The simplest approach for reducing water content in urine is evaporation. Maurer et al. (2006) introduced a thermoelectric integrated membrane evaporation system (TIMES) that involves pre-treating urine with ozone (or UV) and sulfuric acid. The mixture is heated, circulated through hollow fiber membranes, and then subjected to reduced pressure, causing water to evaporate and resulting in a more concentrated urine.

Within air evaporation systems (AES), pre-treated urine undergoes filtration via a particulate filter before being directed to a wick package. Subsequently, applying hot air facilitates water evaporation from the wick, resulting in nutrient-rich solids. This process encounters notable challenges, including losing nitrogen as ammonia gas and high energy consumption. The elevated temperatures in the system lead to urea hydrolysis and subsequent nitrogen loss. Strategies to mitigate ammonia loss involve urine acidification or non-hydrolyzed urine. Energy recovery from the same system can reduce energy usage in the process (Maurer et al., 2006). In a study by Mayer (2002), a pressure set at 200 mbar and 780°C temperature was used to evaporate non-hydrolyzed urine. The study achieved a tenfold urine volume reduction without crystallization and yielded a viscous liquid containing 9.7% nitrogen by weight (Rößler et al., 2015). Antonini et al. (2012) generated heat through a solar thermal technique to concentrate nutrients in urine. Solar radiation generated 360 g of solid fertilizer from 50 L of fresh urine within 26 days.

Employing evaporation for nutrient recovery from urine is practical. However, it requires inhibiting ammonia loss by incorporating additional acid to retain ammonium ions within the solution. Riechmann et al. (2021) showed that virtually all nitrogen (97%) in the urine will be lost during evaporation without urine stabilization. Simha et al. (2020b), in a field test with base-stabilized urine, achieved nitrogen recovery of 30 ± 6% by utilizing hot air (at 60°C) for evaporation, while Dutta and Vinnerås (2016) achieved 74 and 54% nitrogen recovery using air at 35°C and 60°C.

The energy-intensive nature of this technology also makes it a less sustainable method. However, a solar thermal evaporation technique has been investigated to mitigate this problem (Antonini et al., 2012; Bucholtz et al., 2023). Antonini et al. (2012) overcame the energy-intensive challenge faced by the evaporation method. However, the study observed that 32% of the nitrogen fed into the solar evaporated alongside the condensate as dissolved ammonia/ammonium, resulting in a minimal N concentration of 1.84% in the dried matter. When acidified, nitrogen and phosphorus concentrations in the dried urine increased to 8.33 and 17.6%, respectively. Furthermore, the solid matter from this process contained 90% sodium chloride, typical of products recovered from urine through evaporation. In a laboratory scale experiment, 97 ± 2% of the nitrogen in stabilized urine was recovered via the dehydration process with dynamic ventilation on trays within four days (Antonini et al., 2012), and similar results were recorded in a field trial (Riechmann et al., 2021). The study also achieved 91 to 96% phosphate recovery efficiency in the urine stabilization reactor as Ca(OH)₂ precipitated P out of the system. The authors also report that the process leads to an inactivation of the pathogens present, making the product safe for use. No helminths or fecal bacteria were detected in the end product. However, one embryonic Ascaris suum egg was noticed in a 500 mL sample of the stabilization reactor. Senecal and Vinnerås (2017) established comprehensive bacteria and virus deactivation within four days of urine stabilization with wood ash at a pH above 10.5 and a temperature of 20°C.

Nutrient recovery through evaporation has also been plagued with micropollutants. Simha and Ganesapillai (2017) noted that therapeutic groups: non-steroidal anti-inflammatory drugs (ibuprofen, naproxen), b-blockers (bisoprolol, metoprolol), analgesics (acetaminophen), CNS stimulants (caffeine), glucocorticosteroids (hydrocortisone), a decongestant (xylometazoline) and two antibiotics (tetracycline, doxycycline) were detected in the dried fertilizer obtained.

A better approach to the urine evaporation process has been developed by adding ash and lime and drying at low temperatures to minimize nutrient loss, reduce energy consumption, and enhance sustainability (Dutta and Vinnerås, 2016; Senecal and Vinnerås, 2017; Simha and Ganesapillai, 2017). Despite the advantages of this approach, a study observed that after alkaline dehydration treatment, pharmaceuticals such as ibuprofen, caffeine, bisoprolol, metoprolol xylometazoline and naproxen were discovered in the recovered dried matter (Simha et al., 2020b). Using this approach, Simha et al. (2020a) treated 35, 62, and 90 L of urine and generated 18.9 kg, 22.8 kg, and 29.7 kg of total solids from the urine in March, April and May, respectively. The fertilizer value of the recovered solids was reported to be 13-2.3-6.0 NPK, respectively, as % of total solids. The system’s total energy consumption was estimated to be 1,913, 2,315, and 1,185 kWh for each of the three months the process was studied. Riechmann et al. (2021) showed that the minimum Ca(OH)₂ consumption in urine stabilization at full capacity was 6 g/L, and the electricity demand was 150 Wh/kg of water evaporated from urine. This study reported an operational cost of 0.05 EUR•pers.⁻¹•d⁻¹.

Based on the reviewed studies, nutrients recovered from human urine via evaporation contain high levels of pharmaceuticals, pathogens, and salts. A long-term application of this product can result in the buildup of pharmaceuticals in the soil and the eventual uptake by plants. The high salts in these products can cause soil sodification, which is detrimental to soil and plant health.

When ice crystals develop, the water molecules’ structure is re-arranged to a stable and steady tetrahedral network. Ions remain concentrated in the liquid fraction of aqueous salt solutions during the gradual formation of ice and are eliminated from the ice crystals as they grow. As a result of the crystal generation processes, ice is almost salt-free. The faster the ice’s crystallization velocity, the greater the concentrations of ions in the water crystal structure as contaminants (Moriwaki et al., 2020). Because of the density variance between ice and salt, both solids can be easily separated by gravity (Moriwaki et al., 2020). This technology is employed in eutectic freeze crystallization in industrial operations and wastewater purification as a low-energy and effective purification technique (Guessous et al., 2023; Hu et al., 2023).

Lind et al. (2001) observed that by subjecting urine to a temperature of −14°C, it is possible to concentrate around 80% of the nutrients in just 25% of the original volume. This process was applied by Ganrot et al. (2007) to investigate nutrient concentration and urine bulk reduction. The study achieved approximately 80% of nutrient recovery from human urine (Ganrot et al., 2007). In cold climate regions, freeze-thawing is an efficient and economical technique for nutrient recovery from urine (Gay et al., 2003). However, Ganrot et al. (2007) opined that freeze-thawing dramatically increased Daphnia magna survivability. Vinnerås et al. (2008) affirmed that at any temperature below 20°C, human urine poses an elevated risk of carrying live viruses and bacteria. This assertion, therefore, makes freeze-thawing urine to recover nutrients highly susceptible to pathogen contamination.

As stated earlier, in freeze-thawing, all ions in urine stay concentrated at the liquid phase and are omitted from ice crystal development. Consequently, micropollutants, salts, and other unwanted contaminants are concentrated alongside the nutrients, making the recovered nutrient material potentially unsafe for agricultural purposes.

In a recent study, an improved freeze-thawing method was accessed by Courtney and Randall (2023). The study affirmed that the reverse osmosis/eutectic freeze crystallization (RO-EFC) method used could potentially eliminate unwanted salts and excess water, resulting in a pure fertilizer output. The energy requirement of this method (60 kWh m³) was significantly less than that of other urine concentration methods. However, the ion concentration of the solid fertilizer obtained in this study contains 31.2% Na by mass.

Although the method looks attractive, fewer studies have been conducted on its use for nutrient recovery from urine. No data on the pathogen load and pharmaceutical pollution of products generated through this process are available. Given the relatively low-cost potential of this approach in cold areas, this could be used successfully in combination with a second process to reduce pathogen and salt risks.

Under pressure, water permeates membranes with tiny pores and significant parts of salts or organic compounds are retained on the membrane. Recently, a gas-permeable hydrophobic membrane (GPHM) has been used to recover ammonia from various liquid wastes (Kuntke et al., 2016; Vanotti et al., 2017).

The primary membrane techniques for nutrient recovery from wastewater include forward osmosis (FO), reverse osmosis (RO), membrane distillation (MD) and electrodialysis (ED) (Ansari et al., 2017). The effectiveness of these technologies is seen in their ability to reject micro-pollutants, trace elements, pharmaceuticals, and nutrients, resulting in less contaminated water.

In FO, ordinary osmotic pressure is used to power the drive of water molecules from the feed side across a water-permeable membrane to the draw side. Thus, nutrient fractions are concentrated at the feed side (Ansari et al., 2017). This process prevents larger molecules from passing through the membrane, decreasing the supply stream’s overall volume, and the draw solution absorbs the water. FO has the advantage of having a lower fouling tendency and a lower energy requirement, especially when there is no need to renew the draw solution (De Vries et al., 2020).

In urine treatment tests, several FO membranes, such as the HTI cellulose triacetate (CTA) or the Porifera TFC-based membrane, as well as a commercially accessible FO system, have been shown to have significant urea rejection rates (Liu et al., 2016). However, the HTI cellulose triacetate (CTA) demanded nearly 48 h to accomplish 80% water retrieval for 1.5 L urine, making the verified technology unrealistic for large-scale nutrient recovery (Borer et al., 2014). Volpin et al. (2018) investigated the concurrent nitrogen and phosphorous reclamation from source-separated urine using the FO process. The study revealed that after the urine was concentrated to 60%, roughly 40% of the P and 50% of the N in the urine were retrieved.

Engelhardt et al. (2020) investigated ammonium and urea rejection using an aquaporin-based hollow fiber membrane based on disparities in variables such as initial feed solution concentration and pH. The study found that ammonium rejection was much higher at highly acidic feed solution (pH of 3), resulting in 90 and 80% ammonium and water recovery, respectively, while at higher pH, ammonium recovery was relatively lower (37%). The study ascribed the upsurge in ammonium recovery at lower pH to the pH sensitivity of the aquaporin-based hollow fiber membrane due to carboxyl functional groups present on the surface of the membrane, containing fewer charges under acidic pH (Liu et al., 2016). The resulting drop in negative surface charge diminished the bi-directional cation exchanges across the membrane, thereby increasing solute rejection. This study concluded that the small molecular size of urea and uncharged structure obscures urea rejection by aquaporin-based FO membranes. It has been posited that the efficiency of N, NH₄-N and total organic carbon rejection using the FO process is attained when the urine (feed) pH is maintained below 7.5. Ray et al. (2019) used acetic acid, whereas Volpin et al. (2018) used citric acid to maintain an acidic fresh urine pH for better FO efficiency.

A new study investigated urea hydrolysis with pH modification as a pre-FO approach to enhance urea rejection. The survey showed urea hydrolysis and pH modification before FO can significantly improve total nitrogen recovery. For instance, 89% of total nitrogen was recovered, and 75% water recovery was achieved using a 5 M HCl draw solution (Engelhardt et al., 2020). The major setback of this study is the use of high amounts of HCl for pH modification, making the system unrealistic for real-world applications. Despite the possibilities presented by nutrient recovery through FO, membrane efficiency is hampered by the water permeability – solute selectivity balance, an essential property controlling water solute passage via polymeric membranes. A highly selective membrane filter effectively concentrates ammonium and phosphate, leading to a concentrated stream with high nutrient content in nutrient recovery procedures. Jin et al. (2012) examined the diffusion of four medicines across commercial forward osmosis (FO) membranes composed of cellulose triacetate (CTA). The study found that all medications studied were rejected with a high efficiency (>94%). It suggested that the rejection is mainly influenced by size exclusion and the hydrophobic contact between the chemicals and the membrane, particularly under acidic circumstances (Jin et al., 2012).

This study indicates a high likelihood that recovered nutrients through FO will still contain hormones, pharmaceuticals and sodium chloride as the membrane only allows water passage, thereby retaining the micropollutant in the concentrated urine (Yip and Elimelech, 2011). A long-term application of such products will be detrimental to soil and plant health. Based on our search, the pathogen load in urine concentrated through forward osmosis has not been investigated.

While FO does not require high-pressure instruments and resources compared to RO, the cost of the membrane per unit area is greater than that of RO. According to Osipi et al. (2020) the cost of FO membranes accounts for approximately 30% of the total capital expenditure.

On the other hand, RO operates by utilizing the hydraulic pressure that opposes the osmotic pressure across the inputs and draws solutions. During this procedure, water molecules move from a solution with a low solute concentration to a solution with a high solute concentration. When pressure is exerted, a semi-permeable membrane separates the urine and pure water solutions, causing the water to flow back into the urine. The exerted pressure compels the water to transition from the solution with higher concentration to the less concentrated medium, specifically from urine to clean water. Consequently, the impurities accumulate on one particular surface of the semi-permeable membrane, while the cleaned water is located on the opposite side (Ray et al., 2019).

Ammonium retention in reverse osmosis membranes performs better than in neutral form (ammonia). Hence, the holding performance significantly impacts the pH of the medium (Maurer et al., 2006).

In a study by Dalhammar (1997), the pH of urine was kept at pH 7.1 to avoid ammonia gas loss. When a pressure of 50 bar was applied, a high concentration resulted in recoveries of 73% for phosphate, 71% for potassium and 70% for ammonium in the retentate was achieved. Parallel outcomes were recorded by Thörneby et al. (1999) in decreasing the water content in liquid animal manure with reverse osmosis. The nutrient retention was excellent (98%) but slightly lower for ammonia at 93%. The energy usage of the process hinged on the technical and operational parameters (Thörneby et al., 1999).

Reverse osmosis (RO) has a lower energy consumption than thermal technologies (Osipi et al., 2020). Still, the process requires high-pressure containers and instruments, which can negate its economic benefits when dealing with a higher feed salinity, such as urine (Valladares Linares et al., 2016). The major setback of this process is that RO membranes retain micropollutants alongside the nutrient fraction of interest (Hofman et al., 1997). The process presents no separation between nutrients and other unwanted pollutants such as pathogens, hormones and pharmaceuticals. Hence, the utilization of this product on agricultural lands can potentially result in a negative impact on soil and crop health or would require a second treatment procedure.

Another limiting factor for this process application is the fouling of the RO membrane because of the precipitation of salts (Migliorini and Luzzo, 2004). Fouling of the RO membrane will reduce the water outflow and upsurge the system’s total cost. Fouling of RO membrane is very severe in the nutrient recovery process from urine as the urine dynamics is affected by pH change. While optimizing to regulate fouling, chemical formulations may be introduced using acids, surfactants, or blend reagents for urine pre-treatment (Jaffer, 1994; Al-Rammah, 2000). Kuntke et al. (2016) recovered 95% ammonia from urine using a gas-permeable hydrophobic tubular system within 11–15 days.

FO seems more desirable than the pressure-driven reverse osmosis (RO) separation, as it significantly lowers the operation costs.

Membrane distillation (MD) is a separation technology that relies on a hydrophobic mesoporous membrane and temperature increase to concentrate solutions under low-pressure conditions (Ashoor et al., 2016). In this process, the liquid is incapable of passing through the membrane pores as a result of the hydrophobicity of the mesoporous membrane. The variation in the partial vapor pressure initiates the conveyance of water vapor through the membrane pores. It has been noted that because water is transported exclusively in a gaseous state through the membrane, MD has been shown to provide complete elimination (100% rejection) of all non-volatile components in the solution being treated (Ashoor et al., 2016).

Since water vapor movement across the MD membrane cannot substantially influence the feed osmotic pressure, the MD process can achieve high water recovery (Xie et al., 2014). In this process, ammonia transfer through the membrane is primarily affected by temperature. Increased temperature will increase ammonia transfer across the membrane. To reduce the exact ammonia movement across the hydrophobic membrane, it is crucial to acidify the source-separated human urine (Tun et al., 2016). In a study conducted by Tun et al. (2016), PTFE (polytetrafluoroethylene)/PP (polypropylene), PTFE/PP and PVDF (polyvinylidene fluoride) membranes were used for nutrient concentration from source-separated urine. The PTFE/PP membrane gave the maximum water flow rate of 60 L m⁻² h⁻¹ and > 75% total nitrogen rejection at 70°C. Xu et al. (2019) developed direct contact membrane distillation technology using hollow fiber polypropylene as a membrane. Using a temperature gradient of 20°C, 3.6–5.0 L m⁻² h⁻¹ water flux was achieved with percentage TN, P, and K rejection being 31, >97, and > 97%, respectively. Higher rejection of ammonium ions (>90%), ammonia (>95%), sodium (>98%) and potassium (>89%) were achieved by using MD hybrid membrane fabricated from polyvinyl di-fluoride and polytetrafluoroethylene ethylene integrated with methyl functionalized silica nanoparticles (Khumalo et al., 2019). MD has been successfully used for water and reclamation from urine via urine concentration as one solution (Khumalo et al., 2019; Xu et al., 2019).

MD can recover volatile ammonia from hydrolyzed urine. However, the process is hindered by the simultaneous absorption of water vapor, leading to low selectivity for ammonia transport. Pre-treatments for source-separated urine are crucial to prolong the MD membrane’s lifespan as micropollutants stick to the membrane’s surface, invariably block the membrane’s pores and reduce the quality and influx of water (Tun et al., 2016; Khumalo et al., 2019). The fate of organic contaminants is hinged on the membrane selectivity, pollutant properties during membrane separation, and their volatility during vaporization (Alturki et al., 2013; Wijekoon et al., 2014). The process is also plagued with high energy demand, which makes the process less environmentally sustainable.

In the electrodialysis (ED) process, ion-exchange membranes are arranged interchangeably in a direct current field to selectively recover nutrient fractions from wastewater (Li et al., 2020). ED utilizes electrochemical redox processes to produce an electric field and facilitate the movement of ions in solution via ion exchange membranes. In electrodialysis (ED), ion segregation is achieved by employing ion-exchange membranes consisting of cation-selective, anion-selective, and bipolar membranes. Throughout the process, positively charged cationic and negatively charged anionic species move towards the cathode and anode, respectively, under the influence of a direct current acting as the driving force. The selective ion membrane hinders co-ion movement (anions and cations). This is achieved through Donnan’s repulsion (Hu et al., 2018).

Selective phosphate enrichment by ED from urine has been carried out with high recovery efficiency (Escher et al., 2006; Pronk et al., 2007). In a study conducted by Li et al. (2020), an innovative bipolar membrane electrodialysis with membrane contactor (BMED–MC) efficiency for ammonia recovery from synthetic urine with 3,772 mg N L⁻¹ concentration was studied. In the BMED–MC system, electrically induced water dissociation in the bipolar membrane concurrently elevated the pH of the urine stream, providing an acid stream for ammonia stripping. The ammonia gas migrated through the gas-permeable membrane in the membrane reactor as the urine pH increased, and it was collected from the acid stream as ammonium sulphate. The study achieved 90% ammonia recovery with small energy consumption (i.e., 12.5 MJ kg–1 N) using the BMED–MC process.

Nutrient recovered using ED has been observed to be of high purity due to the selective nature of the membrane used in the process. Escher et al. (2006) affirmed that the electrodialysis process was highly efficient concerning toxicity reduction. In their study, urine was spiked with micropollutants containing ethinylestradiol and other pharmaceuticals before ED to ascertain the purity of products recovered from the medium. The study observed that 99.9% of the pharmaceuticals and hormones were removed (Escher et al., 2006).

It can be observed that none of the membrane concentration methods discussed so far is free from challenges. Despite high purity and various products obtained using the ED method, the system is plagued with membrane fouling. Due to permanent fouling, the build-up of fouling deposits in the ED process alters the membranes, increases cell resistance (current drop), and lessens ion migration yield and the ion selectivity of the membranes (Mondor et al., 2008).

Combining different membrane systems to form hybrid systems increases total nutrient recovery efficiency (Xie et al., 2014). For instance, the FO process requires a concentration of diluted wastewater. By coupling forward osmosis (FO) with additional membrane techniques like reverse osmosis (RO) or membrane distillation (MD), it is possible to restore the driving force of FO and simultaneously generate high-quality freshwater8 (Xie et al., 2013, 2014).

Hybrid membrane processes like the FO-RO hybrid system (Coday et al., 2015), FO-MD hybrid system (Xie et al., 2013, 2014), ED-RO hybrid process (Mondor et al., 2008), and ED-MD hybrid process have been used to recover nutrients from various wastewaters (Udert and Wächter, 2012). Recently, FO-MD, as a combined process, has been used to improve nutrient recovery from urine (Volpin et al., 2018). Ray et al. (2019) utilized the FO-MD combination for nutrient recovery from urine. The integrated procedure yielded a product solution with a mean urea concentration of 45–68% while achieving over 90% rejection of TOC. Hybrid membranes such as pressure-driven nanofiltration coupled with reverse osmosis are not often utilized due to inadequate urea and ammonia rejection with high capital and cost of operation of the process (Ray et al., 2019).

The ammonia stripping technique revolves around mass transfer, in which wastewater is exposed to air to strip the ammonia gas present. Ammonia is present in aqueous systems in two states: ammonium ions and ammonia gas. These species’ concentrations depend on the medium pH and temperature (Kinidi et al., 2018). The generation of ammonia gas is enhanced through increased pH, resulting in the movement of the chemical equilibrium to the right, consequently improving the formation of ammonia gas. Elevating the pH of the medium before ammonia stripping is, therefore, critical in enhancing the appearance of molecular ammonia nitrogen for stripping (Figure 2). Lime is often used to raise the pH levels of the medium prior to ammonia stripping (Markou et al., 2017); however, Hidalgo et al. (2016) posited that the extreme rise of pH needed in this process imposes an extra cost for lime and is unprofitable. Hence, it is crucial to maintain an optimal pH level in order to achieve a balance between process efficiency and cost-effectiveness. The study observed that above pH 10.5, the ammonia recovery efficiency was inconsequential as increased pH no longer affects the ionization equilibrium between molecular ammonia and ionic ammonium. Nonetheless, the cost incurred increased due to further lime use to improve the pH value (Hidalgo et al., 2016). Also, Markou et al. (2017) discovered that the type of alkali used (calcium hydroxide, sodium hydroxide and potassium hydroxide) had little or no effect on the ammonia removal efficiency (Markou et al., 2017). Nevertheless, calcium alkali has been found more desirable due to heavy metal concentrations, reduction of solids, and wastewater color after treatment (Guštin and Marinšek-Logar, 2011).

The air-water ratio in the ammonia stripping process has also been observed to be a parameter of interest to achieve better removal efficiency. Airflow into the system affects ammonia concentration in the liquid and air phases (Bonmati and Flotats, 2003). Lei et al. (2007) discovered that a higher ammonia recovery rate was attained after 12 h with an airflow rate of 10 L/min, in contrast with airflow rates of 3 L/min and 5 L/min. However, the study resolved that 5 L/min for 1 L of anaerobic effluent is more economically feasible given that an airflow rate of 10 L/min was found too expensive (Lienert et al., 2007).

Traditional ammonia stripping procedures are usually carried out at room temperature and 50°C, with a pH range of 10–12 (Bonmati and Flotats, 2003; Laureni et al., 2013). A high amount of primary reagent is frequently used to boost the pH value of the medium during ammonia stripping. The pH of the system must be adjusted back to neutral for safe discharge after ammonia removal, making the procedure uneconomical. It has been suggested that increasing the process temperature (high-temperature stripping) will increase ammonia removal efficiency and minimize the amount of reagents necessary to increase the pH (De La Rubia et al., 2010). The higher ammonia removal at higher temperatures is due to the solubility of ammonia in water governed by Henry’s law, which states that the constant of gas depends on solute, solvent, and temperature (Kinidi et al., 2018). Saracco and Genon (1994) discovered that the principal cost of an ammonia stripper at 80°C stripping temperature was half that of a stripping temperature of 40°C. However, increased temperature may result in an impractical rise in the cost of preheating.

The high energy consumption in high-temperature stripping can be minimized by employing wasted heat. For instance, the hydrothermal carbonization (HTC) reaction is performed at 180–260°C (Limoli et al., 2016), and it is possible to raise the temperature of the HTC wastewater to 70–80°C by using HTC heat.

García-González et al. (2015) studied the process of ammonia stripping, which involved raising the pH level of urine by using Ca(OH)₂, which causes the precipitation of P and, at the same time, changes NH₄⁺-N into ammonia gas. The gas was transmitted over a hydrophobic membrane that allowed the passage of gases (gas-permeable hydrophobic membrane, GPHM) and underwent a reaction with H₂SO₄ to generate ammonium sulphate. The recuperated product is a solution of liquid ammonium salt [as described in Equations (2.1, 2.2)]. Both N and P are concurrently retrieved.

Pradhan et al. implemented ammonia stripping and phosphorus precipitation (P) to enhance nutrient reclamation from human urine. The method incorporated Ca(OH)₂ to elevate urine pH, facilitating ammonium conversion into ammonia gas and induced phosphorus precipitation as a calcium-phosphate compound. The investigation achieved notable recovery rates for NH₄-N, ranging from 85 to 99%, and accomplished a 99% phosphorus recovery from human urine (Saracco and Genon, 1994).

In a subsequent study, Pradhan et al. utilized ammonia stripping to extract nutrients from human urine. Their findings revealed that over 8 h at a temperature of 30°C, more than 98% (w/w) of nitrogen (N) and phosphorus (P) were successfully extracted. The resulting ammonium sulphate (liquid) harvested from the process exhibited a nitrogen (N) content of 19% (w/w), while the sediments (solids) contained phosphorus (P) at levels of 1–2% (w/w) (Christiaens et al., 2019; Pradhan et al., 2019).

Evolving methods challenging traditional ammonia stripping are membrane stripping (Christiaens et al., 2019; Pradhan et al., 2019) and vacuum thermal stripping methods (Tao et al., 2019; Tian et al., 2019). These emerging technologies require no air stream for ammonia transfer, an advantage that simplifies the procedural arrangement. Christiaens et al. (2019) evaluated this new method with conventional stripping. They discovered that direct liquid–liquid passage from a concentrated urine solution across a hydrophobic gas membrane could reduce or halve the energy needed for similar air-stripping outcomes.

While conventional ammonia stripping is chemical and energy-intensive, combining electrodialysis and membrane stripping with electrochemical stripping can be a more viable tactic (Tarpeh et al., 2018). Nevertheless, chlorination by-products such as ClO₃⁻and ClO₄⁻ stand as a noteworthy setback of the mediated electro-oxidation of urine because of the high chlorine concentration in the medium (Zöllig et al., 2015).

Pathogens, pharmaceuticals and hormones can be predicted to remain in the urine from which ammonia has been released. This suggests that nutrients obtained via ammonia stripping are generally safe for soil and plant health.

Tables 2, 3 illustrate the assessment of diverse approaches employed for nutrient recovery from human urine.

Ion exchange and adsorption are among the first techniques used for nutrient recovery from domestic effluents in tertiary treatment (Yang et al., 2017). Activated charcoal and biochars showed a capacity for PO₄³⁻ adsorption (Maroušek et al., 2020; Peng et al., 2021) urea adsorption (Kameda et al., 2017), and NH₄⁺ recovery (Cai et al., 2016) due to their microporosity, adsorption capacity, specific surface area and ion exchange capacity. Pyrolysis temperature affects the material’s hydrophobicity, carbon content, microporosity, aromaticity, and surface area. Biochar often contains positively and negatively charged surfaces (zwitterionic) (Windeatt et al., 2014). The anionic functional groups play a crucial role in determining the material’s cation exchange capacity (CEC), which enhances their ability to adsorb ammonium ions from the solution. In contrast, the O-containing functional groups of the biochar contribute to the material’s anion exchange capacity (AEC), which enhances the recovery of nutrient anions such as phosphate.

Ion exchangers are used because of their notable properties, such as fast uptake kinetics and regeneration, high selectivity for NH₄⁺ and high recovery (Malovanyy et al., 2013; Liao et al., 2015). Nutrient recovery through ion exchange is simple and requires less space for operation. This eco-friendly method uses mainly naturally available and modifiable ion-exchanger/adsorbents, like zeolites, and discharges non-hazardous exchangeable cations (Na⁺, K⁺, Ca²⁺, and Mg ²⁺) (Cai et al., 2016; Yang et al., 2017). Ion exchange processes require regenerant to recover the ion exchange after use. Restoration of ion exchangers is carried out by using highly concentrated brines to enhance the desorption of phosphate and ammonium ions in brine solution, which can be precipitated afterward as nutrient-rich solids appropriate for reuse as fertilizers. The common regenerants used in this process include sodium chloride and potassium chloride, with 68 and 94% regeneration efficiency, respectively (Guida, 2023).

Zeolites, particularly clinoptilolite, have generally been used for ammonia and phosphate removals from human urine (Kavvada et al., 2017; Kabdaşlı and Tünay, 2018).

Numerous studies have been conducted to remove ammonia from human urine through ion exchange techniques. Lind et al. (2001) prepared synthetic urine comprising 0.02–1 M NH₄Cl to imitate source-separated human urine to study the adsorption capacity of wollastonite, clinoptilolite, and zeolites. Clinoptilolite uptake of ammonium ions from low-concentration solutions was 70–80 percent, wollastonite achieved 50 percent recovery, and mixed zeolite uptake was 50–60 percent. Ganrot et al. (2007) studied phosphate and ammonia recovery from human urine using a zeolite containing high levels of clinoptilolite. The experiment was carried out by adding MgO, activated carbon and with/without freeze–thaw technique. Results of the experiments showed that phosphorous was largely recovered as struvite, and an enhanced nitrogen recovery was obtained with zeolite addition. In a continuous flow column experiment, Beler-Baykal et al. (2004) employed human urine to appraise ammonium recovery using ion exchange. A concentrated NaCl solution was used to preheat natural clinoptilolite with a 1–2 mm size range. At loadings of up to 3 mg ammonium/g clinoptilolite, removal efficiencies of over 85% were attained.

With diluted urine and clinoptilolite, Kocatürk and Baykal (2012) recovered up to 86% nitrogen, with more significant reductions at higher concentrations. With higher phosphate concentrations, recovery efficiencies rose, reaching 96%. Using struvite precipitation and zeolite adsorption, Bán and Dave (Ban and Dave, 2004) conducted laboratory investigations on the recovery of phosphorous and nitrogen from source-separated urine. With particle sizes of 1.2–2 mm, the zeolite used had a significant clinoptilolite concentration. The reactive material was evaluated at various concentrations ranging from high (0.5–3 g MgO and 30–960 g zeolite per liter of human urine) to low (0.005–0.5 g MgO and 7.5–120 g zeolite per liter of human urine). Phosphorus elimination resulted from struvite precipitation and zeolite adsorption during the low-concentration test.

Zeolite adsorption was the dominating mechanism for pH levels lower than 8.5. In this low-concentration trial, the zeolite was intended to function as a cation exchanger for the residual ammonium ion after the struvite precipitation process.

For phosphate recovery from urine, greywater, and other waste streams, O’Neal and Boyer (2013) used hybrid anion exchange resins (HAIX) comprising hydrous ferric oxide and got excellent recoveries. Fresh urine > hydrolyzed urine > anaerobic digester supernatant > greywater > biological wastewater had the highest phosphorus loading on HAIX resin. HAIX resin had a higher phosphate sorption capacity for source-separated urine than the others. Xia et al. (2017) prepared highly distributed MgO-nanoflake altered diatomite for phosphate adsorption and achieved a high P recovery capacity of 138 mg/g. Kini (2023) removed phosphorus from human urine using granulated blast furnace slag (GGBS). In this investigation, human urine was diluted (50 percent) with tap water to simulate flushing circumstances. The results from the batch experiments demonstrated that a GGBS dosage of 700 g/L, a contact time of 120 min, and a pH range of 6–9 resulted in a phosphorous removal effectiveness above 90%.

While these ion exchange processes appear promising, the considerable challenges impeding their performance in nutrient recovery encompass insufficient media selectivity, bed blockage, and the high cost associated with regenerations.

In the long run, it can be inferred that recurrent regenerations involving fresh chemical solutions and significant expenses for disposing of expended regenerant as hazardous waste are highly uneconomical (Huang et al., 2021). Therefore, the retrieval of regenerants becomes imperative for the preservation of economic viability.

New studies suggest that the appropriate solutions should aim to clean up and reuse the used-up regenerant as well as recover the nutrients simultaneously. Canellas (2023) employed commercially available hollow fiber membrane contactor (HFMC) modules with sulphuric acid to clean up KCl utilized as a regenerant while also recovering ammonia as ammonium sulphate. The phosphorus present in the medium is recovered as calcium phosphate by adding Ca(OH)₂ to the wasted sodium hydroxide regenerate, which results in the creation of hydroxyapatite Ca₅(PO₄)₃(OH), which can then be filtered out of the liquid.

A review study by Beckinghausen et al. (2020) compared industrial nitrogen fertilizer production with nitrogen recovery from waste streams using different technologies, suggesting that nutrient recovery from wastewater is currently uneconomical. However, the cost of NH₄NO₃ recovery through the ion exchange method was observed to be smaller (£1.98 ± 0.23/kg N) compared to other processes, which cost around £2.35/kg NH₄-N. Nonetheless, this cost is still relatively high compared to an industrial synthesis method (£0.28/kg N) (Neale et al., 2010; Tong et al., 2017).

Studies have shown that nutrient recovery through ion exchange from wastewater containing micropollutants such as pharmaceuticals and hormones may result in problems (Jiang et al., 2015; Tong et al., 2017, 2018). For example, micropollutants such as ibuprofen, estrone, and sulfadiazine adsorb on polymeric ion exchangers (Neale et al., 2010; Jiang et al., 2015). The ability of micropollutants to adhere to ion exchangers can increase the levels of micropollutants in the regeneration brines. This can contaminate the nutrient-rich precipitate obtained from the brine after regenerating the ion exchanger (Neale et al., 2010; Tong et al., 2017). Tong et al. (2017) investigated the fate of the micropollutants triclosan (neutral and anionic species in aqueous medium at neutral pH), 17 β-estradiol (neutral at neutral pH), and sulfamethoxazole (anionic at neutral pH) during nutrient recovery by ion exchange-precipitation. The study found that triclosan, 17 β-estradiol, and sulfamethoxazole adsorption to the phosphate-selective ion exchange resins LayneRT and DOW-HFO-Cu ranged from 50 to 71% when employing anaerobic effluent but did not adsorb to the ammonium-selective exchanger clinoptilolite. The hydrophobicity of the micropollutant was also reported to increase the interaction between the resins and the micropollutants.

Charcoal and ion exchange resins (natural and synthetic) remove nutrients from aqueous solutions and eliminate some pathogens (Busscher et al., 2006) and organic contaminants (Nam et al., 2014; Tong et al., 2017). Based on these studies, nutrient recovered from urine via ion exchange is susceptible to pathogens, pharmaceuticals and hormone pollution; thus, micropollutant removal from urine before nutrient recovery through ion exchange or adsorption process is needed to mitigate the contaminants’ risks to the environment. Parsons (2000) suggested that there could be a potential accumulation of bacteria on ion exchange resins, implying that the use of ion exchange for nutrient recovery might introduce concerns related to pathogenic microorganisms.

Chemical precipitation is the most straightforward approach to eliminate excess dissolved phosphorus from wastewater effluent. This is commonly achieved by applying aluminum, calcium, magnesium or iron-rich materials (Tarpeh et al., 2018). The most frequently used reagents for phosphorus precipitation across various wastewater treatment systems are lime (Ca(OH)₂), alum (Al₂(SO₄)₃·18H₂O), and ferric chloride (FeCl₃) (Höglund et al., 2000; Huang et al., 2020).

Nutrients in wastewater have been precipitated in different forms ranging from magnesium ammonium phosphate hexahydrate (MAP) (Lahr et al., 2016; Filho et al., 2017; Zamora et al., 2017), magnesium potassium phosphate (Nakao et al., 2017), aluminum phosphate (Huang et al., 2020), magnesium and sodium phosphates (Huang et al., 2020), calcium phosphate (Saliu et al., 2019) and ferric phosphate (Lin et al., 2017).

The pH and hardness of the medium are the most significant factors often considered in using a chemical precipitation approach for phosphorus recovery. At a pH range of 4.0–6.0, iron phosphate and its oxyhydroxide complexes attain minimum solubility, whereas at a pH range of 5.0–7.0, aluminum phosphate and its oxyhydroxide complexes attain the most negligible solubility. Because this process consumes alkalinity, pH modifications may be necessary, depending on the wastewater influent source. To obtain minimum solubility, calcium apatite (Ca₁₀(PO₄)₆(OH)₂) demands an elevated pH (9.0), which necessitates wastewater softening via the creation of calcium carbonate (CaCO₃). The principal component of the sludge produced by the treatment plant is the creation of CaCO₃, combined with the desired precipitation of calcium apatite (Okano et al., 2013).

In recent years, the struvite precipitation technique has been effectively employed in industrial, municipal, and agricultural wastewater for nutrient recovery (Lahr et al., 2016; Zamora et al., 2017; Figure 3). Struvite is a white crystalline powder characterized by an orthorhombic crystal structure. It comprises equal molar concentrations of magnesium, ammonium, and phosphate (Mg^2+/NH₄⁺/PO₄³: 1/1/1). Depending on the precipitation media, it can also develop as a yellow, brownish, or light grey precipitate (Filho et al., 2017; Kabdaşlı and Tünay, 2018).

The typical reaction for struvite precipitation is given in Equation (2.3) below (Lahr et al., 2016; Kabdaşlı and Tünay, 2018):

It is hypothesized that the primary form of orthophosphates throughout the reaction is HPO₄²⁻ or H₂PO₄⁻ rather than PO₄³⁻ (Jia et al., 2017; Tansel et al., 2018). This means hydrogen ions are discharged into the solution, lowering the medium’s pH during struvite precipitation (Krähenbühl et al., 2016; Kataki et al., 2016; Lorick et al., 2020).

Struvite is formed in a slightly alkaline medium by the reactions listed as Equations (2.4) and (2.5) below (Kumar et al., 2015):

Consequently, the stability of the struvite precipitation process is highly sensitive to pH change. Struvite precipitation has recently been demonstrated to be viable for recovering nitrogen and phosphorus from human urine (Ganrot et al., 2007; Lahr et al., 2016; Filho et al., 2017; Zamora et al., 2017). Ammonium and phosphate in pre-hydrolyzed or fresh human urine are changed into magnesium ammonium phosphate hexahydrate (MAP) (generally recognized as struvite) under the right operating conditions in the struvite precipitation process. Nitrogen, phosphate, and potassium can be reclaimed from human urine in three ways by struvite precipitation. The primary goal of the first strategy is to recover 100% phosphorus and a small amount of ammonia from newly collected human urine. This approach allows for the recovery of around 5% ammonia and 96–98% phosphate (Liu et al., 2013), resulting in effluents with elevated nitrogen concentration in the form of ammonia and urea, requiring an extra treatment step. The other method relies on comprehensive ammonia and phosphate recovery from human urine. Due to the phosphorous deficit in ammonia, this application asks for urea to ammonia conversion by ureolysis and an outside phosphorous supply (Kabdaşlı and Tünay, 2018). An alternate method was recently proposed, which involves precipitating K-struvite (MgKPO₄6·H₂O) from human urine. To reclaim potassium via K-struvite precipitation, total hydrolysis of urea to ammonia is required (Ray et al., 2019; Patel et al., 2020).

In addition to urea hydrolysis and ammonia removal, dilution of human urine is vital in achieving adequate K-struvite precipitation, particularly in incomplete ammonia recovery (NH₄–N > 100 mg N/L) (Xu et al., 2012). Upon meeting these essential conditions and incorporating an external magnesium supply, K-struvite precipitation from human urine, conducted under optimal operational parameters, can recover up to 77% of potassium and nearly all phosphorus. Struvite recovery is feasible exclusively in the presence of Mg²⁺ ion (Figure 3). Magnesium ions from various sources (e.g., Mg(OH)₂, MgCl₂, MgO) have been employed for struvite precipitation from wastewater for many years (Barbosa et al., 2016). Using magnesium supplied from low-cost sources significantly lowered the cost of struvite manufacturing (Luo et al., 2018). Compared to other magnesium sources, such as MgCl₂ and MgSO₄, Hug and Udert (2013) found that using MgO is less expensive. Using MgO or Mg(OH)₂ is also advantageous since they tend to keep the solution pH in a desirable range, promoting struvite crystallization (Hutnik et al., 2013).

Using MgO, Morales et al. (2013) successfully recovered more than 95% of P in the stored urine after 30 min of experimental time. Barbosa et al. (2016) used struvite precipitation to extract nutrients from source-separated urine. The precipitating agents MgO, Mg(OH)₂, and MgCl₂ were evaluated for struvite formation from urine. P recovery efficiency ranged from 82 to 89 percent when MgCl₂ was employed, while P recovery efficiency ranged from 79 to 93 percent when Mg(OH)₂ was utilized. When MgO was used as the Mg²⁺ source, the P recovery efficiency ranged from 90 to 100%. The pH of the solution rose from 8.5 to 9 in the first 3 min and stayed at 9.2 during the stirring period (10 min).

Pinatha (2023) studied nutrient recovery from urine using ashes of calcium and magnesium from power plant solid wastes. The findings indicated that the efficiency of P recovery was enhanced when the pH value increased. Precipitation of P into solid precipitates reached maximum levels of 75, 99, and 99% at pH 7, 9, and 11, respectively. The P content in the form of P crystals was approximately 70, 60, and 50%.

Pinatha et al. (2020) investigated nutrient precipitation from urine with sea salts in a study. Results from the experiments conducted within the pH 7 to 11 revealed that the efficiency of phosphorus precipitation increased to 89–97% with the addition of MgCl₂ and 72–88% with sea salts. The precipitates obtained from both scenarios exhibited a dry-weight phosphorus content ranging from 10.8 to 17.1%, surpassing numerous commercial fertilizers (8.80% P).

The major setback to this technology is that struvite precipitation only allows fractional nitrogen recovery with a non-optimal NPK ratio for plant growth. The process also involves the addition of expensive magnesium supplements whose accumulation in soil could eventually become detrimental to plant growth.

Studies showed that pathogens may accumulate in the precipitate (Decrey et al., 2011; Udert and Wächter, 2012; Lahr et al., 2016) while pharmaceuticals adhere to the surface of deposits instead of being integrated into the crystal structure, which can be eliminated by washing (Escher et al., 2006; Ronteltap et al., 2007; Schürmann et al., 2012). The choice of reagent type in nutrient precipitation from urine significantly impacts process effectiveness and the economic evaluation of the method. The use of large amounts of precipitant in this process offsets the price of the process. To reduce the financial challenges of the technique, it is crucial to pinpoint and appraise the efficacy of less expensive non-traditional reagents (Pinatha et al., 2020).

Microbial fuel cells (MFCs) are bioelectrochemical instruments that degrade organic and inorganic wastes into smaller molecules. Electrons and protons are discharged during the procedure, and electricity is produced. MFCs use microbes or enzymatic catalysis to transform chemical energy directly into electrical energy via bioelectrochemical processes (Deval et al., 2017; Gadkari et al., 2020; Savla et al., 2021).

MFC is a bioelectrochemical device consisting of an anaerobic anode chamber, and an aerobic cathode chamber separated using an anion exchange membrane (AEM). For standard MFC, microorganisms oxidize the substrate in the anode chamber, and the electron generated by the organism is transferred to the cathode through an outer cable. The anode chamber is made up of a microbe (catalyst) and an electrode (anode), nourished with wastewater (anolyte and redox mediator, respectively) (Tamta et al., 2020). Water is formed on the cathode when required protons and electrons are withdrawn during bacterial substrate catabolism and combined with oxygen (Ledezma et al., 2017; Monetti et al., 2019; Jatoi et al., 2021).

Typically, anaerobic bacteria use the anode as the terminal electron acceptor to oxidize organic materials in the wastewater to CO₂ (Ye et al., 2019; Matsuda et al., 2020). As demonstrated in Eqs. (2.6 and 2.7) (Matsuda et al., 2020), the oxidation reactions occurring at the anode are illustrated by acetate or glucose degradation:

To move electrons to the solid electron acceptor (i.e., anode) situated outside the cells, bacteria use extracellular electron transport systems (e.g., direct electron or biofilm matrix). Anaerobic bacteria save the energy required for development and conservation by doing so. The electrons flow to the cathode through an outer electrical circuit and a load (resistor) (Matsuda et al., 2020). The electrons are eventually used at the cathode to reduce electron acceptors like oxygen, nitrate, phosphate, and metal ions. MFCs have been used to recover ammonia nitrogen and electricity from human urine (Tice and Kim, 2014; Kuntke et al., 2016; Tremouli et al., 2018).

In previous studies, MFCs have effectively recovered ammonia nitrogen from diluted urine (Ieropoulos et al., 2012). As a result, using source-separated urine, the concept of ammonia recovery was implemented and confirmed a power output of 0.25Wm⁻² and an ammonia recovery rate of 3.3 g N m²/d in MFCs (Kuntke et al., 2012). A two-chambered BES with a gas diffusion cathode was used for ammonium nitrogen recovery. The movement of the ammonium ion and circulation of ammonia through the cation exchange membrane were the two main methods for transporting ammonia to the cathode chamber (CEM). Because of the system’s alkaline pH, the NH₄⁺ ion is transformed to NH₃ in the cathode chamber (9.5). The catholyte’s ammonia was supposed to be recovered by volatilization and absorption into a dilute sulfuric acid, resulting in liquid ammonium sulphate (Kuntke et al., 2012).

Others have recovered ammonium from urine using the BES in microbial electrolysis cell mode (Kuntke et al., 2014). Ammonium recovery from urine was established at a recovery rate of 162 (±10) g NH₄-N m^_2 d^_1 and a simultaneous H₂ generation in microbial electrolysis cells (MEC), with an energy demand of 8.2 MJ kg^_1 NH₄-N (Kuntke et al., 2014).

Phosphorus removal methods in BES depend on phosphorus solubilization from iron phosphate minerals and soluble phosphorus precipitated as struvite. Moreover, the discharge of iron declines at the cathode, and phosphorus discharge from iron phosphate (Fe-P) crystals improves the electrons obtained from waste organic matter. Direct or electron shuttle within the cathode forces the reduction of iron in the Fe-P. The movement of ammonium ions from the anode to the cathode chamber and alkali generation at the cathode are involved in the precipitation of soluble phosphorus in BES (Zamora et al., 2017; Ye et al., 2019).

In a study conducted by Sharma and Mutnuri (2019), the performance of MFC was assessed by mixing consortia and isolated pure cultures (Firmicutes and Proteobacteria species) from biofilm for nutrient recovery and electricity generation from human urine. The results revealed that microbes only use a fraction of the total phosphorus that is accessible, specifically less than 10%, for their growth, while 90% is recovered as struvite. The study established the concurrent removal of nitrogen and phosphorus from the system as struvite using MFC.

In a recent study, Freguia et al. (2019) studied a self-powered bioelectrochemical nutrient recovery system for nutrient recovery from urine. The investigation demonstrated the world’s first autonomous nutrient recovery mechanism and does not require any chemical or process control. In a microbial fuel cell configuration, a novel air cathode catalyst functional in urine conditions enables in-situ electricity generation and immediate collection of ionic nutrients into a heavy metal-free product stream. Over two months, the system maintained electrical current densities of roughly 3A m–2 while concurrently increasing the concentration of N and K by a factor of 1.5–1.7.

Studies have shown that temperature affects the behavior of the microorganisms used in MFC and the system’s kinetics, activation energy, conductivity of the solution and electrode potentials. Microorganisms have an optimum temperature and pH for growth and metabolism pathways. The pH reduction within the anode impedes the growth of microorganisms that generate electricity. Temperatures between 30–45°C enable biofilm generation and improve the bio-electrocatalytic activity. Heidrich et al. (2014) evaluated the efficiency of MFC at temperatures ranging from 1 to 5°C and discovered that the MFC efficiency (48.7%) was poor and took a long time to attain. However, Jamal et al. (2020) achieved a better result at room temperature.

The financial burden of MFC is tremendously high compared to anaerobic sludge treatment because of the use of expensive electrodes, catalysts, and membranes. Typically, anode and cathode electrodes are costly, and using triggers like Pt increases the inversion cost dramatically, making the process unsuitable for industrial or wastewater treatment (Munoz-Cupa et al., 2021). Despite using carbon-based materials as electrodes, Li et al. (2013) estimated that the cost of MFC operation is roughly 30 times greater than that of conventional wastewater treatment.