With the ever-growing population, water, energy, and resources need to be used carefully, reused, and renewed. There is an increasing global interest in resource recovery from ‘waste’, which is driven by sustainability and environmental concerns and motivated by the potential for economic benefits. A new era in waste (water) management is being realized where wastewater treatment is becoming part of the circular economy by integrating the production of reusable water with energy and resource recovery. In this new perspective, wastewater is no longer seen as a waste to be treated with energy expenditure but rather as a valuable source of freshwater, energy, nutrients (nitrogen and phosphorous), and materials (e.g., bioplastics, cellulose fibres, and alginate). In this review paper, the conversion of wastewater treatment plants (WWTPs) into resource recovery factories (RRFs) is presented as one of the ways forward to achieve a circular economy in the water sector for the Kingdom of Saudi Arabia (KSA). The advanced technologies, some highlighted in the article, can be installed, integrated, or retrofitted into existing WWTPs to create RRFs enabling the recovery of freshwater, cellulose, alginate-like exopolymers (bio-ALE), and biogas from municipal wastewater achieving climate neutrality, decarbonization, and production of new and promising resources. The article highlights the need for modular, adaptive, and/or decentralized approaches using sustainable technologies such as aerobic granular sludge (AGS)-gravity-driven membrane (AGS-GDM), anaerobic electrochemical membrane bioreactor (AnEMBR), and anaerobic membrane bioreactor (AnMBR) for conducive localized water reuse. The increase in reuse will reduce the pressure on non-renewable water resources and decrease dependency on the energy-intensive desalination process. This article also outlines the water challenges that are arising in KSA and what are the major water research programmes/themes undertaken to address these major challenges.

• The conversion of WWTPs into RRFs for the circular economy in Saudi Arabia.

• RRFs enable the recovery of freshwater and resources from municipal wastewater.

• Modular, adaptive, decentralized technologies to promote water reuse.

• Water reuse reduces the pressure on non-renewable water resources and dependency on desalination.

• Seawater toilet flushing to reduce demand for desalination.

### Graphical Abstract

Graphical Abstract
Graphical Abstract

The grand challenges facing society in the current century are water, energy, food security, and climate change (UNDP 2022). These challenges are exacerbated due to the rapidly growing human population, leading to increased resource (water, energy, food, and chemicals) consumption with a concomitant production of large volumes of ‘used resources’, commonly recognized as waste. Water is the most widely used (100-fold greater) natural resource compared to all other natural resources combined (Rittmann & McCarty 2020). It is required to satisfy the domestic, agricultural, and industrial needs of around 8 billion people on the planet. However, readily available freshwater with adequate quality represents less than 1% of the total water on Earth (Gleick 1993). The ever-growing population, urbanization and economic development are putting unprecedented pressure on finite water resources, especially in arid regions. Around 2 billion people are living in countries or regions with ‘absolute’ water scarcity (<500 m3/year/capita), and two-thirds of the world population could be under ‘stress’ conditions (between 500 and 1,000 m3/year/capita) (FAO 2022b). The environmental services and water-related engineered infrastructure cannot handle the wastewaters of all water users and the discharge of untreated wastewaters generated by human activities contaminates freshwater natural resources.

The situation is even worse in the Gulf region, including Bahrain, Kuwait, Oman, Qatar, the Kingdom of Saudi Arabia (KSA), and the United Arab Emirates (UAE). This region is characterized by scarce rainfall (<100 mm/year), high evaporation rates (>3 K mm/year), and scarcity of renewable water resources (ElNesr et al. 2010; Aleisa & Al-Zubari 2017). The average annual per capita renewable water sources have already reached the chronic water scarcity line (<500 m3/capita/year, Table 1), which would adversely impact socio-economic development (FAO 2022b). The region relies mainly on energy-intensive seawater and brackish water desalination, followed by extraction from non-renewable groundwater resources, to satisfy their demand for water.

Table 1

Availability of water resources in GCC countries

DescriptionCountry
BahrainKuwaitOmanQatarKSAUAE
Population [million] 2022 1.8 4.4 5.5 3.0 36.6 10.1
2030 2.1 4.8 6.3 3.4 40.9 10.8
Area [km2778 17,820 309,500 11,610 2,149,690 83,600
Total renewable water resources (TRWR) [km3/year] 0.116 0.020 1.400 0.058 2.4 0.15
TRWR per capita [m3/inhab/year] 65 256 19 71 15
Total water withdrawal [km3/year] 0.434 1.250 1.873 0.881 23.8 5.1
Water withdrawal per capita [m3/inhab/year] 244 285 342 296 656 500
Water withdrawal as % of TRWR 374 6,250 134 1,519 1,000 3,370
Desalinated water produced [km3/year] 0.242 0.420 0.330 0.634 2.18 2.011
Produced municipal wastewater [km3/year] 0.151 0.292 0.108 0.274 2.7 0.751
Treated municipal wastewater [km3/year] 0.090 0.216 0.028 0.274 1.7 0.7
Reuse of treated municipal wastewater [km3/year] 0.0 0.1 0.1 0.1 0.5 0.5
DescriptionCountry
BahrainKuwaitOmanQatarKSAUAE
Population [million] 2022 1.8 4.4 5.5 3.0 36.6 10.1
2030 2.1 4.8 6.3 3.4 40.9 10.8
Area [km2778 17,820 309,500 11,610 2,149,690 83,600
Total renewable water resources (TRWR) [km3/year] 0.116 0.020 1.400 0.058 2.4 0.15
TRWR per capita [m3/inhab/year] 65 256 19 71 15
Total water withdrawal [km3/year] 0.434 1.250 1.873 0.881 23.8 5.1
Water withdrawal per capita [m3/inhab/year] 244 285 342 296 656 500
Water withdrawal as % of TRWR 374 6,250 134 1,519 1,000 3,370
Desalinated water produced [km3/year] 0.242 0.420 0.330 0.634 2.18 2.011
Produced municipal wastewater [km3/year] 0.151 0.292 0.108 0.274 2.7 0.751
Treated municipal wastewater [km3/year] 0.090 0.216 0.028 0.274 1.7 0.7
Reuse of treated municipal wastewater [km3/year] 0.0 0.1 0.1 0.1 0.5 0.5

KSA receives scarce rainfall (59 mm/year), high evaporation rates (>3 K mm/year (ElNesr et al. 2010)) and has limited (2.4 billion m3/year) renewable water resources (Figure 1, https://data.gov.sa/). There is a high reliance on non-renewable groundwater sources (constituting around 35% of KSA's water supply in 2018) and desalinated water accounts for about 59% (SWPC 2020). The total water withdrawal is, however, around 23.8 billion m3/year, out of which 82% is consumed by the agricultural sector, and 18% by the combined domestic and industrial sectors (SWPC 2019; FAO 2022b). Most of the agriculture sector is reliant on non-renewable groundwater, which is rapidly depleting (Famiglietti 2014). The current national municipal water demand is about 3.4 billion m3/year, of which about 2 billion m3/year is supplied through desalination plants (Figure 1, https://data.gov.sa/). The extensive use of non-renewable groundwater sources has contributed to KSA's high water stress level (Figure 1), calculated as a ratio of freshwater withdrawal as a proportion of available renewable freshwater resources. The country's groundwater has been almost depleted in little more than a generation (Griffin et al. 2021). The Ministry of Environment, Water and Agriculture (MEWA) has recently set a directive of reaching an urban water supply mix of 90% desalinated water and 10% non-renewable groundwater by 2030 (SWPC 2020).
Figure 1

The current situation of water availability and demand in KSA. The ratio between water demand and renewable water resources is calculated as the ratio of freshwater withdrawal as a proportion of available renewable freshwater resources (Source: FAO 2022a; KSA 2022; SWPC 2020).

Figure 1

The current situation of water availability and demand in KSA. The ratio between water demand and renewable water resources is calculated as the ratio of freshwater withdrawal as a proportion of available renewable freshwater resources (Source: FAO 2022a; KSA 2022; SWPC 2020).

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KSA, the largest producer of desalinated water, has experienced rapid economic and population growth and development over the past nine decades since oil was discovered. In 1970, KSA had a population of 5.8 million; it increased to 35 million in 2020 and is expected to increase to 40.1 million by 2030 (GAS 2020). Desalination plants will continue to be developed to cater for the growing urban water demand as well as shortages resulting from reduced reliance on ground and surface water. In 2026, the existing and committed desalinated water supply is expected to provide 2.18 billion m3/year (SWPC 2020). On the demand side, KSA is among the world's top five consumers of water on a per capita basis currently standing at 360 L/day, which is significantly higher than other developed countries (SWPC 2020). Total water demand is expected to increase to 14.5 million m3/day by 2026 with a gap of 4.5 million m3/day estimated in 2026 at the national level. Out of this national water gap, a 4.1 million m3/day gap is expected to be served through newly constructed or planned seawater reverse osmosis (SWRO) mainly due to its economic advantages (IDA 2020). SWRO requires 3–8 kWh/m3 of desalinated water for large-to-medium-size plants, respectively (Al-Karaghouli & Kazmerski 2013). For small-size plants, it can be as high as 15 kWh/m3 (Al-Karaghouli & Kazmerski 2013). Assuming an average energy consumption of 6 kWh/m3, more than 15 billion kWh energy is used by desalination plants in KSA. Currently, KSA desalination plants emit on average 2.5 kg CO2/m3 of desalinated water produced (Reddy & Ghaffour 2007), which corresponds to 5 million tonnes of CO2/year through desalination plants (producing about 2 billion m3 water/year) only.

Given the stress on water resources in this region and the number of energy resources allocated to desalination and their carbon footprint, it is imperative to maximize water-use efficiency. In fact, there is an emerging interest worldwide in resource recovery from waste streams towards a circular economy. In this regard, wastewater (i.e., used water) is no longer viewed as a waste material to treat with energy expenditure, but rather as a valuable resource of water, energy, nutrients (nitrogen and phosphorous), and materials (e.g., bioplastics, cellulose fibres, alginate, and metals). Most importantly, the reuse of treated water is becoming mandatory to fully address the water scarcity issue in KSA and many water-scarce countries around the world. Transforming wastewater treatment plants (WWTPs) into resource recovery factories (RRFs) is currently a key driver for research and development of next-generation wastewater treatment technologies. This article outlines the water challenges that are arising in KSA and the major water research programmes/themes undertaken to address these major challenges.

In the standard linear economy model, raw materials are extracted, processed, and turned into products which are then discarded at the end of their life (Figure 2). For example, in the current municipal cycle, wastewater is treated via energy-intensive processes and the treated sewage effluent is discharged back into the sea. In contrast, the circular economy model maximizes the sustainable use and value of resources, eliminating waste and benefiting both the economy and the environment (Kirchherr et al. 2017). In this model, goods and services are produced to minimize waste generation, greenhouse gas (GHG) emissions, energy consumption, and material consumption. Most of the freshwater in KSA is currently produced via desalination processes, which is energy intensive; moreover, the Gulf's waters are already 25% saltier than the global average which increases desalination costs (MEWA 2017). Brine from desalination plants (>10% salinity) is discharged back into the sea, which is detrimental to the marine ecosystem. As regulations on brine discharge become stringent and subsidies on fossil-fuel-based energy wane, future desalination cost is likely to increase. Therefore, energy-efficient desalination technologies with brine management/mining and energy-efficient wastewater treatment technologies are needed to achieve circularity and sustainability in the water sector in KSA.
Figure 2

Going circular—the difference between a traditional linear and a circular economy model.

Figure 2

Going circular—the difference between a traditional linear and a circular economy model.

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Wastewater remains a largely untapped resource for the circular economy. For example, valuable substances in wastewater include water, phosphate, and ammonium that can be used as recovered fertilizer, lipids as biofuel, and cellulose for bio-chemicals and bio-composites. In addition, wastewater provides a source to produce biodegradable bioplastics (i.e., polyhydroxyalkanoates (PHA)), a possible replacement for the petroleum-based plastics that biodegrade very slowly and affect marine life. As a result, these carbon-based elements are no longer directly converted to CO2 and/or CH4 as is happening in the current practice of wastewater treatment but are converted to valuable products leading to a decrease in GHG emissions. Smarter use of secondary raw materials will create a safe and sustainable supply of raw materials to the industry. Smarter use of resources from wastewater will also help protect the environment and at the same time preserve essential resources for current and future generations.

The circular economy model also reduces threats to future growth that come from unsustainable patterns of resource use. It takes account of the depletion and waste of natural resources as a loss to the economy, which is not captured in existing GDP measures. According to a study, a circular economy model could generate up to US$138 billion (Saudi Riyal 518 billion) in economic benefits for the Gulf Cooperation Council (GCC) between 2020 and 2030 (Bejjani et al. 2018). These benefits would largely arise from enhanced competitiveness, a lowering of environmental pressures and costs, making raw material supply chains more secure, boosting industry innovation, and generating new jobs. Desalination is the lifeline of KSA, and it provides a significant amount of fresh water to meet national water demand. There is always a desire to develop innovative desalination technologies which require less energy input and contribute less to global warming. There is rapid growth in KSA desalination capacity, and six new installations based on the reverse osmosis (RO) process will further add 3 M m3/day capacity (Narayanan 2022). The desalination demand can be reduced by using impaired quality water sources, such as seawater, for toilet flushing and treated wastewater for non-potable reuse (discussed in Sections 4 and 6). The main challenge associated with desalination is better management of brine discharge, which is especially critical when desalinating saline groundwater in inland areas. However, RO is still facing challenges, such as limited water recovery and biofouling, especially in areas where feed water has higher salinity and nutrient concentrations such as the Gulf water, further increasing the already intensive energy requirements. An assessment of the concentrations and fate of trace elements during saline water treatment will enable us to evaluate the potential for the development of directed strategies for brine mining combined with zero liquid discharge (ZLD). The current research programme on desalination aims to enhance the total water recovery and reduce energy and chemical consumption by: • Optimizing the RO process through the development of performance control and prediction tools for efficient membrane cleaning and an accurate estimation of the osmotic pressure throughout the pressure vessel to minimize the energy input and reduce (bio)fouling (Desmond et al. 2022). • Integrating RO with solar energy (Bellini 2022). • Treating the RO brine using novel direct contact evaporation and condensation (DCSEC) integrated with membrane distillation (MD) processes (Chen et al. 2021). Both DCSEC and MD operate at moderate temperatures (∼50–70 °C) which makes them suitable for the use of solar energy. The whole hybrid process is targeting close to ZLD. Both RO and the proposed novel DCSEC-MD processes will play an important role in future desalination installations due to their flexible scalability from very small-to-extra-large capacities, which will have a direct socio-economic impact on KSA. Demand for seawater desalination could be reduced if seawater is used for toilet flushing. Over 50% of the world's population resides within 60 km of the coast, and seawater can be directly utilized for non-potable uses such as toilet flushing (∼30% of the total domestic water demand) to reduce the demand on freshwater resources and desalination – an energy-intensive process. The use of seawater for toilet flushing is already in practice in several coastal cities (e.g., Hong Kong), and it is anticipated that more coastal cities may implement seawater for toilet flushing. The saline wastewater generated from seawater toilet flushing practices could produce hyperosmotic stress on the microbiome of WWTPs (Figure 3). Also, the discharge of untreated or partially treated saline wastewater could degrade the water quality of receiving water bodies. Thus, receiving water bodies cannot be directly used for producing potable drinking water (via desalination) (Panagopoulos 2022a, 2022b; Panagopoulos & Giannika 2022). Therefore, there is a need to develop more robust wastewater treatment technologies for the treatment of saline wastewaters before discharging the treated effluent to the sea or reusing it to irrigate salt-tolerant crops (Al-Khudair 2019). Figure 3 The seawater toilet flushing produces saline wastewater with a salinity of 1–1.5%. Figure 3 The seawater toilet flushing produces saline wastewater with a salinity of 1–1.5%. Close modal The aerobic granular sludge (AGS) process has been shown to be a robust technology for the treatment of saline wastewaters having a salinity of 1% (Wang et al. 2017, 2018). Similarly, marine anaerobic ammonium oxidation (anammox) bacteria, which have intrinsic tolerance to salinity compared to freshwater anammox bacteria, enriched in a membrane bioreactor (MBR) showed good removal of ammonium (nitrogen removal rate (NRR) of 0.27±0.01 kg-N/m3/day and ∼92% N-removal efficiency) from saline wastewater with a salinity of 1.2% (Ali et al. 2020a, 2019a, 2020b). These findings have significant implications for applying AGS or marine anammox process for treating saline wastewater generated from toilet flushing practices. It is anticipated that more cities will adopt seawater toilet flushing practices to relieve the stress of the ever-increasing demand for freshwater (Yang et al. 2015). Furthermore, the problems associated with the corrosion of pipelines and equipment induced by seawater toilet flushing practices can now be solved or mitigated satisfactorily with the experience gained from previous practices (Tang et al. 2007). KSA landmass experiences arid and semi-arid climates, characterized by high temperature, intense solar radiation, and extremely low annual precipitation; these conditions, coupled with the lack of surface water resources (e.g., rivers, lakes, and ponds) and nutrient-deficient sandy soils with low water-holding ability, pose significant challenges for growing plants/trees (Sultana & Nasrollahi 2018; Sahour et al. 2020). While plastic sheets have been used to reduce water evaporation from the topsoil in arid climates, their eventual disposal in landfills limits their sustainability (Kasirajan & Ngouajio 2012). In response, Mishra & co-workers have pioneered Superhydrophobic sand (SHS), a mulching approach that combines common sand or sandy soils with a nanoscale coating of paraffin wax (Gallo et al. 2022). When applied as a 5–10 mm layer on the topsoil, SHS acts as a dry diffusive barrier and dramatically curtails evaporative water loss (Figure 4). As a result, the moisture content is significantly higher in the mulched soil that plants utilize for transpiration, nutrients uptake, and so on; consequently, plant health and yield are significantly higher as evidenced from multiyear field trials (Gallo et al. 2022) and controlled environment studies on a variety of crops (Odokonyero et al. 2022). Ongoing experiments in this translational research programme include quantifying the effects of SHS mulching when irrigation is facilitated via treated wastewater effluent from an anaerobic process; time-dependent loss of SHS wax layer and water repellence due to microbial activity in the soil; and scaling up SHS manufacturing. Figure 4 Superhydrophobic sand (SHS). (a–b) Concept illustration: SHS mulch prevents evaporative loss of water from the topsoil and enhances plant health and yield. (c) The water repellence of SHS is underscored by water droplets rebounding from a packed SHS bed (top) and water's inability to infiltrate it (bottom). (d) Water repellence of SHS stems from capillary forces. (e) Field application of SHS on tomato (Solanum lycopersicum) plants in KSA. (f–g) Representative results demonstrating enhancement in tomato fruit production and barley (Hordeum vulgare) grain yield in field experiments in KSA. (h) Representative soil moisture data in a field experiment underscores SHS's ability to reduce water evaporation from the topsoil. (i) Representative image of tomato harvest—more food with limited freshwater (Note: this figure has been adapted from Gallo et al. (2022) under the Creative Commons 4.0 licence from the American Chemical Society). Figure 4 Superhydrophobic sand (SHS). (a–b) Concept illustration: SHS mulch prevents evaporative loss of water from the topsoil and enhances plant health and yield. (c) The water repellence of SHS is underscored by water droplets rebounding from a packed SHS bed (top) and water's inability to infiltrate it (bottom). (d) Water repellence of SHS stems from capillary forces. (e) Field application of SHS on tomato (Solanum lycopersicum) plants in KSA. (f–g) Representative results demonstrating enhancement in tomato fruit production and barley (Hordeum vulgare) grain yield in field experiments in KSA. (h) Representative soil moisture data in a field experiment underscores SHS's ability to reduce water evaporation from the topsoil. (i) Representative image of tomato harvest—more food with limited freshwater (Note: this figure has been adapted from Gallo et al. (2022) under the Creative Commons 4.0 licence from the American Chemical Society). Close modal Regarding the recovery of resources from sewage, there are different options, paths, and priorities. The current focus is on biogas, cellulose, bioplastics, phosphate, and alginate-like exopolymers (also known as bio-ALE). There are currently 91 medium- to large-scale centralized WWTPs in KSA treating 1.7 billion m3/year of wastewater (Table 2). The creation of new RRF is capital intensive compared to upgrading existing WWTPs to RRFs; and some of the WWTPs in KSA are quite big, especially in big cities such as Riyadh, Jeddah, and Mecca. Economies of scale can play an important role in the cost of recovered resources. This implies that RRFs will probably be created in big cities first that are served with large-scale centralized WWTPs. When experience has been gained and the costs of installations have dropped, it is expected that RRFs will also be built in smaller cities. Table 2 Capacities of WWTPs in KSA RegionNo. of WWTPTreatment capacitySludge productiona Methane productionaPhosphatebAmmoniumbEnergy demandc [No.][m3/day]kg-volatile suspended solids/daykg-total suspended solids/day[m3/day][tonne/year][tonne/year]kWh/day Riyadh 14 1,315,404 108,082 133,346 32,425 4,801 21,606 789,242 Mecca 1,181,615 97,089 119,783 29,127 4,313 19,408 708,969 Medina 350,662 28,813 35,547 8,644 1,280 5,760 210,397 Qassim 169,300 13,911 17,162 4,173 618 2,781 101,580 Eastern Province 15 1,107,625 91,010 112,283 27,303 4,043 18,193 664,575 Asir 14 169,015 13,887 17,133 4,166 617 2,776 101,409 Tabuk 112,500 9,244 11,404 2,773 411 1,848 67,500 Hail 45,850 3,767 4,648 1,130 167 753 27,510 Northern borders 19,900 1,635 2,017 491 73 327 11,940 Jazan 13 46,765 3,843 4,741 1,153 171 768 28,059 Najran 5,230 430 530 129 19 86 3,138 Al-Abha 1,440 118 146 35 24 864 Al-Jouf 36,500 2,999 3,700 900 133 600 21,900 Total 91 4,561,806 374,828 462,441 112,448 16,651 74,928 2,737,084 RegionNo. of WWTPTreatment capacitySludge productiona Methane productionaPhosphatebAmmoniumbEnergy demandc [No.][m3/day]kg-volatile suspended solids/daykg-total suspended solids/day[m3/day][tonne/year][tonne/year]kWh/day Riyadh 14 1,315,404 108,082 133,346 32,425 4,801 21,606 789,242 Mecca 1,181,615 97,089 119,783 29,127 4,313 19,408 708,969 Medina 350,662 28,813 35,547 8,644 1,280 5,760 210,397 Qassim 169,300 13,911 17,162 4,173 618 2,781 101,580 Eastern Province 15 1,107,625 91,010 112,283 27,303 4,043 18,193 664,575 Asir 14 169,015 13,887 17,133 4,166 617 2,776 101,409 Tabuk 112,500 9,244 11,404 2,773 411 1,848 67,500 Hail 45,850 3,767 4,648 1,130 167 753 27,510 Northern borders 19,900 1,635 2,017 491 73 327 11,940 Jazan 13 46,765 3,843 4,741 1,153 171 768 28,059 Najran 5,230 430 530 129 19 86 3,138 Al-Abha 1,440 118 146 35 24 864 Al-Jouf 36,500 2,999 3,700 900 133 600 21,900 Total 91 4,561,806 374,828 462,441 112,448 16,651 74,928 2,737,084 Source: https://www.stats.gov.sa/en and https://data.gov.sa/ accessed on 13 June 2022. aSludge production was estimated in kg of volatile suspended solids (VSS) and total suspended solids (TSS) per day, and methane production in m3 per day using Lenntech webpage: https://www.lenntech.com/wwtp/calculation-summary.htm bBased on typical domestic wastewater composition as described in Chen et al. (2020). cCalculated at 0.6 kWh/m3 as described by Bengtsson et al. (2019). The different resource recovery technologies are in different stages of development. Normally the stages of pre-development, development, take-off, acceleration, and stabilization are discerned (Figure 5). In the RRF, some resources are in the take-off stage such as cellulose and bio-ALE, and phosphate is in an acceleration stage. Biogas from anaerobic digestion (AD) is in the stabilization stage, whereas bioplastics are in the early stages of development (van Leeuwen et al. 2018). The total value of the recovered resources for KSA is potentially about US$3 billion/year from 2030 (Table 3). Probably many of the existing WWTPs will have been transformed into RRFs and markets for the recovered resources will be more developed by 2030.
Table 3

Forecasted rough quantities and values of recovered resources from wastewater in 2030 for KSA

DescriptionQuantities (2030)Unit cost (US$)Revenue (US$ M)
Wastewater treateda 2.8 km3/year –
Energy used for treatmentb 1.68 B kWh/year 0.085/kWhc −143
Water reuse 1.96 km3/year 1.6/m3d 3,136
Methane productione 69 M m3/year 0.2/m3f 13.8
Phosphate-Pg 28 k tonnes/year 0.4 M/k tonnesf 11.2
Cellulose 100 k tonnes/year 0.3 M/k tonnesf 30
Bio-ALE 0.73 k tonnes/year 2 M/k tonnesf 1.46
Total (US$M/year) 3,050 DescriptionQuantities (2030)Unit cost (US$)Revenue (US$M) Wastewater treateda 2.8 km3/year – Energy used for treatmentb 1.68 B kWh/year 0.085/kWhc −143 Water reuse 1.96 km3/year 1.6/m3d 3,136 Methane productione 69 M m3/year 0.2/m3f 13.8 Phosphate-Pg 28 k tonnes/year 0.4 M/k tonnesf 11.2 Cellulose 100 k tonnes/year 0.3 M/k tonnesf 30 Bio-ALE 0.73 k tonnes/year 2 M/k tonnesf 1.46 Total (US$ M/year) 3,050

aAssuming 100% wastewater will be treated, the amount of wastewater generated by 40 million people (forecasted population of KSA) is estimated at 212.5 L/capita/day.

bCalculated at 0.6 kWh/per/m3 as described by Bengtsson et al. (2019).

cElectricity tariffs as mentioned on the Saudi Electric website, accessed on 22 June.

dWater tariff as mentioned on the Marafiq website, accessed on 22 June 2022.

eMethane production from anaerobic digestion was estimated from the Lenntech website, accessed on 22 June 2022. Generated methane can be utilized to offset the energy demand for the treatment of wastewater.

fEstimates are based on the number mentioned elsewhere (van Leeuwen et al. 2018).

gBased on typical domestic wastewater composition as described in Chen et al. (2020). Phosphate will be harvested as struvite using chemical precipitation.

Figure 5

The level of development for recovery of different resources from wastewater.

Figure 5

The level of development for recovery of different resources from wastewater.

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### Water reuse

The most important resource in wastewater is water, once it has been adequately treated to prevent harm to the consumer (Rittmann & McCarty 2020). KSA uses desalination, an energy-intensive process, with average energy consumption of 6 kWh/m3 (Reddy & Ghaffour 2007; Shahzad et al. 2017), to produce freshwater to meet its municipal water demand. The used water is then treated using aerobic wastewater treatment processes such as the activated sludge (AS) process where ∼50% of the total energy (0.6 kWh/m3) (Bengtsson et al. 2019) is consumed for aeration and most of the treated effluent is released back to the sea without reuse. The need for energy for wastewater treatment might drop significantly due to the introduction of more energy-efficient technologies, e.g., AGS (Pronk et al. 2015; Ali et al. 2019b). This linear water cycle should be changed to a circular water cycle, where the treated effluent should be reused for beneficial purposes such as agriculture and irrigation, groundwater replenishment, industrial processes, and environmental restoration. The average wastewater generated in KSA is estimated at 212.5 L/capita/day (SWPC 2020), roughly generating 2.7 billion m3/year of wastewater. The Saudi National Water Strategy (NWS) highlights the need for treated sewage effluent reuse, which was 17% in 2016 and is set to reach 70% in 2030 (SWPC 2020). An increase in reuse will reduce the pressure on non-renewable water resources and decrease dependency on energy-intensive desalination processes. However, water reuse will not be able to cover the demand for agriculture and other sectors, so KSA will still rely on desalination, but future desalination plants will be coupled with solar to reduce CO2 emissions and energy consumption.

Current centralized wastewater networks and treatment plants in KSA cannot handle the increasing population demand for sanitation. KSA has vast water distribution network coverage spanning over 122,780 km in length. Whereas wastewater collection sewer lines are 47,639 km in length which is around 40% of the water distribution network (KSA 2022). Expanding the infrastructure in many cases is impractical due to physical constraints and requires prohibitively expensive capital investment. Currently, 40% of the households in KSA are not connected to the centralized wastewater network. Also, current centralized WWTPs are not designed for recycling wastewater. One approach to address this gap is by proposing new decentralized plants which would also avoid the construction of centralized sewage collection networks and could facilitate treated effluent reuse for the same community. It is proposed to integrate the AGS process with gravity-driven membrane (GDM) filtration to achieve energy-efficient decentralized domestic wastewater treatment and reuse (Figure 6) (Ali et al. 2022). The majority of full-scale centralized AGS installations worldwide are in countries where water scarcity is not an issue. To address the water scarcity issue and achieve effluent quality suitable for non-potable reuse, further polishing of the effluent from AGS would be required. Currently, MBR technology based on the AS process is the standard technology for wastewater treatment and reuse. AS-MBR plants consume on average 1 − 2 kWh/m3, where a large fraction of this energy is used for aeration, filtration process, and scouring of the membranes to minimize fouling (Bengtsson et al. 2019; Matar et al. 2016). In contrast, AGS technology requires significantly less (0.2 kWh/m3) energy (Pronk et al. 2015) than AS-MBR. Since the GDM process is a self-driven membrane filtration process operated at sub-critical flux, it does not cause frequent fouling and requires no energy input as the process is driven by the natural gravity pressure (Ali et al. 2022). Also, AGS technology can greatly reduce the footprint by 75%, hence making it suitable for decentralized wastewater treatment and reuse.
Figure 6

A schematic illustration of AGS-GDM for decentralized wastewater treatment and reuse. The AGS reactor contains microbial granules. Each granule is made of different microbial layers allowing the removal of different constituents in wastewater. Outside layers (oxic zone) are more exposed to oxygen than inner layers, which are anoxic/anaerobic. The GDM tank is a self-driven membrane filtration process operated at sub-critical flux which does not cause frequent fouling and requires no energy input as the process is driven by the natural gravity pressure.

Figure 6

A schematic illustration of AGS-GDM for decentralized wastewater treatment and reuse. The AGS reactor contains microbial granules. Each granule is made of different microbial layers allowing the removal of different constituents in wastewater. Outside layers (oxic zone) are more exposed to oxygen than inner layers, which are anoxic/anaerobic. The GDM tank is a self-driven membrane filtration process operated at sub-critical flux which does not cause frequent fouling and requires no energy input as the process is driven by the natural gravity pressure.

Close modal

The pilot testing of the AGS-GDM unit has now been successfully completed with real wastewater. The AGS-GDM system produced superior effluent quality to the neighbouring full-scale AS-MBR with higher concentrations of NO3-N and PO43-P (data unpublished). Currently, the AGS-GDM process is being scaled up under the KAUST Research Translation Program (Near Term Grand Challenge) to specifically tackle the above grand challenges. Under this project, the AGS-GDM unit for wastewater treatment and recycling will be built and demonstrated at a real-world application scale. Also, it addresses many critical aspects, including infrastructure development, water savings, economy, job growth, and impact on the people of KSA. Other technologies that integrate the production of reusable water with energy recovery are discussed in the next section.

### Integrating the production of reusable water with energy recovery from wastewater

Annually, the 91 centralized WWTPs in KSA consume about 1 million MWh—assuming an average-size WWTP requires 0.6 kWh/m3 of wastewater (Bengtsson et al. 2019). These WWTPs have the potential for biogas production of about 112 million m3/year which can cover almost 40% of the energy consumption of the water authorities (Wan et al. 2016; Bengtsson et al. 2019). It is estimated that biogas production at 112 million m3/year leads to a total saving of US$25 million/year (van Leeuwen et al. 2018). It will also lead to a reduction in the use of fossil fuels to drive WWTPs. The current research programme on wastewater treatment aims to achieve a circular economy in wastewater treatment by integrating energy production and resource recovery (energy) into the production of clean water. In this regard, anaerobic microbial technologies hold great promise for resource recovery (energy) from wastewater. As such, microbial processes based on methanogenesis or electrogenesis offer an opportunity to biologically treat the organic carbon in wastewater with the concomitant recovery of energy as methane (methanogenesis) in AD or hydrogen in microbial electrolysis cell (MEC). In MECs, efficient electroactive bacteria (EAB) such as Geobacter sulfurreducens and Desulfuromonas acetexigens (Katuri et al. 2020; Sapireddy et al. 2021), which have a unique respiratory pathway, namely extracellular electron transfer (EET) pathway, transfer the metabolically generated electrons from the oxidation of organic waste to insoluble extracellular electron acceptors such as anode (Figure 7). The bioanode in MECs is electrically connected to a cathode where proton reduction to H2 occurs using an inexpensive inorganic catalyst (e.g., nickel) (Katuri et al. 2019). One key problem in integrating MECs or AD for mainstream urban wastewater treatment is that they do not produce high-quality water suitable for reuse. To address this, MEC and AD are integrated with membrane filtration processes to generate permeate suitable for reuse. The integrated biotechnologies are referred to as anaerobic electrochemical membrane bioreactor (AnEMBR) for the case of MEC (Katuri et al. 2014, 2016; Werner et al. 2016; Sapireddy et al. 2019; El Kik et al. 2021) and anaerobic membrane bioreactor (AnMBR) for the case of AD (Wei et al. 2014; Harb et al. 2015; Shoener et al. 2016; Cheng et al. 2019). AnEMBR and AnMBR have been demonstrated successfully at the lab-scale for achieving energy-efficient treatment of domestic wastewater while simultaneously generating good quality water suitable for agriculture reuse (Katuri et al. 2016; Werner et al. 2016; Cheng & Hong 2017; Harb & Hong 2017; Sapireddy et al. 2019; Zaouri et al. 2021). However, findings from the lab-scale may not be directly applied to full-scale systems as the environmental and temporal variability in the incoming stream of raw wastewater was not accounted for in the lab studies. Currently, experiments are underway to demonstrate the feasibility of AnEMBR at the pilot-scale and AnMBR at the demonstration-scale (Figure 8) and under real conditions. Figure 7 A schematic representation of a microbial electrolysis cell for wastewater treatment with the recovery of energy. The small voltage (0.8–1.0 V) required to drive the MEC process can be supplied by renewable energy such as solar or wind (Katuri et al. 2019). Figure 7 A schematic representation of a microbial electrolysis cell for wastewater treatment with the recovery of energy. The small voltage (0.8–1.0 V) required to drive the MEC process can be supplied by renewable energy such as solar or wind (Katuri et al. 2019). Close modal Figure 8 A schematic representation of the demonstration-scale AnMBR-based wastewater treatment plant that is commissioned and currently under operation in Jeddah, Saudi Arabia. The treatment capacity for this plant is 50 m3/day and is entirely powered off-grid. Figure 8 A schematic representation of the demonstration-scale AnMBR-based wastewater treatment plant that is commissioned and currently under operation in Jeddah, Saudi Arabia. The treatment capacity for this plant is 50 m3/day and is entirely powered off-grid. Close modal Roughly, 500 tonnes of dry waste sludge is produced every day in KSA (Table 2). Off-site landfilling is the current practice for waste sludge management. Waste sludge generated from WWTPs is being collected and transported to the landfilling sites. Trucking waste sludge is costly (each trip costs around US$ 50–75), and results in traffic congestion, air pollution, and GHG emissions (methane) from landfills. Instead, the waste sludge can be used to recover resources such as cellulose, bioplastics, phosphate, and bio-ALE and later sludge can still be digested to produce biogas. Currently, in the KSA, there is only one anaerobic digester operational in a WWTP in Riyadh. The produced biogas (a mixture of CH4 and CO2), after desulphurization, is either flared or used for electricity generators in emergencies.

It is pertinent to mention that the conventional AD process often suffers from operational instability and low biogas yield, caused by substrate characteristics, short-chain fatty acids (SCFAs) accumulation, and/or ammonium ions (Hobbs et al. 2018). MECs could potentially be integrated with existing AD processes to enhance stability and facilitate SCFAs degradation (Hari et al. 2016a, 2016b). The MEC-assisted AD (MEC-AD) process successfully avoided SCFA accumulation during waste-activated sludge (WAS) degradation and suggested that exoelectrogens were much more efficient in SCFA degradation (Sun et al. 2014). Another advantage of integrating MEC with AD is the enhancement of CH4 production and purity (Feng et al. 2015a, 2015b).

Alternatively, waste sludge production can be minimized by directly cleaning wastewater using AnMBR. Over a course of a 30-week operation, we have observed that our demo-scale AnMBR generates at least 20 times lower volume of sludge compared to the aerobic wastewater treatment process that is operated at the same site and receiving the same type of influent. The AnMBR minimizes the volume of sludge that needs to be disposed of or digested, in turn bypassing the problems mentioned earlier. Instead, energy is directly obtained by converting the organic carbon in the wastewater to methane. In the demonstration-scale wastewater treatment plant (50 m3/day capacity), the AnMBR functions as the core technology for carbon removal, while UV/hydrogen peroxide is used for disinfection and further degradation of organic micropollutants (Augsburger et al. 2021). As AnMBR cannot remove ammonium and phosphate, these nutrients can either be captured first by biochar to convert into solid fertilizers (see section 6.3) or directly reused for urban farming or green landscapes grown in a controlled system (e.g., blue-green walls, (Prodanovic et al. 2017, 2019, 2020)). The nutrients and CO2 within the biogas may also be suitable to sustain microalgae farms to derive value-added products such as bioplastics, lipids, biomass-based fuels, etc. In a proof-of-concept study, mixed microalgae-bacteria photobioreactors were successfully operated to utilize nutrients that remained in effluents from AnMBR and CO2 from the biogas (Xiong et al. 2018), in turn enhancing the purity of the effluent stream and the produced CH4.

### Ammonium

Nitrogen is generally present in the form of ammonium or organic nitrogen in wastewater. In KSA, about 75 tonnes of ammonium is wasted every year in the WWTPs (Table 2). Important to realize is that, in order to feed the world's growing population, N2 in the atmosphere is converted into ammonium for use as a crop fertilizer using the Haber-Bosch process, which is an energy-intensive process (12.1 kWh/kg NH3-N), consumes about 7% of the world's natural gas, and is one of the fossil fuels causing climate change (McCarty et al. 2011). The ammonium in municipal wastewater comes from the food we eat, and conventional wastewater treatment for nitrogen removal based on nitrification and denitrification generally converts the ammonium back into N2 gas at an energy expenditure of 4.6 kWh/kg-N (Cruz et al. 2019). The current routes for ammonium removal not only neglect the energy embedded (about 12 kWh/kg NH3-N) in ammonium (Trimmer et al. 2017) but can also produce N2O, which is a potent GHG with a 300-fold stronger greenhouse effect than CO2 and is one of the most dominant ozone-depleting gases (Ali et al. 2016). Therefore, there is a need to develop alternative NH4+ management approaches that centre on the recovery of energy from NH4+ in domestic wastewater rather than dealing with its ‘destruction’ into dinitrogen (N2) (Cruz et al. 2019).

Recently, it has been demonstrated that freshwater and marine anammox bacteria can oxidize ammonium directly to N2 using an anode in an MEC as an electron acceptor. The electrons from the oxidation of ammonium are then harvested at the cathode for H2 generation in MEC (Katuri et al. 2019; Shaw et al. 2020). A key research programme is to integrate the electro-anammox process with mainstream wastewater treatment to achieve an energy-neutral or possibly energy-positive wastewater treatment (Ali & Okabe 2015; Shaw et al. 2020). This can be done by having a two-stage treatment with the anaerobic treatment of organic C by heterotrophic EAB with the recovery of energy as H2 in stage 1, followed by treatment of ammonium by autotrophic EAB (electro-anammox) with the recovery of energy as H2 in stage 2 (Figure 9). This process will be coupled with the GDM unit to achieve energy-neutral or (positive) treatment of domestic wastewater while simultaneously generating reclaimed water suitable for non-portable reuse. The proposed process has important implications for the efficient treatment of nitrogen-rich wastewater with energy recovery. This agrees with the emerging interest worldwide for resource recovery from waste streams in the context of a circular economy.
Figure 9

A schematic illustration of an AnEMBR for wastewater treatment and non-potable reuse with the recovery of energy. In MEC reactor 1, electroactive bacteria on the anode oxidize organics in wastewater and the electrons and protons generated from organic oxidation are converted to H2 at the cathode. In MEC reactor 2, electro-anammox bacteria oxidize ammonium in wastewater and the electrons and protons generated from ammonium oxidation are converted to H2 at the cathode.

Figure 9

A schematic illustration of an AnEMBR for wastewater treatment and non-potable reuse with the recovery of energy. In MEC reactor 1, electroactive bacteria on the anode oxidize organics in wastewater and the electrons and protons generated from organic oxidation are converted to H2 at the cathode. In MEC reactor 2, electro-anammox bacteria oxidize ammonium in wastewater and the electrons and protons generated from ammonium oxidation are converted to H2 at the cathode.

Close modal
Both AnMBR and AnEMBR do not remove nitrogen (10–50 mg N/L) present in wastewater, which represents an opportunity if they can be concentrated and recovered and then applied as a fertilizer. Given the low concentration of ammonium in domestic wastewaters, it is not economically cost-effective to recover ammonium using existing technologies (Cruz et al. 2019). For ammonium recovery to become economically attractive, its concentration in the waste stream should be above 2 g-N/L. One approach to concentrate ammonium from the effluent of anaerobic technologies (i.e., AnMBR or AnEMBR) is by utilizing biochar – produced by pyrolysis of waste organic biomass such as date palm leaves in the absence of oxygen – for adsorbing N in wastewater followed by the application of ‘loaded’ biochar as a fertilizer. This can be achieved using a configuration for mainstream wastewater treatment where A-stage is specifically designed to maximize the recovery of energy from organic matters in domestic wastewater as methane in AD or hydrogen in MEC, whereas B-stage is mainly dedicated to biological nitrogen removal or recovery (Wan et al. 2016; McCarty 2018; Shaw et al. 2020). Ammonium in stage B can be recovered through biochar after fine-tuning its material properties via reactor engineering and surface functionalization (Figure 10). Nutrient-loaded biochar can be utilized as a slow-release fertilizer (Zhang et al. 2022) that also locks carbon in the soil, thereby combating climate change. In doing so, a net energy-neutral wastewater treatment process can be realized.
Figure 10

A schematic representation of a general configuration of the A–B process for wastewater treatment and reuse with the recovery of energy. A-stage is specifically designed to maximize the recovery of energy from organic matters in domestic wastewater as methane in AD or hydrogen in MEC, whereas B-stage is mainly dedicated to nitrogen removal/energy recovery as in electro-anammox or nitrogen recovery using biochar.

Figure 10

A schematic representation of a general configuration of the A–B process for wastewater treatment and reuse with the recovery of energy. A-stage is specifically designed to maximize the recovery of energy from organic matters in domestic wastewater as methane in AD or hydrogen in MEC, whereas B-stage is mainly dedicated to nitrogen removal/energy recovery as in electro-anammox or nitrogen recovery using biochar.

Close modal

Phosphate (PO43−) is a scarce resource on the global scale with limited reserves, especially found in Morocco and China today; almost all the phosphate is mined and used in agriculture. Most of that phosphate ends up in waterways due to run-off and wastewater discharges (Rittmann & McCarty 2020). Phosphate causes eutrophication of surface waters (Morée et al. 2013; Rittmann & McCarty 2020). Like ammonium, phosphate in wastewater must be treated to very low concentrations (<1 mg/L) before being discharged to aquatic environments or when treated water is recycled for non-potable reuse. Currently, enhanced biological phosphorous removal (EBPR) is considered the main biological process for the removal of phosphate from wastewater (Dueholm et al. 2022). Instead of biological removal through the enhanced biological phosphorus removal (EBPR) process, phosphate can be recovered as, e.g., struvite (MgNH4PO4·6H2O). In this way, phosphate becomes available again since the recovered struvite can be applied as a fertilizer. Struvite is formed in digested sludge based on a process with the addition of magnesium chloride (MgCl2). Raw domestic wastewater contains approximately 10 mg-P/L/person/year (Chen et al. 2020), which corresponds to approximately 17,000 tonnes of phosphate/year in KSA (Table 2). Recently, a phosphate recovery facility was introduced in a WWTP in Amsterdam and other communities, which can produce 900 tonnes of struvite (van der Hoek et al. 2015, 2016). The plant treats approximately 2,000 m3 sludge/day, while the expected savings are €400,000/year (van der Hoek et al. 2016). The savings of this process consist of selling the struvite, lowering the sludge handling costs, and better sludge dewatering. Extrapolated for KSA, the recovery of phosphate could generate approximately US$11.2 million/year from the year 2030 (Table 3). ### Cellulose Cellulose is a material which ends up in sewage due to the use of toilet paper, and cellulose can be recycled from toilet paper (Ruiken et al. 2013). This recovered cellulose can be used in road construction, but further markets need to be developed (Crutchik et al. 2018). A person uses an average of 10 kg of tissue paper/year (Ruiken et al. 2013), which means that the total potential volume is approximately 420,000 tonnes/year in KSA. Nevertheless, not all tissue paper can be recovered at WWTPs. It is estimated that about 25% of cellulose can be recovered with the introduction of fine-sieving at WWTPs (Roest 2017), which will lead to approximately 100,000 tonnes/year. Although there are currently no revenues from cellulose recovery at all, as considerable investments are necessary, the potential revenue is interesting due to the expected price of this material, i.e., approximately US$300/tonne (van Leeuwen et al. 2018).

In the context of digesting cellulose to produce methane, the fermentation process suffers from low yields, and instability caused by the accumulation of acids from the degradation of complex substrates such as cellulose (Lusk et al. 2018). Coupling fermentation with MECs could be a viable approach for efficient biogas production from renewable biomass (Lu & Ren 2016). Several studies reported the integration of MEC with fermentation for cellulose waste degradation (Lalaurette et al. 2009; Wang et al. 2011; Lusk et al. 2018).

### Alginate-like exopolymers (bio-ALE)

Alginate is a naturally occurring anionic polymer. It is a polysaccharide, typically obtained from brown seaweed and has been extensively investigated and used for many biomedical applications which include wound healing, drug delivery, and tissue engineering due to its biocompatibility, low toxicity, and relatively low cost (Roest et al. 2015). Bio-ALE is an alginate-like polymer of sugars and proteins and can be used in agriculture and horticulture, paper industry, medical, and construction industries. Bio-ALE can be applied as fibres, gel, or foam. Fibres can be used for the production of absorbing tissues, gels can be used as glue for the production of fertilizer pellets, and foam can be used for the production of fire-resistant boards. The liquid version can be used for thickening inks, improving paper quality, and increasing the quality of curing concrete (Roest et al. 2015). The first bio-ALE production installations were built in Zutphen and Epe in 2019 in the Netherlands. In a field test, it was demonstrated that 0.2 kg bio-ALE can be produced per kg of Nereda AGS (HaskoningDHV 2017) and a tonne of ALE has market values of approximately US$2,000 (van Leeuwen et al. 2018). The extraction of bio-ALE from AGS can be considered up-cycling because a more valuable product is produced from waste. Recently, the National Water Company (NWC) tender a Nereda plant of 50,000 m3/day capacity in Jeddah. It is expected that more WWTPs in KSA will be retrofitted or upgraded to Nereda technology by 2030. Assuming the total capacity of Nereda plants would be 150,000 m3/day by 2030, it would potentially produce 730 tonnes bio-ALE (from 3,650 tonnes of excess sludge)/year. The sale of bio-ALE would generate revenue of about US$1.46 million/year by 2030 (Table 3).

### Bioplastic

Bioplastic can be produced from sewage sludge as a form of up-cycling. Bioplastics are formed using a complicated process during which volatile fatty acids (VFAs) are produced first and then later fed to microbes which produce the building blocks for bioplastics (i.e., PHA). There have been some experiments and tests that showed an average extraction amount of 30 g-PHA/100 g of biomass was achieved with 98% purity (de Souza Reis et al. 2020). The production cost of this material, however, is currently still rather high, and it is twice as much as the regular market price. Furthermore, there is no available stable industrial production process yet (van Leeuwen et al. 2018). Currently, plastics are produced through petroleum products made from fossil fuels. In future, there will be more demand for bioplastics to minimize reliance on fossil fuel-based plastics.

The existing centralized wastewater-related infrastructure is not able to be modified easily. The creation of new RRFs from scratch is capital intensive compared to upgrading existing centralized WWTPs to RRFs. Therefore, upgrading or retrofitting existing centralized facilities would be the most feasible route at the beginning to achieve a circular economy in the water sector. Once critical knowledge and experience have been gained, resource recovery technologies would also be implemented for newer and smaller-scale installations.

Water reuse should be preferred to achieve circularity. Currently, a centralized approach is used for wastewater management, collection, and treatment. Upgrading centralized WWTPs with water reclamation and reuse facilities could support the increase of water reuse in KSA. However, many centralized WWTPs are located far away from city centres and in areas where water demand for treated effluent is low. In such cases, physical infrastructure (such as pipelines) is needed to transport the treated effluent to areas with high water demand. There is a lack of centralized infrastructure for water reuse and building centralized water reuse infrastructure could be capital intensive. Therefore, modular, adaptive, and/or decentralized approaches should be implemented using sustainable technologies such as AGS-GDM, AnEMBR, and AnMBR, which could enable treated water reuse at a neighbourhood scale. Implementation of a modular, adaptive, and/or decentralized approach for newer installations would certainly promote water reuse and help to achieve the 2030 water reuse target (70%) by the Kingdom of Saudi Arabia.

There are already established technologies for the recovery of CH4, phosphate, cellulose, and bio-ALE from sludge, which could be easily integrated into existing installations. As mentioned above, conventional AD or modified MEC-AD could be easily integrated into existing facilities to recover CH4 from waste sludge. However, it should be highlighted that there is no infrastructure available for CH4 distribution and it could be expensive to build such infrastructure. Therefore, the generated CH4 should rather be utilized to offset the energy requirement of the wastewater treatment to make the system more energy-efficient or energy-neutral.

This work was supported by the Center Competitive Funding Program (FCC/1/1971-33-01, FCC/1/1971-44-01), Near Term Grand Challenge Funding Program (REI/1/4254-01-01), Research Translation Funding Program (REI/1/4223-01-01), and Impact Acceleration Funding Program (REI/1/4224-01-01) from King Abdullah University of Science and Technology.

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.

Aleisa
E.
&
Al-Zubari
W.
2017
Wastewater reuse in the countries of the Gulf Cooperation Council (GCC): the lost opportunity
.
Environ. Monit. Assess.
189
(
11
),
553
.
Ali
M.
,
Rathnayake
R.
,
Zhang
L.
,
Ishii
S.
,
Kindaichi
T.
,
Satoh
H.
,
Toyoda
S.
,
Yoshida
N.
&
Okabe
S.
2016
Source identification of nitrous oxide emission pathways from a single-stage nitritation-anammox granular reactor
.
Water Res.
102
,
147
157
.
Ali
M.
,
Shaw
D. R.
&
Saikaly
P. E.
2019a
Draft genome sequence of a novel marine anaerobic ammonium-oxidizing bacterium, ‘Candidatus Scalindua sp.’
.
Microbiol. Resour. Announce.
8
,
e00297
e00219
.
Ali
M.
,
Wang
Z.
,
Salam
K. W.
,
Hari
A. R.
,
Pronk
M.
,
van Loosdrecht
M. C. M.
&
Saikaly
P. E.
2019b
Importance of species sorting and immigration on the bacterial assembly of different-sized aggregates in a full-scale aerobic granular sludge plant
.
Environ. Sci. Technol.
53
(
14
),
8291
8301
.
Ali
M.
,
Katuri
K. P.
&
Saikaly
P.
2022
Waste Water Treatment System Using Aerobic Granular Sludge Gravity-Driven Membrane System
.
Office, U.S.P.a.T. (ed)
.
King Abdullah University of Science and Technology KAUST
,
US
.
Al-Karaghouli
A.
&
Kazmerski
L. L.
2013
Energy consumption and water production cost of conventional and renewable-energy-powered desalination processes
.
Renewable Sustainable Energy Rev.
24
,
343
356
.
Al-Khudair
D.
2019
Startup of the Week: Red Sea Farms in Saudi Arabia Aims to Provide Viable Solutions to Water Scarcity
.
Saudi Arabia
.
Augsburger
N.
,
Zaouri
N.
,
Cheng
H.
&
Hong
P. Y.
2021
The use of UV/H2O2 to facilitate removal of emerging contaminants in anaerobic membrane bioreactor effluents
.
Environ. Res.
198
,
110479
.
Bejjani
M.
,
Anouti
Y.
,
Klat
A.
,
Batal
J.
,
Beukel
J.-W. v. d.
,
Khalifa
T.
&
M.
2018
Putting GCC Cities in the Loop – Sustainable Growth in A Circular Economy
.
Ideation Center Insights
,
Abu Dhabi, UAE
.
Bellini
E.
2022
Solar Partly Powering World's Largest Reverse Osmosis Desalination Plant
.
Bengtsson
S.
,
de Blois
M.
,
Wilen
B. M.
&
Gustavsson
D.
2019
A comparison of aerobic granular sludge with conventional and compact biological treatment technologies
.
Environ. Technol.
40
(
21
),
2769
2778
.
Chen
G.
,
Ekama
G. A.
,
Loosdrecht
M. C. M. v.
&
Brdjanovic
D.
2020
Biological Wastewater Treatment: Principles, Modelling and Design
.
IWA Publishing, London, UK
.
Chen
Q.
,
Burhan
M.
,
M. W.
,
Ybyraiymkul
D.
,
Akhtar
F. H.
,
Li
Y.
&
Ng
K. C.
2021
A zero liquid discharge system integrating multi-effect distillation and evaporative crystallization for desalination brine treatment
.
Desalination
502
,
114928
.
Cheng
H.
,
Cheng
D.
,
Mao
J.
,
Lu
T.
&
Hong
P. Y.
2019
Identification and characterization of core sludge and biofilm microbiota in anaerobic membrane bioreactors
.
Environ. Int.
133
(
Pt A
),
105165
.
Crutchik
D.
,
Frison
N.
,
Eusebi
A. L.
&
Fatone
F.
2018
Biorefinery of cellulosic primary sludge towards targeted Short Chain Fatty Acids, phosphorus and methane recovery
.
Water Res.
136
,
112
119
.
Cruz
H.
,
Law
Y. Y.
,
Guest
J. S.
,
Rabaey
K.
,
Batstone
D.
,
Laycock
B.
,
Verstraete
W.
&
Pikaar
I.
2019
Mainstream ammonium recovery to advance sustainable urban wastewater management
.
Environ. Sci. Technol.
53
(
19
),
11066
11079
.
Desmond
P.
,
Huisman
K. T.
,
Sanawar
H.
,
Farhat
N. M.
,
Traber
J.
,
Fridjonsson
E. O.
,
Johns
M. L.
,
Flemming
H. C.
,
Picioreanu
C.
&
Vrouwenvelder
J. S.
2022
Controlling the hydraulic resistance of membrane biofilms by engineering biofilm physical structure
.
Water Res.
210
,
118031
.
de Souza Reis
G. A.
,
Michels
M. H. A.
,
Fajardo
G. L.
,
Lamot
I.
&
de Best
J. H.
2020
Optimization of green extraction and purification of PHA produced by mixed microbial cultures from sludge
.
Water
12
(
4
),
1185
.
Dueholm
M. K. D.
,
Nierychlo
M.
,
Andersen
K. S.
,
Rudkjobing
V.
,
Knutsson
S.
,
Mi
D. A. S. G. C.
,
Albertsen
M.
&
Nielsen
P. H.
2022
MiDAS 4: a global catalogue of full-length 16S rRNA gene sequences and taxonomy for studies of bacterial communities in wastewater treatment plants
.
Nat. Commun.
13
(
1
),
1908
.
El Kik
O.
,
Issa
L.
,
Katuri
K. P.
,
Saikaly
P. E.
,
Alameddine
I.
&
M.
2021
Coupling anaerobic fluidized membrane bioreactors with microbial electrolysis cells towards improved wastewater reuse and energy recovery
.
J. Environ. Chem. Eng.
9
,
5
.
Famiglietti
J. S.
2014
The global groundwater crisis
.
Nat. Clim. Change
4
(
11
),
945
948
.
FAO, F.a.A.O.o.t.U.N.
2022a
AQUASTAT Database: Municipal Wastewater Production, Collection, Treatment and use
.
FAO, F.a.A.O.o.t.U.N.
2022b
Water Scarcity
.
Gallo
A.
,
Odokonyero
K.
,
Mousa
M. A. A.
,
Reihmer
J.
,
Al-Mashharawi
S.
,
Marasco
R.
,
Manalastas
E.
,
Morton
M. J. L.
,
Daffonchio
D.
,
McCabe
M. F.
,
Tester
M.
&
Mishra
H.
2022
Nature-inspired superhydrophobic sand mulches increase agricultural productivity and water-use efficiency in arid regions
.
ACS Agric. Sci. Technol.
2
(
2
),
276
288
.
GAS, G.A.o.S.
2020
Population by Age Groups, and Gender Mid-Year 2020
.
General Authority for Statistics (GAS), Kingdom of Saudi Arabia
.
Gleick
P. H.
1993
Water in Crisis: A Guide to the World's Fresh Water Resources
.
Oxford University Press
,
Oxford, UK
.
Griffin
C.
,
Menani
M.
,
Haider
S. A.
,
Mawlawli
A. A.
,
Hughes
K.
&
Olivereau
D.
2021
Boosting Circularity Across Saudi Arabia
.
The Economist Intelligence Unit
,
Dubai, UAE
.
Hari
A. R.
,
Katuri
K. P.
,
Gorron
E.
,
Logan
B. E.
&
Saikaly
P. E.
2016a
Multiple paths of electron flow to current in microbial electrolysis cells fed with low and high concentrations of propionate
.
Appl. Microbiol. Biotechnol.
100
(
13
),
5999
6011
.
R.
2017
Kaumera Nereda® Gum
.
Hobbs
S. R.
,
Landis
A. E.
,
Rittmann
B. E.
,
Young
M. N.
&
Parameswaran
P.
2018
Enhancing anaerobic digestion of food waste through biochemical methane potential assays at different substrate: inoculum ratios
.
Waste Manage.
71
,
612
617
.
IDA, I.D.A.
2020
IDA Desalination and Reuse Handbook 2021–2022
.
International Desalination Association (IDA), Danvers, MA, USA.
Kasirajan
S.
&
Ngouajio
M.
2012
Polyethylene and biodegradable mulches for agricultural applications: a review
.
Agron. Sustainable Dev.
32
(
2
),
501
529
.
Katuri
K. P.
,
Werner
C. M.
,
Jimenez-Sandoval
R. J.
,
Chen
W.
,
Jeon
S.
,
Logan
B. E.
,
Lai
Z.
,
Amy
G. L.
&
Saikaly
P. E.
2014
A novel anaerobic electrochemical membrane bioreactor (AnEMBR) with conductive hollow-fiber membrane for treatment of low-organic strength solutions
.
Environ. Sci. Technol.
48
(
21
),
12833
12841
.
Katuri
K. P.
,
Bettahalli
N. M.
,
Wang
X.
,
Matar
G.
,
Chisca
S.
,
Nunes
S. P.
&
Saikaly
P. E.
2016
A microfiltration polymer-based hollow-fiber cathode as a promising advanced material for simultaneous recovery of energy and water
.
28
(
43
),
9504
9511
.
Katuri
K. P.
,
Ali
M.
&
Saikaly
P. E.
2019
The role of microbial electrolysis cell in urban wastewater treatment: integration options, challenges, and prospects
.
Curr. Opin. Biotechnol.
57
,
101
110
.
Katuri
K. P.
,
Kamireddy
S.
,
Kavanagh
P.
,
A.
,
Conghaile
P. O.
,
Kumar
A.
,
Saikaly
P. E.
&
Leech
D.
2020
Electroactive biofilms on surface functionalized anodes: the anode respiring behavior of a novel electroactive bacterium, Desulfuromonas acetexigens
.
Water Res.
185
,
116284
.
Kirchherr
J.
,
Reike
D.
&
Hekkert
M.
2017
Conceptualizing the circular economy: an analysis of 114 definitions
.
Resour. Conserv. Recycl.
127
,
221
232
.
KSA, K.o.S.A.
2022
Saudi Open Data Portal
.
Lalaurette
E.
,
Thammannagowda
S.
,
Mohagheghi
A.
,
Maness
P.-C.
&
Logan
B. E.
2009
Hydrogen production from cellulose in a two-stage process combining fermentation and electrohydrogenesis
.
Int. J. Hydrogen Energy
34
(
15
),
6201
6210
.
Lu
L.
&
Ren
Z. J.
2016
Microbial electrolysis cells for waste biorefinery: a state of the art review
.
Bioresour. Technol.
215
,
254
264
.
Lusk
B. G.
,
Colin
A.
,
Parameswaran
P.
,
Rittmann
B. E.
&
Torres
C. I.
2018
Simultaneous fermentation of cellulose and current production with an enriched mixed culture of thermophilic bacteria in a microbial electrolysis cell
.
Microb. Biotechnol.
11
(
1
),
63
73
.
Matar, G., Gonzalez-Gil, G., Maab, H., Nunes, S., Le-Clech, P., Vrouwenvelder, J. & Saikaly, P. E. 2016 Temporal changes in extracellular polymeric substances on hydrophobic and hydrophilic membrane surfaces in a submerged membrane bioreactor. Water Res. 95, 27–38
.
McCarty
P. L.
2018
What is the best biological process for nitrogen removal: when and why?
Environ. Sci. Technol.
52
(
7
),
3835
3841
.
McCarty
P. L.
,
Bae
J.
&
Kim
J.
2011
Domestic wastewater treatment as a net energy producer–can this be achieved?
Environ. Sci. Technol.
45
(
17
),
7100
7106
.
MEWA, M.o.E.W.A.
2017
National Environment Strategy
.
Ministry of Environment, Water and Agriculture (MEWA), Kingdom of Saudi Arabia
.
ElNesr
M.
,
Alazba
A.
,
Abu‐Zreig
M.
2010
Spatio-temporal variability of evapotranspiration over the Kingdom of Saudi Arabia
.
Appl. Eng. Agric.
26
(
5
),
833
842
.
Morée
A. L.
,
Beusen
A. H. W.
,
Bouwman
A. F.
&
Willems
W. J.
2013
Exploring global nitrogen and phosphorus flows in urban wastes during the twentieth century
.
Global Biogeochem. Cycles
27
(
3
),
836
846
.
Narayanan
S. G. N.
2022
SWCC Head Expects Water Assets Transfer to PIF-Owned WSC to Take up to 3 Years
.
Panagopoulos
A.
&
Giannika
V.
2022
Comparative techno-economic and environmental analysis of minimal liquid discharge (MLD) and zero liquid discharge (ZLD) desalination systems for seawater brine treatment and valorization
.
Sustainable Energy Technol. Assess.
53
,
102477
.
Prodanovic
V.
,
Hatt
B.
,
McCarthy
D.
,
Zhang
K.
&
Deletic
A.
2017
Green walls for greywater reuse: understanding the role of media on pollutant removal
.
Ecol. Eng.
102
,
625
635
.
Prodanovic
V.
,
McCarthy
D.
,
Hatt
B.
&
Deletic
A.
2019
Designing green walls for greywater treatment: the role of plants and operational factors on nutrient removal
.
Ecol. Eng.
130
,
184
195
.
Prodanovic
V.
,
Hatt
B.
,
McCarthy
D.
&
Deletic
A.
2020
Green wall height and design optimisation for effective greywater pollution treatment and reuse
.
J. Environ. Manage.
261
,
110173
.
Pronk
M.
,
de Kreuk
M. K.
,
de Bruin
B.
,
Kamminga
P.
,
Kleerebezem
R.
&
van Loosdrecht
M. C.
2015
Full scale performance of the aerobic granular sludge process for sewage treatment
.
Water Res.
84
,
207
217
.
Reddy
K. V.
&
Ghaffour
N.
2007
Overview of the cost of desalinated water and costing methodologies
.
Desalination
205
(
1–3
),
340
353
.
Rittmann
B. E.
&
McCarty
P. L.
2020
Environmental Biotechnology: Principles and Applications
.
McGraw Hill
,
New York, US
.
Roest
H. v. d.
,
Loosdrecht
M. v.
,
Langkamp
E. J.
&
Uijterlinde
C.
2015
Recovery and reuse of alginate from granular Nereda sludge. Water21 April
.
Roest
K.
2017
WWTP Capacity Increases by 10% with Fine Screen Technology
.
KWR Water
,
Nieuwegein, Netherlands
.
Ruiken
C. J.
,
Breuer
G.
,
Klaversma
E.
,
Santiago
T.
&
van Loosdrecht
M. C.
2013
Sieving wastewater–cellulose recovery, economic and energy evaluation
.
Water Res.
47
(
1
),
43
48
.
Sahour
H.
,
Vazifedan
M.
&
Alshehri
F.
2020
Aridity trends in the Middle East and adjacent areas
.
Theor. Appl. Climatol.
142
(
3
),
1039
1054
.
Sapireddy
V.
,
Ragab
A. a.
,
Katuri
K. P.
,
Yu
Y.
,
Lai
Z.
,
Li
E.
,
Thoroddsen
S. T.
&
Saikaly
P. E.
2019
Effect of specific cathode surface area on biofouling in an anaerobic electrochemical membrane bioreactor: novel insights using high-speed video camera
.
J. Membr. Sci.
577
,
176
183
.
Sapireddy
V.
,
Katuri
K. P.
,
A.
&
Saikaly
P. E.
2021
Competition of two highly specialized and efficient acetoclastic electroactive bacteria for acetate in biofilm anode of microbial electrolysis cell
.
NPJ Biofilms Microbiomes
7
(
1
),
47
.
M. W.
,
Burhan
M.
,
Ang
L.
&
Ng
K. C.
2017
Energy-water-environment nexus underpinning future desalination sustainability
.
Desalination
413
,
52
64
.
Shaw
D. R.
,
Ali
M.
,
Katuri
K. P.
,
Gralnick
J. A.
,
Reimann
J.
,
Mesman
R.
,
van Niftrik
L.
,
Jetten
M. S. M.
&
Saikaly
P. E.
2020
Extracellular electron transfer-dependent anaerobic oxidation of ammonium by anammox bacteria
.
Nat. Commun.
11
(
1
),
2058
.
Shoener
B. D.
,
Zhong
C.
,
Greiner
A. D.
,
Khunjar
W. O.
,
Hong
P.-Y.
&
Guest
J. S.
2016
Design of anaerobic membrane bioreactors for the valorization of dilute organic carbon waste streams
.
Energy Environ. Sci.
9
(
3
),
1102
1112
.
Sultana
R.
&
Nasrollahi
N.
2018
Evaluation of remote sensing precipitation estimates over Saudi Arabia
.
J. Arid. Environ.
151
,
90
103
.
Sun
R.
,
Xing
D.
,
Jia
J.
,
Liu
Q.
,
Zhou
A.
,
Bai
S.
&
Ren
N.
2014
Optimization of high-solid waste activated sludge concentration for hydrogen production in microbial electrolysis cells and microbial community diversity analysis
.
Int. J. Hydrogen Energy
39
(
35
),
19912
19920
.
SWPC, S.W.P.C.
2019
SWPC 7 Year Statement 2019–2025
.
SWPC, S.W.P.C.
2020
SWPC 7 Year Statement 2020–2026
.
Tang
S. L.
,
Yue
D. P. T.
&
Ku
D. C. C.
2007
Engineering and Costs of Dual Water Supply Systems
.
IWA Publishing
,
London, UK
.
Trimmer
J. T.
,
Cusick
R. D.
&
Guest
J. S.
2017
Amplifying progress toward multiple development goals through resource recovery from sanitation
.
Environ. Sci. Technol.
51
(
18
),
10765
10776
.
UNDP, U.N.D.P.
2022
Annual Report 2021
.
United Nations Development Programme
.
van der Hoek
J. P.
,
Struker
A.
&
de Danschutter
J. E. M.
2015
Amsterdam as a sustainable European metropolis: integration of water, energy and material flows
.
Urban Water J.
14
(
1
),
61
68
.
van der Hoek
J. P.
,
de Fooij
H.
&
Struker
A.
2016
Wastewater as a resource: strategies to recover resources from Amsterdam's wastewater
.
Resour. Conserv. Recycl.
113
,
53
64
.
van Leeuwen
K.
,
de Vries
E.
,
Koop
S.
&
Roest
K.
2018
The energy & raw materials factory: role and potential contribution to the circular economy of the Netherlands
.
Environ. Manage.
61
(
5
),
786
795
.
Wang
A.
,
Sun
D.
,
Cao
G.
,
Wang
H.
,
Ren
N.
,
Wu
W. M.
&
Logan
B. E.
2011
Integrated hydrogen production process from cellulose by combining dark fermentation, microbial fuel cells, and a microbial electrolysis cell
.
Bioresour. Technol.
102
(
5
),
4137
4143
.
Wang
Z.
,
Dunne
A.
,
van Loosdrecht
M. C. M.
&
Saikaly
P. E.
2018
Effect of salt on the metabolism of ‘Candidatus Accumulibacter’ Clade I and II
.
Front. Microbiol.
9
,
479
.
Wei
C. H.
,
Harb
M.
,
Amy
G.
,
Hong
P. Y.
&
Leiknes
T.
2014
Sustainable organic loading rate and energy recovery potential of mesophilic anaerobic membrane bioreactor for municipal wastewater treatment
.
Bioresour. Technol.
166
,
326
334
.
Werner
C. M.
,
Katuri
K. P.
,
Hari
A. R.
,
Chen
W.
,
Lai
Z.
,
Logan
B. E.
,
Amy
G. L.
&
Saikaly
P. E.
2016
Graphene-coated hollow fiber membrane as the cathode in anaerobic electrochemical membrane bioreactors-effect of configuration and applied voltage on performance and membrane fouling
.
Environ. Sci. Technol.
50
(
8
),
4439
4447
.
Xiong
Y.
,
Hozic
D.
,
Goncalves
A. L.
,
Simões
M.
&
Hong
P.-Y.
2018
Increasing tetracycline concentrations on the performance and communities of mixed microalgae-bacteria photo-bioreactors
.
Algal Res.
29
,
249
256
.
Yang
M.
,
Liu
J.
,
Zhang
X.
&
Richardson
S. D.
2015
Comparative toxicity of chlorinated saline and freshwater wastewater effluents to marine organisms
.
Environ. Sci. Technol.
49
(
24
),
14475
14483
.
Zaouri
N.
,
Cheng
H.
,
Khairunnisa
F.
,
Alahmed
A.
,
Blilou
I.
&
Hong
P.-Y.
2021
A type dependent effect of treated wastewater matrix on seed germination and food production
.
Sci. Total Environ.
769
,
144573
.
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