilisation into the ambient environment.⁴¹ For instance, it is reported that more than forty volatile organic compounds (VOCs) with a total yield of 545–591 mg per kg dry weight are released during the thermal drying of sewage sludge.⁴² Low-temperature (60–80 °C) thermal drying of sewage sludge was observed to increase the emissions of malodorous VOCs, thereby changing the odour categories and increasing the odour intensity from moderate level (8–9) for raw sludges to strong (>10) for dried sludge.⁴³ It was also reported that higher temperatures (150–200 °C) can remove labile organic matter and convert chemically extractable organic matter into more stable organic forms.⁴¹ Shanahan et al. noted that the evaporative solar dryer reduced the viruses, helminths and bacterial indicators such as Salmonella sp. and E. coli in dried Australian biosolids.⁴⁴ Yang et al. also reported that thermal effect during thermal hydrolysis of sewage sludge played a critical role in the elimination of human pathogenic bacteria by more than 86.3% and increased to >99.9% reduction when the thermal hydrolysis was combined with mesophilic anaerobic digestion.⁴⁵ Most metals are thermally stable and not expected to be substantially volatilised during typical thermal drying conditions. However, mercury is a highly volatile metal and can be released during thermal drying at high temperatures (>150 °C). For example, Dziok et al. reported that about 34% and 62% of mercury present in biosolids was lost during thermal drying at ≤200 °C and 200–300 °C, respectively.⁴⁶ Lastly, it has been demonstrated that the thermal drying of biosolids under typical conditions can induce certain extent of PFAS transformation and volatilisation, requiring the monitoring of dryers off-gas for PFAS levels.^47,48 A positive linear relationship was observed between thermal dryer inlet temperature and quantified PFAS reduction (high temperature indicates higher removal) with an average reduction of 15–30%.⁴⁹

2.1.4 Capital and operating costs. In the USA, dried biosolids were sold in the range of US$0–36 per t in 2006, having a median price of US$10 per t.³⁰ Substantial additional value can be created by blending with mineral fertilizers. For example, milorganite sells at US$800 per t, with demand exceeding supply. Due to limited supply and sales channels not being fully developed, dried biosolids are currently not sold in Australia. These limitations could further hamper the estimated net drying cost of US$323 per dry tonne for indirect and US$441 for direct drying and the capital expenditures (CAPEX) being around 30% of the net cost.⁵⁰ More recent information has indicated that costs of biosolids thermal drying have increased by 20% in 2006, with CAPEX doubling and OPEX rising in line with energy pricing.³⁰ Operational costs are dominated by thermal energy input, which is strongly related to the feed. Drying can achieve economic break-even when the product has brand recognition, but generally, the sale price of dried biosolids is lower than the net present cost of drying.

2.2 Incineration

Incineration is a high-temperature (800–1200 °C) thermal process that completely combusts biosolids in the presence of air (with or without other fuels), converting them to ash, heat, and gases. This process significantly reduces the volume of biosolids by over 90%, produces recoverable heat energy, and can also destroy harmful contaminants and pathogens.⁵¹ Incineration is widely employed in Asia, North America, and Europe for managing biosolids. The number of incinerators in Japan was around 1173 in 2013, whereas this number in Korea was close to 476 in 2014.⁵² In the USA, 16% of biosolids are disposed of via incineration in about 340 incinerators,⁵³ and in Europe, approximately 28% of biosolids are disposed of by incineration.⁵⁴ Incineration has received serious disapproval in the recent past primarily due to the introduction of EPA emission regulations in 2011 and enforced from 2016.⁵⁵ In the USA, nearly 30 incinerators (mainly multiple hearth furnaces) have been decommissioned in the last 10 years.

Incineration is effectively a two-stage process consisting of drying wet biosolids followed by combustion. Initial drying is required to achieve the necessary flame temperatures required for emission control. Drying can be done initially in the incinerator but with lower efficiency than a dedicated dryer unit, which is economically justified in plant sizes above 20 ML d⁻¹.⁵⁶ The most common incinerator types in the US are multiple hearth furnace (MHF) and fluidised bed furnace (FBF).⁵³ A MHF consists of 5–12 hearths, and the solids are fed at >20% dry solids. MHFs are relatively cheaper (low capital cost), and reasonably energy efficient. However, they are difficult to operate, mechanically complex, and produce high-emission loads (due to variable combustion zones). Fluidised bed furnaces utilise an inert bed (normally sand) as a heat reservoir and fluidizing agent. Sand is fluidized by the added air, which ensures uniform drying and combustion and hence a consistent combustion temperature. Employing fluidising media in FBF furnaces facilitate easy operational control but reduce energy efficiency (as the inert bed material must be preheated to 490 °C). While an FBF is more easily controlled than an MHF, it is more difficult than a standard combustion process, since both air–fuel mixture and combustion temperature must be controlled to minimize emissions. To address this, even when sludge is sufficiently dry to maintain combustion auto-ignition, gaseous fuel must be continuously fed, which reduces energy efficiency.

2.2.1 Energy requirements. Electrical energy production from the flue gas generated from the incineration of mechanically dewatered biosolids is not practically possible. Even thermally dried sludges have a marginal ability to recover electrical energy, as shown in Fig. 4. Modern incinerators with high efficiency can achieve net thermal efficiencies of 50–70% at 60–80% combustion thermal efficiency. That is, 50–70% of the net calorific energy in the dry sludge can be recovered internally and utilised for drying. A dedicated waste gas heat recovery is required to achieve a 50% net thermal efficiency, and a dedicated indirect dryer, utilising steam from the furnace, may be required to achieve higher efficiencies. A minimum cake solid of 28–33% is required to sustain combustion and avoid the requirement of extra natural fuel.⁵⁷ Supplementary fuel is required (in substantial amounts) for startup/shutdown and for emission control.

Fig. 4 Feed dry solids vs. energy recovery for achieving 70% and 50% net thermal efficiency.

Thermal energy balance is also impacted by the gross calorific value (GCV) or higher heating value (HHV) of the biosolids feed material. GCV or HHV is the maximum recoverable thermal energy including the latent heat recovery of the generated water vapour. On the other hand, net calorific value (NCV) or lower heating value (LHV) is the recoverable thermal energy excluding the latent heat recovery of the generated water vapour. Calorific value is a measure of energy content and mineral content and can vary substantially depending on the redox of the inputs. HHV impacts many other design and operational parameters including combustion flame temperature, optimal air mix, and net thermal efficiency. The LHV or NCV can be calculated from the HHV using eqn (2) (the second term represents latent heat of vaporisation of the evaporated water):⁵⁸

LHV (MJ kg⁻¹) = HHV − 0.02454(M + 9H) (2)

where H and M are hydrogen and moisture contents in wt%, respectively. For most sludge materials, the hydrogen content is reported in the range of 5–8% of the total organic solids.^59,60 It is generally more convenient to utilise HHV for energy balances, incorporating the energy of vaporisation or condensation where necessary through explicit calculation. The values from studies that recorded multiple sample analyses have identified HHV as 16–21 MJ kg⁻¹ DS for primary sludges, 15–17 MJ kg⁻¹ DS for activated sludge, 10–14 MJ kg⁻¹ DS for digested sludge, and 7–10 MJ kg⁻¹ DS for lagoon sludges.⁵⁹

HHV is ideally measured in a bomb calorimeter, though there are a large number of correlations that use proximate compositions (ash, volatile matter, and fixed carbon) or ultimate compositions (CHONS), or both.^61,62 These have mean absolute errors of the order of 4–7% on plant biomass.⁶² In general, proximate analysis correlations are biomass specific, while ultimate analysis correlations consider variation in biomass compositions more thoroughly. Ideally, correlations which include a minimum of CHON should be utilised (since these vary in biosolids). Assessing the correlations reported in previous studies,^61,62 which are general for biomass, eqn (3) from a previous study⁶² (source ref. 63) fulfils the above-mentioned criteria (and also includes an ash correction), and predicts HHV of both reference materials (CH₄, ethanol) and sludges effectively:^59,60

HHV (MJ kg⁻¹) = 0.3491C + 1.1783H + 0.1005S − 0.1034O − 0.0151N − 0.0211Ash (3)

where C is carbon, H is hydrogen, S is sulfur, O is oxygen, and N is nitrogen.

HHV is fundamentally linked to chemical oxygen demand (COD) of sludge and 1 kg COD is correlated to an HHV of 13.875 MJ kg⁻¹ (corrected from ref. 64). However, accurately measuring sludge COD can be difficult as the chemical COD test does not oxidise all materials in the sludge matrix. This means that HHV calculations using COD correlations can show substantial variation. In a recent study, Schaum et al. reported a HHV coefficient of 14.98 MJ kg_COD⁻¹.⁶⁴

2.2.2 Fate of nutrients. At moderate temperatures (∼450 °C) and in an aerobic atmosphere, organic phosphorus in sewage sludge gets converted into inorganic phosphorus, whereas the majority of the phosphate combines with the metal ions in the temperature range of 600–800 °C and forms phosphorus-containing minerals.³⁶ Nitrogen in sewage sludge is partly retained in the incinerated sewage sludge ash (ISSA) and mostly released in the gas phase contributing to nitrogen-based emissions during the incineration process.¹⁵ Nitrogen is first released as NH₃ during the drying stage of the incineration process. NH₃ further converts into N₂ and NO to produce N₂O through a series of reactions, depending on the incineration temperature and air-fuel ratio, as depicted in Fig. 5.⁶⁵ Agricultural application of the ISSA is generally not feasible due to the presence of heavy metals, which are heavily concentrated in the ISSA. In addition, the phosphorus in ISSA cannot be absorbed by plants due to the presence of aluminium, iron and formation of stable calcium at high temperatures.³⁶ Globally, the majority of ISSA is sent to landfill. Other applications such as utilisation as construction materials (bricks and cement) can also be considered as promising options.⁶⁶

Fig. 5 Nitrogen transformation during sewage sludge combustion (adapted with permissions⁶⁵).

2.2.3 Fate of organic and metal contaminants. Emissions during sewage sludge incineration include nitrogen oxides, ammonia, carbon monoxide, sulphur dioxide, hydrogen chloride, hydrogen fluoride, metals, VOCs, polycyclic aromatic hydrocarbons (PAHs), polychlorinated dioxins and furans, and dioxin-like polychlorinated biphenyls.⁶⁷ A study on full-scale sludge incineration reported about 3.9–525 and 1.2–127.8 μg m⁻³ PAHs in the raw flue gas and stack flue gas, respectively.⁶⁸ Many contaminants mainly carbon, sulfur, and nitrogen gas emissions must be continuously monitored. Removing pollutants from exhaust gas can be expensive when in significant concentrations. Hence, emissions must be minimised via process control, with exhaust gas treatment to remove residual gas pollutants. Biosolids' biochemical compositions such as volatile matter, inert materials, mineral matter and ash elements also impact the product distribution and emissions profile. The majority of persistent organic pollutants (POPs), including PCBs and PFASs are oxidised above 500 °C with short residence times (<2 s) for temperatures above 800 °C.⁶⁹ Dioxin and furans are destroyed during biosolids combustion but may re-form during flue gas combustion via de novo synthesis. The relatively low levels of halogens in biosolids make dioxin and furan control easier during incineration compared to MSW incineration.^70,71 Carbon monoxide and VOC emissions can be controlled by regulating combustion temperature and air-to-fuel ratio. Maintaining a 30% increase in stoichiometric oxygen concentration and a combustion temperature >700 °C minimise CO emissions from biosolids incineration in the order of 1000 ppmv.⁷² Sulfur progresses through a stepwise combustion process from reduced sulfur (H₂S and C–S) to SO₂, SO₃ and H₂SO₄. SO₂ can be minimised by employing control measures, such as post-combustion using natural gas. Uniform combustion temperature is difficult to achieve in MHFs, so FBFs is dominating newer incineration installations. The alkali and alkaline earth metals in sludge can cause slagging and agglomeration in the furnace. Most of the heavy metals are retained in the ash during the incineration of sewage sludge. However, highly volatile metals such as Hg, Cd, As and Pb may be released in the gas phase.⁷³ Of these, only Hg emissions from sewage sludge incineration could exceed the legal limits in certain cases.^74,75

2.2.4 Capital and operating costs. The incineration plant's capital investment and operating cost depend on several factors, including the sludge characteristics (such as solids content and calorific value) and plant capacity. Therefore, the CAPEX of the incineration plant can be estimated using eqn (4):⁷⁶

CAPEX = 15 797 × P^0.82 (4)

where P is the estimated electrical energy of the plant in kWh per year, which is dependent on the lower calorific value of sludge and can be calculated using eqn (5):where LCV, W and N denote the lower calorific value of sludge (kJ kg⁻¹), estimated sludge generation (tonne per day) and number of working hours, respectively. η, k and C_f represent the sludge to electricity generation efficiency, unit adjustment constant and capacity factor, respectively.

The CAPEX and OPEX, in million USD per year, of the incineration plant can also be estimated as a function of plant capacity (C) in 1000 tonnes per year using eqn (6) and (7), respectively.⁷⁷

CAPEX = 2.3507 × C^0.7753 (6)

OPEX = 0.0744 × C^0.8594 (7)

The CAPEX of incineration plants was estimated at US$190–1000 for a capacity of 1 tonne per year with OPEX of US$12–61 per wet tonne. The capital investment and operational expenditures of incineration plants vary with location. For instance, the capital cost of incinerators for waste-to-energy generation was between US$190 and 400, 600 and 830, and 600 and 1000 per tonne per year in China, USA and Europe, respectively.⁷⁸ The operating cost of similar incinerators was estimated between US$12 and 22, 44 and 55, and 25 and 30 per tonne of waste in China, the USA and Europe, respectively. On the contrary, the CAPEX and OPEX of MSW incineration was estimated between US$400 and 700 and US$40 and 70, respectively, per tonne of MSW per year.⁷⁹ It is worth noting that the operational and maintenance costs significantly vary with the solid content of the sludge. For instance, the operating and maintenance costs of incinerating biosolids having solid contents of 33 and 47% were about US$60.55 and US$22.78 per wet tonne, respectively. The OPEX per tonne of dry biosolids incinerated was US$183.49 and US$48.47 for 33 and 47% solid contents, respectively.

2.3 Dry thermal processes (torrefaction, pyrolysis, and gasification)

Torrefaction, pyrolysis, and gasification are thermal conversion processes designed to operate under atmospheric conditions and elevated temperatures (300–800 °C) under different gas environments, such as inert and reactive. The three processes operate similarly over a spectrum of temperature and time scale. Pyrolysis is generally conducted in the absence of air or oxygen (inert environment), while gasification may be with oxygen or air (reactive environment) or without (thermal cracking). The basic difference between pyrolysis and torrefaction is that torrefaction employs lower temperatures (200–300 °C) in comparison to pyrolysis at 400–700 °C, and are both conducted at low heating rates in an inert atmosphere.^80,81

Operating below the biochar temperature/timeline produces bio-coal and highly volatile organics (as in the torrefaction process). Operating above the biochar temperature/timeline results in biochar, either via slow pyrolysis or via fast pyrolysis. Gasification is conducted at elevated temperatures and results in minimal residual carbon with the main product referred to as syngas. The residual carbon can be partly oxidised with the addition of steam or other oxidising agents, resulting in the production of activated carbon.^80,81

Pyrolysis and gasification have been extensively demonstrated for biochar and bioenergy production from biosolids. However, these works are mostly based on laboratory-scale batch operation rather than continuous pilot or full-scale demonstrations. It is worth mentioning that though the literature in the area is informative, key technical challenges around reactor design and scale-up, technology variation on energy requirement, product quality, and emission profile as well as product application limitations are often underreported. For example, the most important environmental impact of using biochar on land is for carbon sequestration as well as renewable bioenergy and greenhouse gas emission reductions from utilising pyrolysis- and gasification-derived products as opposed to fossil-based fuels. Nevertheless, pyrolysis/gasification could have negative impacts on the environment including direct and indirect greenhouse gas emissions, energy requirements for pre-processing of feedstock and flue gas cleaning, and the production of pollutants with human or environmental toxicity.

2.3.1 Energy requirement. Pre-processing high water content sludge is one of the most energy-intensive steps for pyrolysis and gasification. For pyrolysis and gasification to achieve a neutral thermal energy balance, the moisture content of the solid feed must be less than 25%.⁸² Dewatering and drying stages are major contributors to greenhouse gas emissions and energy consumption during sewage sludge treatment.⁸³ Therefore, improvement in sludge dewatering technology is essential for the widespread adoption of pyrolysis and gasification as sustainable biosolids treatment techniques.

Nevertheless, pyrolysis and gasification processes generate product streams whose energy content exceeds that of the parent feedstock helping to generate net positive thermal energy.⁸² It has been reported that recycling the excess heat for drying incoming feedstock and/or for supplying the process energy can significantly improve the overall energy performance of these technologies. Using the bioenergy recovered from the syngas or bio-oil product instead of fossil-derived energy can significantly reduce greenhouse gas emissions and lower the carbon footprint of pyrolysis or gasification processes. The heat released during biosolids pyrolysis or gasification can also be captured and used to dry feedstock, thus lowering the need for external energy.⁸¹ Furthermore, integrating pyrolysis or gasification systems with existing sewage treatment units such as anaerobic digestion and dewatering/drying could be more efficient as it allows for the effective reuse of excess heat, biosolids feedstock, and products within the treatment system.⁸¹

2.3.2 Fate of nutrients. During pyrolysis and gasification of biosolids, nitrogen is redistributed into biochar, bio-oil/tar, and gas products with nitrogen transforming from organic nitrogen to NH₃, HCN, N-heterocyclics, char-N, and NOx (at higher temperatures).⁸⁴ At moderate temperatures (<550 °C), char has similar or slightly higher N contents to the biosolids feed material and as the temperature increases, the nitrogen content in char decreases. Furthermore, nitrogen in char is not plant available but is mostly converted to non-volatile organic N via the Maillard reaction. During pyrolysis/gasification, high temperatures and the presence of oxygen facilitate nitrogen volatilisation and transformation. The organic N is released and volatilised via amino acid decarboxylation,⁸⁵ and HCN is formed mainly from the reaction between ammonia and reduced carbon.⁸⁶

Volatile organic nitrogen in biosolids is mostly converted to ammonia, but the relative fraction changes from 50% conversion of feed nitrogen to 80% conversion as the process temperature increases from 400 to 800 °C. As the temperature increases, other volatile nitrogen compounds progress from volatile organic nitrogen compounds to HCN, ammonia, and finally, nitrogen gas.⁸⁶ Elevated levels of HCN require post pyrolysis volatile combustion or reforming of mixed pyrolysis gases. Ammonia recovery is possible in high-temperature pyrolysis or gasification but is more complex due to the availability of toxic compounds such as cyanide and sulfides.

Phosphorus and potassium are other critical nutrient elements in biosolids. During pyrolysis/gasification, the majority of P and K are retained in the resulting char product with generally more than 80% recovery reported. In addition, the concentration of these elements in the char typically increases by at least 1.5-fold compared to the feed, depending on the process temperature.⁸⁷ The increase in the concentration of P and K in the char residues during thermal treatment of biosolids is mainly due to volume reduction and the relative thermal stability of these elements. However, the retention of P and K is accompanied by some transformation induced by temperature, chemical speciation of the elements, and mineral composition of the biosolids.⁸⁸ For example, labile P and K in biosolids are transformed to reducible and oxidisable fractions in the biochar, forming complex minerals with major inorganic elements in the char such as Ca, Si, Al, and Fe.⁸⁹ Notably, high-temperature pyrolysis and gasification (>700 °C) can cause K volatilisation to the gas phase, thereby reducing the K content in the biochar product. It was reported that K would volatilise at 700–800 °C from the carbon matrix having a C/K ratio of 6.2–9.5 wt%.⁹⁰ It is generally reported that the bioavailability of the N, P, and K content of biochar is generally lower compared to that of the biosolids feed attributed to the thermally induced stabilisation of the elements in the char matrix.⁹¹ It is important to optimise the thermal treatment process for maximum retention of nutrient elements in the biochar.

2.3.3 Fate of organic and metal contaminants. Pyrolysis/gasification of biosolids produces biochar with a high concentration of ash-forming elements and minerals. Increasing the pyrolysis/gasification temperature increases the proportion of inert and basic cations in the biochar product.⁹² Most heavy metals except for highly volatile elements such as Hg, As, and Cd are concentrated in the biochar with metal retention efficiency increasing with the treatment temperature.⁸⁷ As such, char from the pyrolysis/gasification of biosolids generally contain stable but higher heavy metal contents than the parent biosolids.⁹³ Most organic contaminants are expected to be decomposed during high-temperature pyrolysis/gasification treatment.⁹⁴ However, it is important to maintain adequate retention time in the primary thermal treatment (pyrolysis/gasification) unit, and subsequently in the secondary thermal treatment (gas-phase combustion) to destroy the volatilised organic compounds.⁹⁵ Halogenated organics, particularly fluorinated compounds, are reactive under reducing conditions, resulting in halide ion production.⁹⁶ This makes the reformation of halogenated organics (including dioxins and furans) less likely in the gas phase post-treatment. Reductive dehalogenation is catalysed by metals (which form metal hydrides, and subsequently react with the organic to form metal halides), which means this reaction mainly occurs in the char.

Microplastics and PFASs are substantially decomposed under pyrolysis/gasification environment as low as 500 °C.^48,97 PFAS levels in biochar are generally observed to be below the detection limit following biosolids pyrolysis and gasification at ≥500 °C. For example, more than 90% removal of PFOA and PFOS from biosolids having an input concentration of 0.2–87 ng g⁻¹ for PFOA and 13–30 ng g⁻¹ for PFOS was observed under pyrolysis conditions operating at 500–700 °C. Residual PFOS/PFOA concentration was <0.2 ng g⁻¹ or in most cases below detectable levels in the biochar.^98–101 Similarly, organic micropollutants such as pesticides, pharmaceuticals, and surfactants are also sufficiently eliminated during biosolids pyrolysis.^95,102 Polyaromatic hydrocarbons (PAHs) are formed during biosolids pyrolysis between 450 and 650 °C, reaching a peak value around 550 °C.¹⁰³ Most 5-ring or lower PAH compounds are expected to get volatilised at these temperatures and potentially destroyed in the gas-phase combustion process. Long-chain PAH compounds (above 5-ring) are largely retained in the biochar with naphthalene being the most prominent PAH compound detected in biosolids biochar.¹⁰⁴ Generally, the char product has a low level of POP contamination due to the thermal decomposition of the organic material in a reducing environment, relatively high thermal volatility of the compounds, and longer solid retention times in pyrolysis environment. However, gas phase contamination can be an issue due to shorter reactive retention time and the volatility of POPs into the gas phase, with most POPs being condensable with the tar. Notably, the quality of bio-oil from biosolids is relatively low, containing high-molecular weight aromatic nitrogenated and oxygenated compounds including PAHs.¹⁰⁵ The Diels–Alder reactions result in PAH synthesis, particularly at >500 °C, and PAHs would have a high fraction of C₅–C₆ nitrogenous compounds, including nitriles, pyridines, amides, amines and polyaromatic nitrogenated compounds.¹⁰⁶ Sulfur is largely volatilised as hydrogen sulfide but also forms carbonyl sulfides, methyl sulfides and other organosulfides, imparting a strong odour to pyrolysis tar.¹⁰⁷

Overall, the utilisation of biosolids tar product can be problematic. For this reason, volatiles are either converted post-pyrolysis in a reformer or combusted for energy in a thermal oxidiser. The gas phase emissions from the combustion of volatile products from biosolids can be carefully managed by following the best available practices for thermal plants using appropriate gas cleaning units.⁹⁵ Beyond the environmental aspects of pyrolysis and gasification, it is critical to examine the economic impacts to determine their cost-effectiveness.

2.3.4 Capital and operating costs. Primarily, pyrolysis and gasification demand a large initial capital expenditure.¹⁰⁸ The costs involved in building pyrolysis and gasification plants imply that the technologies are likely to be viable only for large-scale plants, limiting their widespread use. Compared to other technologies, the investment recovery cycle is described as 10 years or less, and the energy payback period as 2.7 years.^109,110 Generating commercially viable products, such as electricity, biochar, biofuels, and valuable chemicals will influence the economics of pyrolysis or gasification.¹¹¹ The profitability of pyrolysis or gasification products will depend on their selling price, which varies globally.¹¹² In addition, by diverting biosolids from landfills, pyrolysis/gasification gains the additional benefit of eliminating landfill tax and transport costs. The revenue generated from the sale of by-products and avoiding environmental fees will partially offset the high CAPEX costs associated with building pyrolysis and gasification plants.

EnerSludge™ set up a commercial plant in Perth, Australia, to transform 45% of the energy in the sludge feed into bio-oil. However, the plant was shut down after 16 months of operation because it was not cost-effective and raised concerns over the bio-oil quality.¹¹³ In 1986, Battelle Memorial Institute in the United States proposed a pilot plant process to produce bio-oil from sewage sludge and patented it as the sludge-to-oil-recovery system (STORS). ThermoEnergy in 2007 commercialised STORS with the following operational conditions: 275–315 °C, 11.4–14.8 MPa, and 1–3 h. The process converted sludge (20% dry solids) into a liquid fuel having 90% of diesel's HHV and solid char similar to coal.¹¹⁴ The engine could use the bio-oil to make electricity and heat. Molton et al. provided a detailed cost estimate of STORS technology for small WWTPs processing 10 000 population equivalent dry tonnes per year sludge, medium WWTPs processing 100 000 population equivalent dry tonnes per year sludge, and large WWTP processing 1 million population equivalent dry tonnes per year sludge.¹¹⁵ The estimated energy used in the process were 1410, 1394 and 1394 kWh per dry MT, and the electricity produced from oil for the small- (1898 kWh per dry MT), medium- (1480 kWh per dry MT) and large-sized population (1480 kWh per dry MT). However, no full-scale installation is currently in operation.

2.3.5 Scale-up of pyrolysis technologies. Slow pyrolysis has been effectively operated in Japan since 2007.¹¹⁶ The KORE Infrastructure has completed a six year pilot programme in Carson with the Los Angeles County Sanitation District. Following the completion of the pilot programme, KORE obtained permits to construct a commercial-scale waste conversion facility in Rialto to develop renewable natural gas for transportation.¹¹⁷ KORE's modular high-temperature pyrolysis system uses a plug flow reactor and operates at >600 °C. KORE's first commercial-scale biosolids pyrolysis facility is fully operational in Los Angeles since August 2021.

Goldhöfe in Germany has a biosolids pyrolysis demonstration plant that uses a pyrolysis drum as the reactor.¹¹⁸ The heat from natural gas combustion is used to pyrolyze biosolids in the reactor. Pyrolytic gas is used to provide heat for gas purification and recycling.

The Silicon Valley Clean Water Authority in California commissioned the BioForceTech Corporation BioDryer integrated with PYREG Pyrolysis technology in 2018. BioForceTech's “biosolids to energy plant” consists of 6 BioDryers coupled with a P-FIVE pyrolysis system. The system has a rated capacity of 14 wet tonnes per day (wtpd) of dewatered, digested biosolids with 25% total solids. The BioDryers consume only 220 kWh of energy to remove moisture from 1 tonne biosolids. The process can produce up to 320 MJ h⁻¹ heat energy used to generate hot water for the drying purposes. The pyrolysis reactor temperature ranges between 450 and 750 °C. The reported mass output of the system is 1.1 tpd biochar at 90% solid content.⁸¹

In 2015, Anaergeia's Technologies commissioned a pyrolysis system (PyroSys™) at the Encina Wastewater Authority's Water Pollution Control Facility in California. The system has a rated capacity of 12 wtpd and is designed to process digested biosolids.

RMIT University in collaboration with South East Water, Greater Western Water, and Intelligent Water Networks has developed an innovative efficient fluidised bed heat exchanger reactor (PYROCO™) with optimised heat and mass transfer for biosolids pyrolysis, which is the first of its kind in Australia. The PYROCO™ pyrolysis system was trialled on a semi-pilot scale with a capacity of 0.25 kg h⁻¹ in 2019.¹¹⁹ The pilot-scale with a capacity of 1 tonne wet biosolids per day in 2021 and 3 tonne per day wet biosolids in 2023 were trialled at a Greater Western Water treatment site in Melbourne, Australia.⁸² The fluidised bed pyrolysis reaction in a PYROCO™ reactor was carried out at 500–600 °C using oxygen lean syngas generated in the preceding gas producer unit for fluidisation and heat carrier media. The pilot-scale PYROCO™ plant demonstrated no detectable PFAS and other organic contaminants in the biochar product from biosolids pyrolysis.⁹⁵ The commercial-scale PYROCO™ demonstration plant is expected to be ready by the end of 2026.

2.3.6 Scale-up of gasification technologies. Fixed bed and fluidised bed are the popular reactor designs for pilot- and commercial-scale gasification reactors. A pilot plant operated by the Dokuz Eyrul University in Turkey used a downdraft fixed-bed gasifier for dried biosolids gasification at 1150 °C.¹²⁰ The University of Seville in Spain has developed a 100 kW pilot plant for dry biosolids gasification using a bubbling fluidised bed reactor.¹²¹ A 3 MW commercial biosolids bubbling fluidised bed gasification plant in Taiwan consists of a feeding zone, gasifier, combustion chamber, boiler and air pollution control systems.¹²²

Low-temperature circulating fluidised bed has gained popularity as a fuel-flexible gasification technology. A classic example is a 6 MW demonstration project developed by the Technical University of Denmark at DONG Energy's Asnaes Powerplant in Kalundborg, Denmark.¹²³ High ash content in biosolids feedstock necessitates ash removal, and a high flow rate for fluidisation degrades syngas quality during fluidised bed gasification. Recovery of energy-rich syngas via a commercial EBARA-fluidised bed gasification technology was demonstrated in Japan. Six TwinRec process lines were functioning in September 2002 and 14 more were planne.¹²⁴ A recent report has demonstrated that only three commercial sludge gasification plants are in operation in Europe, all located in Germany. The smaller one, located in Balingen, has a capacity of 1955 dry solids per year, the medium size, located in Koblenz, has a capacity of 4000 dry solids per year and the largest, located in Mannheim, has a capacity of nearly 5000 dry solids per year.^125,126 All plants are based on a two-stage gasification process. The KOPF gasification technology is an example of a two-stage process and includes a post-combustion chamber and a solar drying unit.¹²⁷ The solar unit dries the wet-digested sludge to a solid content of 70–85%. The fluidized bed operates at 900 °C for 30 minutes. The syngas product is combusted at 850 °C and the gas engine generates around 70 kW electricity. A post-combustion chamber is used to flare unused gas.

Logan City Council in Queensland, Australia, conducted a biosolids gasification demonstration project using a multiple hearth gasifier with a bed temperature ranging from 400 °C to 700 °C, and a 100 s retention time.¹²⁸ With a dried-biosolids feed rate of 480 kg h⁻¹ (74% of maximum capacity at 650 kg h⁻¹), soot and tar buildup in the air manifolds reduced system throughput during longer runs. However, these deposits were easily burned off during feed pauses, and an automated burn-off sequence is planned for future operation. Modifications were made to the spray absorber scrubber and barrier filter due to overwhelming non-sticky carbon and dust carryover from the pyrolysis off-gas. A venturi device was installed, and a wet electrostatic precipitator is planned for improved dust control.

At the Morrisville Municipal Authority in Pennsylvania, the Ecoremedy Fluid Lift Gasification technology, commissioned in 2019, has a rated capacity of 32 wtpd (with 27% total solids).¹²⁹ It produces a maximum energy of 2640 MJ h⁻¹ in the form of heat for thermal drying. The reported mass output of the system is 2.4 wtpd.

The City of Lebanon, Tennessee commissioned an Aries Clean Technologies Downdraft Gasification system in 2016.¹³⁰ The system has a rated capacity of 29 wet tons per day and processes a blend of waste wood, scrap tires, and dewatered, digested biosolids. It produces a maximum energy output of 420 kW electricity using a flue gas-driven organic Rankine cycle generator. The reported mass output is 1.5 wtpd.^130,131

2.4 Wet thermal processes

Raw biosolids has about 95–99% water, which can be reduced to 70–85% through mechanical dewatering.¹³² The cost of sludge management in WWTPs is nearly 50–60% of the total operating costs, and moisture removal is one of the main contributors, given the huge amount of energy required to evaporate water.¹³³ Incineration and other dry thermal treatment processes (torrefaction, pyrolysis, and gasification) require a feedstock with a moisture content below 20% to achieve net positive thermal energy.¹³⁴ Consequently, the dry thermal processes require prohibitive energy for biosolids drying, which might be avoided in wet thermal processes. This is because, in wet thermal processes, the inherent water content in biosolids is simultaneously used as the reaction medium, solvent and catalyst to decompose the organic fraction into a hydrophobic solid char product (hydrochar).¹³⁴

Wet thermal processes cover hydrothermal carbonisation (HTC), hydrothermal liquefaction (HTL), hydrothermal oxidation (HTO), and supercritical water gasification (SCWG), which vary over a spectrum of temperature and oxygen inputs (Fig. 1). Hydrothermal techniques are ideally suitable for processing highly wet biosolids or even raw sewage sludge.¹³⁵ Autogenous pressure builds up during the process, helping to keep the water in its liquid state and avoiding evaporation. Treatment in such a water-rich medium not only alleviates the loss of carbon caused due to coke formation but also takes advantage of hydrogen donated by water to enhance the quality of products. Hydrochar, bio-oil, aqueous phase and gas products are the main products of hydrothermal treatment; however, their distribution depends on the process conditions.¹³⁵ The water-soluble intermediates are formed relatively rapidly and are then converted to bio-oil and hydrochar. The liquefaction reactions can occur under mild conditions; for instance, the hydrolysis of hemicellulose and cellulose occurs at 140 and 220 °C, respectively.¹³⁶ Hydrolysis can contribute to liquid-phase products or be followed by isomerization, dehydration, decarboxylation and condensation/repolymerization to mainly produce secondary char at low temperatures (<250 °C).¹³⁷ Increasing the temperature up to the critical point of water, i.e., 374 °C, may decrease the chances of re-polymerization and accelerate the decomposition of protein, lignin, and lipid in the sludge.¹³⁷ Under HTL conditions, more liquefied products, including bio-oil or aqueous phases, are generated. Bio-oil yield depends on the extraction solvent and can substantially vary depending on the solvent used. Bio-oil formation is favoured with microbial, protein, and lipid-rich feedstock such as sewage sludge over lignocellulosic biomass feedstock.¹³⁸ The aqueous phase contains largely polar and water-soluble organic acids, alcohols, ammonia, sulfides, and other compounds, which cannot be extracted using a solvent. The aqueous phase is organic-rich with chemical oxygen demand (COD) exceeding 30 g L⁻¹, which may be anaerobically digested to produce biogas or further processed to recover energy and nutrients.¹³⁹ Furthermore, increasing the temperature above water supercritical temperature (>375 °C) generates mainly H₂, CO₂, CH₄ and CO as the products of SCWG. Fig. 6 summarises the hydrothermal processes based on their conditions and potential products. In general, for HTC (<250 °C), hydrochar is the dominant product, with 50–70% yield. For the HTL, the process occurs in the range of 250–370 °C and bio-oil is the main product with 30–60% yield. In SCWG (>370 °C), gas is the predominant product.¹⁴⁰ Inherent catalytic hydrothermal conversion occurs in sewage sludge due to the presence of high concentrations of alkali and alkaline earth metals.¹⁴¹ Additionally, hydrothermal treatment of sewage sludge or biosolids offers many advantages including reduction in waste volume and complete sterilization and eradication of microbial pathogens for biosolids stabilisation.¹⁴²

Fig. 6 Ranges of processing conditions for HTC, HTL and SCWG.

2.4.1 Fate of nutrients. A high portion of macronutrients present in biosolids, such as N, P, K, Ca, Mg, and S, are fractionally solubilised during hydrothermal processing, mostly in proportion to overall solid conversion.^143,144 Hydrochar obtained from the hydrothermal treatment of primary, waste-activated and digested sewage sludges had at least 70% retention of P, Ca and Mg contents compared to the content in the respective sludge feed. However, there was almost 100% solubility of Na and at least 50% solubility of K from the sludge feeds to the liquid products.¹³⁵ In another study, the respective N, P and K concentrations increased from 19, 1.2, and 0 mg L⁻¹ in the raw slurry feed to 2000, 50, and 200 mg L⁻¹ in the liquid products from the hydrothermal treatment.¹⁴⁵ Nahar et al. reported a similar finding during the hydrothermal treatment of various sewage sludge where the concentrations of total N, P, and K in the aqueous phase product were in the range of 850–3700, 270–730 and 1000–3500 mg L⁻¹, respectively.¹³⁵ This indicates solubilisation of macronutrients in the hot pressurised water environment at 180–270 °C. Generally, sewage sludge feedstock sacrifices its N and K contents to the aqueous-phase product during hydrothermal treatment, while P is largely accumulated in the solid hydrochar product with a significantly higher concentration than that in the feed material. The positive accumulation of P in hydrochar during the hydrothermal conversion of sewage sludge has also been reported.¹⁴⁶

Biosolids are also rich in micronutrients. For example, Australian biosolids comprise Mn, Cu, Fe, Zn, Ni, at a concentration of 50–500, 92–1996, 13 824–18 026, 210–3060, 20–250 mg kg⁻¹, respectively.¹⁴⁷ Given the main application of biosolids as soil amendment in Australia, the concentration of micronutrients must be within the environmental loading limits and helps plant growth.¹⁴⁸ Hydrothermal conversion of sewage sludge at 230 °C could slightly increase the concentration of micronutrients such as Mn, Cu, Zn and Ni. One plausible reason could be the subsequent precipitation of initially solubilized micronutrients with other chemical species.¹⁴⁹ Similar observations were reported by Wang et al.¹⁵⁰ Although some elements such as Zn and Cu are beneficial plant micronutrients, at higher concentrations, they are toxic to plants and/or animals.

2.4.2 Fate of organic and metal contaminants. In general, hydrothermal conversion produces relatively clean product streams with reduced generation of toxic gas phase emissions and combustion byproducts. Hydrolysis and reducing reactions as well as the generation of water-phase reductive radicals under sub- and super-critical conditions promote the removal of halogenated organics at lower temperatures. Non-polar compounds tend to migrate to the oil phase and, to a lesser extent, to char, while polar compounds migrate to the aqueous phase. Sewage sludge contains various undesirable components such as pathogens, heavy metals, PFAS, and POPs, namely polychlorinated biphenyls and organochlorine pesticides.¹⁵¹ Hydrothermal conversion can be effective in degrading a range of microbial and organic contaminants due to the combined effect of temperature, pressures, and OH radicals in a hydrothermal environment.^152,153 Recently, Nahar et al. conducted a review on transformation and fate of contaminants such as heavy metals, PFAS, microplastics, pharmaceuticals and personal care products, and microbial pathogens during hydrothermal treatment of sewage sludge.¹⁵⁴ It was reported that most of the microbial and organic contaminants can be effectively degraded under hydrothermal conditions, but the degradation efficiencies are component specific with some compounds exhibiting high recalcitrance to degradation. Ducey et al. demonstrated that hydrothermal processing of sewage sludge at 100 °C for 30 minutes could destroy all pathogens. In a recent study, the model microorganism (Ec DH10B) was not recovered after hydrothermal treatment, and plasmid-borne DNA was neither amplified nor detected.¹⁵⁵ PFAS is another contaminant of emerging concern in sewage sludge for which the current biological and chemical sludge treatment processes cannot be sufficiently addressed. High temperatures typically of hydrothermal processes can destabilize PFAS, facilitating their desorption from sewage sludge. Higher temperatures up to 300 °C can entirely degrade perfluoroalkyl carboxylic acids (PFCAs) but can significantly enhance the release of perfluoroalkyl sulfonic acid (PFSA) and perfluoroalkyl acid (PFAA) precursors. It was observed that the concentration of PFAAs dramatically increased from around 5.5 ng g⁻¹ in untreated sewage sludge to 80 ng g⁻¹ when the sludge was hydrothermally treated at 300 °C for 0.5 h. PFOA, PFHpA, and PFHxA compounds were completely degraded when maintained at 250 °C for 2 h or 300 °C for 0.5/2 h. However, the concentration of PFOS increased when maintained at 300 °C for 2 h. It has been suggested that more severe operational conditions are required to ensure complete PFAS transformation and destruction particularly for the more recalcitrant PFOS.¹⁵⁶ For instance, it was demonstrated that near-critical hydrothermal treatment at 350 °C in the presence of NaOH degraded more than 90% of PFOS within 15 minutes.¹⁵⁷ Nonetheless, there is need for more studies on the fate and transformation of PFASs during hydrothermal treatment of sewage sludge.

Hydrothermal treatment of sewage sludge concentrates and immobilises heavy metals in the hydrochar. However, despite the increased heavy metal concentration in hydrochar, the direct bioavailability of the metals was exceedingly low.¹⁵⁰ Under severe processing conditions, as shown in HTL and SCWG, more heavy metals may be released to the liquid phase and the hydrochar may be lean in heavy metals.¹⁵⁸ The heavy metals during hydrothermal treatment transform from weakly bounded fractions to stable fractions with reduced toxicity and bioavailability, possibly due to the formation of a complex with inert minerals and immobilisation in the crystal lattices of the hydrochar.¹⁵⁰ A detailed review of the fate of heavy metals during the hydrothermal conversion of sewage sludge was conducted by Leghari et al.¹⁵⁹ Co-hydrothermal treatment of sewage sludge with other feedstock like food and garden waste, alum sludge, or lignocellulosic biomass can help to reduce the total concentration of heavy metals by dilution effect as well as enhance metals immobilisation, reduce bioavailability and generate stable precipitates via beneficial synergistic interactions of feed materials.^160,161

2.4.3 Capital and operating costs. Since hydrothermal process needs to handle waste with high moisture content, the size of the primary reactor may be larger than processes that treat dry waste. However, dry thermal processes will require big and extensive back-end gas cleaning units, whereas the hydrothermal process will only need lean back-end gas cleaning units. Hence, the overall footprint of a wet thermal process may not necessarily be higher than that for a dry thermal process. Nevertheless, wet thermal processes operate under high pressure and require additional safety measures and design consideration (generally tubular reactors in a plug-glow configuration) with special materials needed to operate under supercritical conditions (above 374 °C and 220 bar). These factors significantly increase the cost of the hydrothermal reactor, its installation and maintenance. The capital investment of hydrothermal plant comprises the costs associated with the reactor and its accessories, slurry pump, feed pre-heater, and separators, as well as other direct and indirect capital costs. Fig. 7 shows a typical process flow diagram for HTL process for treating sewage sludge to four product streams, namely hydrochar, gas, aqueous phase and biocrude.¹⁶²

Fig. 7 Process diagram for the HTL of sewage sludge (adapted with permission¹⁶²).

Starting with the pump, several vendors reported that the pumps could effectively transfer feedstock having up to 15 wt% solid content.¹⁶³ A higher solid content means that less water must be heated and pressurized, offering a more compact plant. However, the resulting high viscous feed slurry may create problems for pumps and the low heat transfer coefficients may significantly increase the operating costs. Furthermore, pumps will likely be subjected to downtime and regular maintenance to preserve efficiency. The viscosity of the feed and transport rheology can be improved by pre-heating or thermal hydrolysis before pumping. A hydrothermal reactor is operated at high pressures. Therefore, to avoid installing excessively thick walls requiring larger diameter pipes in a single pipe train, it is suggested to install four process trains, which not only reduce the pipe diameter but also provide process redundancy.¹⁶⁴ The holding time represented by liquid hourly space velocity (LHSV) can significantly impact the reactor size and the overall capital investment. According to a report from Pacific Northwest National Laboratory (PNNL) published by the US Department of Energy, a reactor system comprising a pump, heat exchanger and plug flow reactor, handling a sludge flow rate of 139 tons h⁻¹ at 4 LHSV costs around $18.7 million.¹⁶⁵ The price of a pre-heater, usually referred to as a hot-oil system, is reported as $4.7 million and is installed to heat the feed mixture to the processing temperature. Downstream separation and material processing are fundamentally different. While phase separation can be achieved by depressurisation and flashing, the material can also be selectively separated through variation in the dielectric constant of water and using the pressure differential (for example, to extract water and gases) through membranes. However mineral and organic fouling can be a substantial issue. A phase separator, including a solid filter and an oil/water separator, costs around $4 MM for a flow rate of 553 tons kg⁻¹ solids. Ideally, the electrical work required to pressurise 1 ton of water to 220 bar is 5.56 kWh (or 40 kWh per ton-dry solids). Modelled energy consumption of HTL plants is estimated at 100 kWh per ton dry solids.¹⁶²

2.4.4 Scale-up of hydrothermal technology. The commercial deployment of hydrothermal processes is largely impacted due to the issues associated with sludge pumping, safety measures for processing at high pressures, large volume footprint, and the immature market of products.¹⁶⁶ Therefore, only a handful of companies have been able to operate successfully on a larger scale. For example, efforts are made by Southern Oil Refining Pty Ltd in collaboration with Queensland Urban Utilities and Melbourne Water to produce biodiesel by processing bio-oil from the sludge HTL.

Aalborg University in Denmark collaborated with Steeper Energy to design and construct a continuous hydrothermal liquefaction (HTL) research facility.¹⁶⁷ This facility was utilised to convert non-digested sewage sludge into HTL biocrude, water, and gas under supercritical conditions, reaching a temperature of around 400 °C and 320 bar pressure. The HTL process effectively transformed the sewage sludge into valuable biocrude, along with the generation of water and gas as byproducts.

The Altaca Energy facility, located in Gönen, Turkey, is a demonstration plant that was constructed and completed in 2016.¹⁶⁸ It utilizes the catalytic hydrothermal liquefaction technology known as CatLiq®, developed by SCF Technologies based in Denmark. This innovative plant is designed to convert diverse biomass sources, such as biogas plant digestate, forest waste, sewage sludge, agricultural waste, food plant waste, and organic household waste, into HTL bio-crude.

Licella Pty Ltd, an Australian company, has developed advanced catalytic hydrothermal liquefaction technology (Cat-HTR™).¹⁶⁹ Licella commissioned their first Cat-HTR™ pilot-scale plant in 2007. Two commercial Cat-HTR™ demonstration plants are in operation in Somersby, Australia since 2008 and 2015 which are producing 350 tonnes per year bio-oil from lignocellulosic biomass and 1 tonne per year bio-oil from plastics, respectively. Mura Technology is constructing the first commercial HTL plant for waste plastic in the UK using the Cat-HTR™ technology. Mura also aims to build a Cat-HTR™ technology-based advanced recycling plant in Japan in collaboration with Mitsubishi Chemicals. Licella is developing a Cat-HTR™ technology-based facility in Altona, Australia, for converting 20 000 tonnes per year plastics into liquid and gaseous hydrocarbon products, which has received approval from the Environmental Protection Agency Victoria.

Generally, hydrothermal processing has a low technology readiness level (TRL) of 4–6.5 when compared with a TRL of 6–9 for other thermochemical processes across the fossil and biofuels industries due to the need for higher capital investment to handle high amounts of water for processing.^169–171 Hydrothermal process is a complex treatment generating multiple product phases and has substantial safety and operational challenges due to the high pressures employed in the process.

3. Integration potential of thermal technologies with anaerobic digestion

Anaerobic digestion (AD) is a well-established biochemical process in WWTPs for treating sewage sludge by degrading volatile organic solids and converting them into biogas leaving bio-recalcitrant residues as digestate. The integration of AD being a critical sludge treatment unit with thermal technologies such as pyrolysis, gasification, and hydrothermal processes can offer promising solutions for sludge volume reduction, bioenergy recovery, contaminant destruction, and valuable product generation.¹⁷² This integrated approach combines the strengths of biological and thermal treatments to achieve more efficient sludge management and resource recovery outcomes. Coupling AD upstream of thermal process offers a promising cascaded approach, converting the residual organic matter in the AD digestate into syngas, bio-oil, and biochar via thermal processes.¹⁷³ The syngas can be used for energy production, and the organic phase of bio-oil can be used for electricity and heat, or upgraded to petroleum products or hydrogen fuel. Studies have demonstrated that integrating AD with the thermal process increases biogas and electricity production and reduces greenhouse gas emissions.^174,175 There are several potential process configurations for integrating thermal processes downstream of AD for intensive treatment of wastewater sludge and their by-products. The different cases have been grouped into AD-gasification, AD-pyrolysis, AD-HTC, and AD-HTL, as well as their process variant as summarised in Fig. 8. The practical feasibility of the proposed integration will depend on the choice of thermal treatment technology, operating requirements, quality of the sewage sludge feed (physicochemical and biochemical properties), quality of side product streams, auxiliary processing requirements, and overall conversion efficiency.¹⁷⁶

Fig. 8 Process configurations showing mass balance for processing 1000 kg of mixed sludge containing 4.9% total solids through various integrations of AD with thermal treatment (pyrolysis, gasification and hydrothermal processes). Dotted red line refers to the energy stream for heat recovery.

3.1 Description of scenarios

Scenario 1: AD-pyrolysis. In this scenario, digested sludge is dewatered and dried before being fed into the pyrolysis unit where biochar is obtained as the main product. The volatile product (oil and gas vapour) is combusted in a thermal oxidiser, and the resulting flue gas is used for energy recovery. Depending on the solid content and calorific value of the dried biosolids, the energy recovered from the flue gas can be used for AD process heat and for drying biosolids, and any excess thermal energy can be used for external heating purposes. The coupling of pyrolysis to AD of sewage sludge has been investigated in previous studies, and the synergy of the dual process has been summarised elsewhere.^177–179

Scenario 2: AD-pyrolysis-recycled pyro-oil. This process configuration is similar to scenario 1 except that the condensable volatiles is wholly or partly collected as pyro-oil and is recycled as organic substrates to the AD unit. Any remaining volatiles (mainly gas) are then used as bioenergy following combustion in a thermal oxidiser. The recycling of pyro-oil to AD can increase biogas production and shift the energy to more biogas and less to flue gas.

Scenarios 3–4: AD-gasification. In this process configuration, digested and dewatered sludge is dried to the required moisture level before being fed to a gasifier where it is converted into biochar and syngas. For scenario 3, the gasification process is considered as an endothermal low-temperature process occurring at 600 °C, while scenario 4 considers that the gasification is an autothermal high-temperature process at 800 °C. The amount and quality of the char and syngas products will depend on the gasification conditions (temperature, steam or air gasifying agents, equivalence ratio, and reactor configuration).¹⁸⁰ The generated syngas is sent to a thermal oxidiser where it is combusted, and energy is recovered from the flue gas. The recovered energy can be used in the drying unit and supplying AD process heat. The gasification of solid digestate has been demonstrated in previous works.^180,181

Scenario 5: AD-HTL-recycle HTL-oil. This configuration involves the HTL of digested and dewatered sludge. The liquid product (mixture of bio-oil and aqueous phase) is recycled as a carbon substrate to AD, potentially boosting biogas production. Any residual volatile and non-condensable gas stream are sent to a thermal oxidiser unit for energy recovery while the solid product (hydrochar) is obtained as a product stream alongside biogas.

Scenario 6: AD-HTC-recycle aqueous phase. This is similar to scenario 5 except that the hydrothermal treatment is operated in the HTC mode to generate mainly hydrochar and aqueous phase. The latter is recycled to AD to boost biogas production. Hydrochar is withdrawn as the main product stream.

Scenario 7: HTL-AD. In this configuration, the HTL of dewatered sewage sludge is first performed and the product stream (containing hydrochar, oil, and aqueous phase) after useful energy recovery is subsequently treated in AD to produce biogas. The energy recovery efficiency of coupling HTL to AD, namely HTL-AD and AD-HTL, has been assessed in a previous work, and it was observed that the overall energy shift was mainly between bio-oil and biomethane products.¹⁸²

Scenario 8: HTC-AD. This is similar to scenario 7 where hydrothermal treatment is situated upstream of AD. In this case, the HTC of dewatered sludge is performed and following heat recovery from the product stream (containing hydrochar and aqueous phase), the stream is sent to AD for biogas production.

3.2 Considerations for process integration

To determine the most efficient process integration from energy point of view, we modelled the various configurations shown in Fig. 8 (scenarios 1–8) focusing on mass and energy balance. The investigations considered the processing of 1000 kg (1 tonne) of mixed sludge (combination of primary sludge and waste activated sludge), which is the typical sludge stream being used in WWTPs for biogas production. The mixed sludge has 4.9% total solid content containing 80% volatile solids and 20% ash content. The study considered several combinations of AD with pyrolysis, low- and high-temperature gasification, HTC, and HTL. The volatiles generated during pyrolysis and gasification are combusted in the thermal oxidiser for energy recovery utilised in digestate drying, endothermic energy required for pyrolysis/gasification. Char was considered as the primary product of the thermal treatment process, and the biogas product from AD has not been combusted for energy recovery. The study also evaluated the possibility of recycling bio-oil and aqueous phase from pyrolysis and hydrothermal units for biogas production in the AD unit. The calculation only accounts for thermal energy and does not consider electricity consumption in the dewatering unit.

The AD process in this study has several key parameters and considerations. First, at a temperature of 35 °C, 400 mL of biogas is produced per gram of volatile solids (mL g-VS⁻¹) added. The biogas composition is 59 wt% CH₄ and 41 wt% CO₂, with a higher heating value (HHV) of 21.55 MJ Nm⁻³ and a density of 1.159 g L⁻¹. Additionally, the study assumed that AD converts 75% of the pyrolysis oil and 75% of the HTL oil to biogas.¹⁸³ The aqueous phase produced from HTC at 200 °C generates 400 mL of biogas per gram of COD when subjected to AD. However, the aqueous phase produced from HTL at 339 °C generates 120 mL of biogas per gram COD due to the accumulation of refractory compounds at higher temperatures, resulting in a lower biomethane potential than that of HTC.¹⁸⁴ The hydrochar produced from the HTC can be anaerobically digested to produce 400 mL of biogas per g-VS added, while that from the liquefaction process was not digested.

The study also examined the dewatering process, which increases the total solids to 27% for pyrolysis/gasification and 12% for HTC/HTL. The reason for maintaining this level of water content is because hydrothermal treatment uses water as the medium of reaction, and having enough water content is necessary to ensure a favourable conversion.

As for pyrolysis, the study assumed that at a temperature of 500 °C, the product yield for biochar, bio-oil, and gas is 40%, 35%, and 25%, respectively. Additionally, it is assumed that the heat of pyrolysis would be around 1.8 MJ kg⁻¹ which is in the range for similar biomass feedstock.¹⁸⁵ In the case of gasification, the process is assumed to operate in energy-neutral mode (autothermal), which means that no external heat energy is required. At a low temperature of 600 °C, it is assumed that 35% of the feedstock will be converted to biochar, while 65% will be converted to syngas. Tar production is considered to be negligible at this temperature. At a higher temperature of 800 °C, it is assumed that the yield of biochar will decrease to 27%, while the yield of syngas will increase to 73%. Tar production is still assumed to be negligible.

The assumptions for the dryer are that all the water will be evaporated, and the specific heat capacity of water and sludge solids is 4.18 and 1.6 kJ kg_K⁻¹, respectively.¹⁸⁶ Additionally, the latent heat of evaporation of water is assumed to be 2257 kJ kg⁻¹.¹⁸⁷ In the thermal oxidiser, the heating value of the mixed volatiles from pyrolysis is 20 MJ kg⁻¹, while that of the non-condensable gases is 12 MJ kg⁻¹.¹⁸⁸ The syngas produced from low- and high-temperature gasification has a heating value of 6 and 8 MJ Nm⁻³, respectively,¹⁸⁹ and the HTL gas product has a heating value of 2.84 MJ kg⁻¹.

HTC at 200 °C for 30 min results in a yield of 67% hydrochar, 20% aqueous phase, 10% bio-oil, and 3% gas.¹⁹⁰ The aqueous phase has a COD concentration of 21.5 g L⁻¹,¹⁸⁴ while the gas phase is mainly composed of CO₂.¹⁹¹ The enthalpy of water was obtained from the literature¹⁸⁹ and verified by the STEAMNBS method in Aspen Plus. In the HTL reaction, operating at 340 °C for 30 min results in an ash-free yield of 3.6% hydrochar, 40.2% oil, 34.6% aqueous phase, and 21.6% gas.¹⁹² The aqueous phase has a COD content of 41.2 g L⁻¹ (ref. 184) and the gas phase is assumed to consist of 30 mol% H₂ and 70 mol% CO₂.¹⁹³ The heat exchanger is also used to pre-heat the feedstock using the thermal energy of the product stream from the hydrothermal reactor. Exiting product stream temperature decreases to 100 °C and 170 °C after heat recovery in HTC and HTL scenarios, respectively.

3.3 Thermal energy performance of the integrated processes

Integrating different thermal treatment techniques with AD for primary sludge processing had different thermal energy efficiencies and environmental impacts. On the one hand, pyrolysis and gasification require a pre-dried feedstock, which adds to the energy requirements, while hydrothermal treatment does not need feedstock pre-drying. On the other hand, the requirement for high pressures in hydrothermal treatment poses a major disadvantage in terms of safety precautions and design complexity. The energy requirements across each unit operation and the overall thermal energy balance in all the scenarios investigated are summarised in Table 2. Notably, all scenarios investigated in this study resulted in a net positive energy balance ranging from 92.1 to 405.9 MJ per tonne of mixed sludge processed. It should be noted that the energy data are obtained from theoretical calculations based on certain assumptions, regarding the product yield and biomethane production potential of different organic substrates. This range of energy balance data need to be validated through experimental investigations involving the integration of various thermal processes and AD, as there are limited pilot- and industrial-scale studies reported. Nevertheless, based on the net energy balance (MJ), the energy integration efficiency of AD with thermal techniques can be ranked as high temperature gasification (405.9) > low temperature gasification (316.3) > pyrolysis (268.9) > HTL (150) > HTC (92.1). Gasification downstream of AD (scenario 3 and 4) gave the highest net positive energy due to their autothermal operation and the generation of high-volume and high-quality syngas for energy recovery, albeit at the expense of quality char product. In contrast, AD upstream of HTC/HTL (scenarios 7 and 8) gave the least net positive energy due to the low organic matter conversion, yielding low volumes of energy-carrying volatiles and gas streams. Pyrolysis generates a net positive energy value (192.4–268.9 MJ), which is in between that of gasification (316.3–405.9 MJ) and hydrothermal (228.3–92.1 MJ)-integrated processes (Table 2). This was attributed to the moderate conversion and generation of volatiles, which are either combusted for energy recovery (as in scenario 1) or condensed as bio-oil and partly recycled as organic substrates to the AD unit (as in scenario 2).

Table 2 Summary of the thermal energy balance (MJ) in various integration scenarios^a

	Scenarios	AD	Dryer	HTC	HTL	Pyrolysis	Gasification	Oxidiser	Balance
a All values are expressed in MJ for processing 1000 kg (1 tonne) of mixed sludge at 4.9% TS.
1	AD-pyrolysis	+232.7	−108.0	—	—	−390.9	—	+183.2	+268.9
2	AD-pyrolysis (recycled bio-oil)	+286.0	−108.0	—	—	−390.9	—	+534.5	+192.4
3	AD-low temp. Gasification	+232.7	−108.0	—	—	—	0	+191.6	+316.3
4	AD-high temp. gasification	+232.7	−108.0	—	—	—	0	+281.2	+405.9
5	AD-HTL-(recycled bio-oil)	+303.9	—	—	−826.6	—	—	+698.6	+228.3
6	AD-HTC-(recycled aqueous phase)	+249.4	—	−434.4	—	—	—	—	+206.0
7	HTL-AD	+244.6	—	—	−253.8	—	—	+242.3	+150.1
8	HTC-AD	+216.9	—	−124.8	—	—	—	—	+92.1

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Gasification gave the highest net positive energy value available for use in non-process heating or for grid integration. The process produced low yield and low-quality char compared to the char product from pyrolysis due to the partial oxidation of the biochar carbon. Pyrolysis sequesters carbon in biochar with a higher calorific value, suggesting that the combustion of biochar will add to the overall net energy value as in scenarios 1 and 2. In addition, the economic attractiveness of biochar from pyrolysis is higher than that from gasification considering the widely demonstrated applications of biochar, thanks to their physicochemical and structural properties.¹⁹⁴ The recycling of pyrolysis bio-oil to anaerobic digestion could help convert inherent nitrogenated compounds into biogas, a cleaner form of energy compared to bio-oil. However, the recycling of bio-oil resulted in a decrease in the total energy balance since 75% of bio-oil is converted to a biogas stream containing 41% CO₂ (as in scenario 2). Meanwhile, in scenario 1 without bio-oil recycle, the use of thermal oxidiser for bio-oil combustion may result in the production of nitrogen oxides (NOx) in the flue gas. This is an important consideration in terms of environmental impact, as NOx can contribute to air emissions. Therefore, downstream gas treatment in the scrubbing unit may be necessary to reduce NOx contents in the flue gas released into the environment.

The HTL/HTC downstream AD with recycling of bio-oil/aqueous phase (scenarios 5 and 6) shows a slightly lower energy output than that of pyrolysis (Table 2). However, the partial return of one of the liquid product streams to the upstream AD unit increases the biogas production volume, thereby generating a higher net surplus energy than in scenarios 7 and 8. This integrated process is expected to have a less emission load, particularly SOx and NOx in the gas stream due to the milder treatment conditions producing negligible volume of bio-oil for combustion. Moreover, hydrochar produced by this method has a better quality and economic value than pyrolysis and gasification char due to the preservation of surface functional groups and energy densification at low to medium temperatures. The last two scenarios involving hydrothermal processes before AD showed the lowest net surplus energy (92.1–150.1 MJ), attributed to the energy-intensive nature of processing a large volume of wet feedstock containing 88 wt% water to elevated temperatures and pressure. However, all the generated product streams, namely hydrochar, aqueous phase, bio-oil, and gas, are available for use in their raw or refined form, which can increase the economic prospects of these integrated configurations (HTC/HTL-AD).

In selecting an optimum process configuration for treating wet sewage sludge via integrated AD and thermal techniques, it is important to consider various factors, such as surplus energy available, environmental footprint, and the quality of end products (economic potential). Careful consideration of these factors holistically is crucial to achieve an effective, efficient, and sustainable treatment of wet sewage sludge. Combining AD and thermal processes presents a promising pathway for maximizing energy recovery from sewage sludge while generating value-added products such as biochar, syngas, nutrient-enriched aqueous phase, bio-oil, and platform chemicals. Other potential synergies of coupling AD upstream of thermal units include:

1. The utilisation of dewatered and dried digestates as thermal/hydrothermal feedstock can eliminate microbial and organic contaminants without the need for traditional sludge stabilisation methods.

2. Increased overall plant capacity by reducing the solid residence time in the digester by directly feeding biosolids to the downstream thermal unit (gasification/pyrolysis/hydrothermal).

3. Efficient utilisation of process heat from thermal units in other processes like anaerobic digestion, biosolids drying, and gas upgrading.

4. Pre-separation of soluble plant nutrients, such as potassium and sulfur, into the liquid fraction, mitigating corrosion and deposition issues in thermal gasification/pyrolysis.

5. Application of the thermal process char product (biochar/ash/hydrochar), which contains macro- and micro-nutrients (N, P, and K), as fertiliser and soil amendment.

6. Diversion of condensation tar and hydrothermal process water to the digester, enhancing the overall gas yield.

7. Elimination of direct gaseous emissions from biosolids stockpiling and land application, contributing to environmental sustainability.

8. The integrated approach enhances waste-to-energy conversion, improves resource efficiency, and promotes a circular economy model for the water industry.

4. Perspectives for future works

1. Development of technology assessment tool

While qualitative comparisons can be established from the current study and many other similar literature reviews, a detailed benchmarking, which is crucial for the technology/process selection, must be developed. The optimum technology/process configuration may be site specific. Therefore, developing a technology assessment tool covering technology and site-specific parameters is crucial.

2. Advances in the technologies

Commercial thermal technologies suffer from a range of bottlenecks such as poor heat and mass transfer, limited scalability and higher capital and operating costs. Future research efforts should be carried out to improve the reactor design to address the above-mentioned bottlenecks.

3. Pilot-scale and industrial-scale studies

Limited data are available in the open literature on large-scale trials including pilot plants, demonstration production plants and industrial-scale studies of biosolids thermal treatment. The availably of data from large-scale field trials is essential to understand the range of technical, environmental, and techno-commercial challenges associated with biosolids thermal treatment. This will help improving the fundamental understanding and adoption of these technologies for biosolids management in water industries.

4. Advances in energy and process modelling

Energy and resource recovery estimates published in the literature vary significantly. One of the reasons may be the empirical nature of the published models. Efforts should be made to derive process models based on the first principles to accurately predict product distribution, emission profile, mass and energy balance, and heat recovery potential.

5. Fundamental understanding of contaminant destruction/removal

Though the volume of work published on fate of contaminants during the thermal treatment of biosolids is on the rapid increase, data on the fate and transport of multiple contaminants (as inherent in biosolids) across all effluent streams are still sparsely available. Siloxane has been identified as an emerging contaminant of concern during biosolids' thermal treatment and requires deeper investigations. The studies in the literature have mainly focused on the fate of contaminants in the solid stream, mostly biochar. Therefore, extensive research is required to understand the fate and distribution of contaminants in other product streams including liquid (scrubber effluent) and gaseous streams (flue gas and stack gas), during the thermal process, and evaluate their concentrations in line with relevant environmental regulation guidelines. Thermodynamic equilibrium calculations of contaminants using computational software can provide a good understanding, though the calculations are not based on real-time conditions. This will help to take appropriate measures to overcome the environmental and other practical challenges during the design and operation of scale-up facilities.

6. Life cycle analysis

Many studies have focused on the life cycle assessment (LCA) of thermal technologies for energy recovery from different wastes. However, only a handful of studies considered different sludge compositions in LCA. It is worth noting that biosolids composition can significantly affect the technical and environmental performance of the thermal technology. Therefore, future research should focus on the comprehensive LCA of biosolids thermal processing, considering the goals of water industries including volume reduction, contaminant reduction, biochar production, and bioelectricity generation, for biosolids transformation, considering spatial and temporal variation in wastewater treatment process and biosolids production.

5. Conclusions

The following key conclusions have been reached from this critical review.

1. Drying is a critical thermal treatment process for biosolids management either as a standalone biosolids stabilisation method or as a pre-treatment step for high-temperature thermal processes. Dryers are necessary for drying and/or pelletising biosolids (if necessary) before feeding into incineration, pyrolysis and gasification reactors. The viability of indirect dryers for the thermal drying of biosolids is enhanced by energy efficiency improvements, exhaust management, and the ability to capture and utilise evaporated moisture. Thermally dried and pelletised biosolids products are potentially saleable, and a market has been established particularly in the US. However, the thermal drying of biosolids alone without high thermal treatment via pyrolysis or gasification is not sufficient to remove persistent organic contaminants in biosolids.

2. Although incineration is a well-established thermal treatment technique for biosolids management, it has been and will continue to be challenged by emission control. In addition, incineration does not allow for beneficial carbon sequestration from biosolids in the form of biochar. The solid-phase combustion process and the extensive nature of gas cleaning and emission capture equipment increase the capital and operational costs and reduce the thermal efficiency. Incinerators need external energy for startup and require sludge cake solids of 28–32% fed into a dryer to avoid any external input of thermal energy for the process.

3. Thermal treatment, particularly pyrolysis, should be regarded as an advanced, thermally efficient technique for organic pollutant destruction methods rather than a resource recovery platform. Biosolids pyrolysis at 500 °C integrated with the combustion of the pyrolysis volatile products can make pyrolysis plants achieve thermal breakeven (autothermal operations) with 22–23% cake solids. In addition, emissions can be more readily and economically controlled since combustion occurs in the gas phase at high temperatures.

4. Hydrothermal liquefaction (HTL) produces a hydro-oil that is usable than recovered pyrolysis oil but requires high temperatures and pressures (though in a small footprint due to the liquid nature of the process and less rigorous back-end gas cleaning units). The main barrier to HTL is the low technological readiness level and high capital and operating cost due to the need for specialised construction material for the reactor suitable to handle the process autogenous pressure.

5. Biochar and hydrochar are usable products, mainly for agriculture. Pyrolysis char has stable carbon structures, while hydrochar will break down, making it unsuitable for carbon fixation in soils. Both biochar and hydrochar from biosolids thermal treatment have several non-agricultural applications.

6. The integration of AD with thermal treatment processes offers beneficial synergy in organic matter conversion, energy recovery, and valuable product generation (biogas, biochar, and bioenergy). The most advanced integration of AD with thermal processes from thermal energy and contaminant destruction perspectives is AD-gasification and AD-pyrolysis. However, the integration of AD with hydrothermal treatment (HTL/HTC) can be highly favourable for sewage sludge treatment given the milder processing conditions. However, this type of configuration might require extensive downstream secondary treatment of the products.

Conflicts of interest

There are no conflicts to declare.

Data availability

No primary research results, software or code have been included and no new data were generated or analysed as part of this review.

Acknowledgements

This work is supported by the Australian Research Council (ARC) Training Centre for the Transformation of Australia's Biosolids Resource, RMIT University, Australia, under the ARC Industrial Transformation Training Centre scheme (grant number IC190100033).

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