Technologies for Recovering Nutrients from Wastewater: A Critical Review

Abstract

Phosphorus (P) is a non-renewable resource, production of nitrogen (N) fertilizer is energy intensive, and discharge of these nutrients in treated wastewater causes environmental eutrophication. Hence, recovery of nutrients from municipal wastewater has attracted attention. In this article, current technologies for such recovery are reviewed, with synthesis in terms of wastewater characteristics, recovery goals, effluent discharge limits, constraints on chemical usage, treatment plant scale, operational complexity and applicability, and analysis of energy demands. Phosphorus recovery processes applicable for centralized plants include enhanced biological phosphorus removal (EBPR) combined with chemical and electrochemical struvite precipitation and chemical precipitation alone, whereas electrochemical and chemical precipitation and ion exchange (IE) may be adapted to onsite and packaged treatment plants. Many processes can be used for N concentration; however, N recovery has been reported only by struvite precipitation and acid absorption following separation by gas stripping or gas permeable membrane. Only chemical and electrochemical precipitation can produce fertilizer requiring minimal post-processing beyond filtration. Electrochemical precipitation of struvite and calcium phosphate is further capable of such recovery with minimal chemical addition. Direct microbiological recovery as protein is an emerging technology, and algal recovery is being developed for livestock and fuel production. Although reactive filtration can achieve very low P discharge concentrations, the only processes reported to be capable individually of removing P in secondary effluent to below 10 μg/L, for example, for discharge to surficial waters, were adsorption and IE. Several authors point to EBPR as a currently preferred approach, and further development of electrochemical processes appears warranted.

Introduction

Phosphorus is a non-renewable resource currently mined for fertilizer, and conversion of nitrogen to ammonia fertilizer by the Haber–Bosch process is energy intensive, accounting for 1% of world's energy consumption and attendant greenhouse gas emissions (Smith, 2002; Galloway and Cowling, 2002; Williams et al., 2015). Much of this nutrient load is taken up in food and ultimately delivered to treatment plants in municipal wastewater which, when collected separately from storm water, may contain ∼5.6 mg/L of total phosphorus, as P, and ∼35 mg/L of total nitrogen, as N, (Tchobanoglous et al., 2014). Conventional wastewater treatment plants generally remove nutrients from wastewater as phosphate-rich sludge and convert nitrogen-containing compounds into nitrogen gas by nitrification–denitrification, to control eutrophication (De-Bashan and Bashan, 2004). However, the embodied energy of nitrogen-containing compounds is then lost in their conversion to stable nitrogen gas, and phosphate in the sludge is relatively dilute and often lost to large water bodies where it is impractical to recover.

The purpose of this article is to review current technologies for recovery of phosphorus, and nitrogen, directly from municipal wastewater. Applicability in terms of wastewater characteristics, recovery goals, effluent discharge limits, constraints on chemical usage, treatment plant scale, and operational complexity are then evaluated and compared. We first consider P-recovery technologies in Phosphorus recovery Section and then recovery of N in Nitrogen recovery Section.

Phosphorus recovery

Williams et al. (2015) indicate that ∼15% of phosphate fertilizer demand can be satisfied by recovering phosphates from wastewater treatment plants. Only 2% and 6% of foodborne nitrogen (N) and phosphorus (P), respectively, are retained in the body during human growth, and there is no net retention once the body is fully grown (Jönsson et al., 2004). Hence, at 1.2 g/(person•day), the global population would excrete 3 million tons of P per year in urine and feces (Cordell et al., 2009, 2012). Recovery of this wasted P is critical considering the diminishing P resources and growing demand for fertilizer.

Ion exchange/Adsorption

Ion exchange (IE) is a process in which PO₄³⁻ is adsorbed on to a P-selective media, exchanging with cations such as Cl⁻ to produce P-depleted effluent while maintaining charge neutrality in the solid media (Williams et al., 2015). Typically these media are porous materials with high surface area and active sites to selectively adsorb ions (Mehta et al., 2015). A variety of commercial media such as Purolite A500P and A520E (Johir et al., 2011), hybrid anion exchangers (HAIXs), Amberlite IRA-410 (Martin et al., 2009), Hydrotalcite compounds (Kuzawa et al., 2006), and laboratory synthesized media, such as DOW-Cu, DOW-FeCu, DOW-NAB-Cu, and DOW-HFO (Sengupta and Pandit, 2011; Williams et al., 2015), have been tested on both synthetic and real wastewater and have demonstrated good P-removal capability (Table 1).

Table 1.

Properties of Different Adsorbent Materials Used in P Recovery

Material	Wastewater type	Influent conc.	Effluent conc.	Max adsorption capacity	EBCT	Desorption solution	Reference
HTAL	Synthetic	200 mg-P/L	—	47.3 mg-P/g	Batch	10 w/v% NaCl +3w/v% NaOH regeneration using	Kuzawa et al. (2006)
HTAL	Synthetic	200 mg-P/L	—	47.3 mg-P/g	Batch	25 w/v% MgCl₂	Kuzawa et al. (2006)
A500P	MBR effluent	3.11–4.18 mg-PO₄/L	0.31–0.42 mg-PO₄/L	—	2.4^c h	1 (w/v)% NaCl (20 BVs)	Johir et al. (2011)
Dow-Cu	Synthetic	100 mg-PO₄/L	<0.1 mg-PO₄/L	0.5 g-P/L_media	17 min	6% NaCl, pH = 4.3	Williams et al. (2015)
Dow-FeCu	Synthetic	100 mg-PO₄/L	<0.1 mg-PO₄/L	1.5 g-P/L_media	17 min	6% NaCl, pH = 8	Williams et al. (2015)
LayneRT™	Synthetic	100 mg-PO₄/L	<0.1 mg-PO₄/L	2.0 g-P/L_media	17 min	2% NaCl +2% NaOH, pH >14	Williams et al. (2015)
HAIX	Synthetic	200 μg-P/L	<10 μg-P/L	2 mg-P/g-HAIX	2.2 min	2% NaOH +2% NaCl	Blaney et al. (2007)
HAIX	Secondary effluent	260 μg-P/L	<10 μg-P/L	—	2.2 min	2% NaOH +2% NaCl	Blaney et al. (2007)
Zr-FPS	Synthetic	400 μg-P/L	4 μg-P/L	1.45^c mg-P/g	1.2^c min	0.1 M NaOH, regeneration with 0.5 M H₂SO₄	Awual et al. (2011)

calculated value.

EBCT, empty bed contact time; FPS, bifunctional cation exchange fiber; HAIX, hybrid anion exchanger; HTAL, hydrotalcite compounds.

Many studies provide evidence of P removal in both synthetic and actual wastewater over wide ranges of P concentration (1–2,000 mg/L (Mehta et al., 2015)), including very low concentrations (Sengupta and Pandit, 2011). The process is suitable for removing trace P concentrations from effluents (Blaney et al., 2007) with low solid concentration (<2,000 mg/L) (Kuzawa et al., 2006; Sengupta and Pandit, 2011; Mehta et al., 2015). Williams et al. (2015) have demonstrated PO₄³⁻ removal from 100 mg/L to <0.1 mg/L in synthetic wastewater. Johir et al. (2011) showed that over 95–97% P can be recovered from real membrane bioreactor (MBR) effluent with a P concentration of 3.11–4.18 mg/L. Awual et al. (2011) and Blaney et al. (2007) demonstrated P removal down to 10 ppb, as recommended for discharges to South Florida waters to protect against eutrophication (Florida Everglades Forever Act, 1994), using zirconium(IV)-loaded fibrous adsorbent and HAIX media, starting from 216 and 260 μg/L, respectively.

The number of bed volumes (BVs) before breakthrough depends on several factors, dominated by the typically limited selectivity of the media, species of competing anions (Puchongkawarin et al., 2015), and their influent concentration, as well as on PO₄³⁻ concentration. Sengupta and Pandit (2011) observed that SO₄²⁻ influenced the adsorption capacity of HAIX media at high concentrations (160 mg/L), and the influence depends on the concentration of both P and the competing anion. According to Martin et al. (2009), the type of influent had a significant influence on the number of BVs before breakthrough of HAIX media. Studies indicate that the adsorption capacity of a media is limited by the number of active sites; hence, media regeneration is an important consideration in minimizing cost of operation and waste generation.

Media regeneration is often carried out using a comparatively small volume of brine or alkaline solution, or a combination thereof. Typical desorption solutions (regenerants) include NaCl (1–6%), MgCl₂, and mixtures of NaCl and NaOH (Martin et al., 2009; Johir et al., 2011; Sengupta and Pandit, 2011; Williams et al., 2015). Regenerant volumes can range from one (Martin et al., 2009) to 20 (Sengupta and Pandit, 2011) BVs, depending on the media. However, the amount of P desorbed depends on the volume of regenerant used (Martin et al., 2009; Sengupta and Pandit, 2011). According to Martin et al. (2009), 20% of P remained in HAIX media after one BV, while about 10 BVs would be required to desorb almost all of the P in the media. In fact it is thermodynamically impossible to desorb all the adsorbed P, and in addition, adsorption capacity is impaired upon media regeneration. For example, Kuzawa et al. (2006) observed 70% decrease compared to initial adsorption capacity after three regenerations. More recent studies indicate minor loss (∼1.5%) in adsorption capacity even after 10 cycles (Martin et al., 2009; Sengupta and Pandit, 2011), although only ∼90% of P was desorbed (Sengupta and Pandit, 2011). Furthermore, the capacity for desorption of P strongly depends on the type of regenerant and the regeneration process. Regeneration can be either a one-step (using one solution) or two-step (using two solutions) process. One-step regeneration has been found applicable in large-scale applications due to reduced chemical cost, foot print, and operational complexity. Cost can be further reduced by reusing regenerant solution following the recovery of desorbed P.

Extraction of P from the regenerant solution is very important from a nutrient recovery perspective, as well as for cost reduction. Although some authors suggest direct use of brine solution as fertilizer, P is commonly recovered from used regenerant by precipitation as struvite or calcium phosphates (Sengupta and Pandit, 2011). Precipitation is carried out by addition of a Ca salt [e.g., Ca(NO₃)₂] or a mixture of MgSO₄·7H₂O and NH₄Cl. Even though precipitation can produce high-purity fertilizer with HAIX media (Sengupta and Pandit, 2011), leaching of metal ion (e.g., Al³⁺ and Ca²⁺) can reduce the product purity (Williams et al., 2015). Furthermore, reuse of spent regenerant following PO₄³⁻ precipitation would require the addition of NaOH to compensate for OH⁻ losses, and loss of performance of recycled regenerant is observed due to trace of P concentrations after precipitation.

Magnetic microsorbents

P removal and recovery using magnetic microsorbents is based on similar principles as adsorption; only magnetic microsorbents are adsorbent materials with magnetic properties suspended in wastewater. Nutrients adsorbed on magnetic media are recovered by capturing the suspended media using high-gradient magnetic separators, with subsequent regeneration and precipitation from regenerant (Mehta et al., 2015).

Various types of magnetic microsorbents were tested on synthetic and actual wastewater to remove trace P concentrations of ∼1 to 0.03–0.005 mg/L PO₄-P (Ito et al., 2009; Ishiwata et al., 2010; Drenkova-tuhtan et al., 2017b). Reported magnetic microsorbent materials include ferromagnetic zirconium ferrite (ZrFe₂(OH)₈) (Ishiwata et al., 2010), carbonyl iron particles (Merino-Martos et al., 2011), magnetite and iron particles (de Vicente et al., 2010), and Fe₃O₄ nanoparticles embedded in SiO₂ matrix coated with P selective ZnFeZr (Drenkova-tuhtan et al., 2017b) (Table 2). Ding et al. (2012) indicated that magnetic IE resins which are typically used for removal of organic compounds can be also used for P removal from secondary effluent. Similar to conventional adsorption, most of these materials were able to reduce P levels to below 0.05 mg/L P_total (<0.005 mg/L PO₄-P) (Drenkova-tuhtan et al., 2017b). However, the absorbent dosage had to be increased and adsorbent material had to be washed in between cycles to achieve >90% removal (Drenkova-Tuhtan et al., 2017a).

Table 2.

Properties of Different Magnetic Microsorbents Used in P Recovery

Material	Wastewater type	Influent conc. (mg/L)	Effluent conc. (mg/L)	Max adsorption capacity (mg/g)	Recommended adsorption capacity (mg/g)	Contact time (min)	Reference
Zirconium ferrite	Treated wastewater	1.12^a	0.03^a	—	—	5	Ishiwata et al. (2010)
Zirconium ferrite	Treated wastewater	1.12^a	0.01^a	—	—	15	Ishiwata et al. (2010)
MgFeeZr LDH	Spiked treated wastewater	10^b	—	35^b	—	—	Drenkova-Tuhtan et al. (2013)
ZnFeZr-adsorbent	Spiked treated wastewater	∼10^b	0.05^b	—	17^a	20	Drenkova-Tuhtan et al. (2017)
ZnFeZr	Treated wastewater	0.08–1.9^c	0.05^c	75^a	10^a	20	Drenkova-tuhtan et al. (2017)
Magnetite	Synthetic wastewater	155–1,860^a	—	27.15^a	—	Eq^m	de Vicente et al. (2010)
Iron	Synthetic wastewater	155–1,860^a	—	18.83^a	—	Eq^m	de Vicente et al. (2010)
Iron-ATPS	Synthetic wastewater	155–1,860^a	—	25.06^a	—	Eq^m	de Vicente et al. (2010)

as P.

P-PO₄³⁻

calculated value.

LDH, layered double hydroxide; ATPS, surface treated with amino surface groups; Eq^m, equilibrium.

As with IE, subsequent regeneration is required to reuse the magnetic microsorbents, and regeneration is carried out in either single solutions or mixtures thereof. Drenkova-Tuhtan et al. (2013) demonstrated 88% and 95% adsorption and desorption respectively, using MgFe-Zr in a pilot scale experiment using treated sewage effluent. Desorption was carried out in a 1 M NaOH + 1 M NaCl solution for 30 min. In a recent study Drenkova-tuhtan et al. (2017b) used 1 M NaOH solution to regenerate ZnFeZr adsorbent and achieved 95.2% and 86% adsorption and desorption efficiencies, respectively, in a pilot-scale experiment using spiked treated sewage effluent. The authors achieved a regenerant PO₄³⁻P concentration (382 mg/L) 38 times the initial concentration (∼10 mg/L PO₄³⁻P). Ishiwata et al. (2010) demonstrated 83.8% desorption from ferromagnetic zirconium ferrite using a 7% NaOH solution. Even though Drenkova-Tuhtan et al. (2017a) showed reusability of the adsorption media to some extent, de Vicente et al. (2010) and Merino-Martos et al. (2011) observed 20% and 36% reduction of P adsorption capacity after regeneration. It should also be noted that many of these authors used highly alkaline chemicals in regeneration of adsorption material, as recommended in previous reports.

Reactive filtration

Reactive filtration is a process whereby P is removed by multiple mechanisms, including coprecipitation, adsorption, and physical filtration. Both orthophosphates and particulate phosphorus are removed due to the multiple mechanisms involved. Newcombe et al. (2008) demonstrated 90.3% of P removal at the Hayden Area Regional Wastewater Treatment Facility in Hayden, Idaho. The process involved moving-bed sand filters filled with iron oxide-coated sand (Newcombe et al., 2008), which was continuously regenerated with FeCl₃ (Newcombe et al., 2006). Sutton et al. (2011) indicated that backwash from reactive filtration can be used as a potential slow release P fertilizer.

Studies indicate that reactive filter materials can be used to recover phosphorus. Polonite is such a filter material (Gustafsson et al., 2008; Renman and Renman, 2010; Nilsson et al., 2013), which can be used as a soil amendment (Renman et al., 2009). Recent studies have shown that activated zeolites can remove and recover phosphates (Mehrez Hermassi et al., 2016) and simultaneously recover ammonia (You et al., 2017; Hermassi et al., 2018). These activated reactive zeolites remove phosphates mainly by precipitation with Ca and Mg, to form brushite and struvite, and surface complexation with Al and Fe (Mehrez Hermassi et al., 2016; You et al., 2017). The authors indicate that these zeolites can be potentially used as slow release fertilizers (Mehrez Hermassi et al., 2016; You et al., 2017).

Waste materials such as steel slag (Shilton et al., 2006; Gustafsson et al., 2008; Lee et al., 2010; Barca et al., 2012, 2014), oil shale ash (Mo et al., 2010), and fly ash (Ragheb, 2013) have been investigated for P removal capabilities. Results indicated that the principal P removal mechanism involved dissolution of CaO and precipitation of Ca-phosphate complexes or hydroxyapatite. Several authors pointed out the high pH of these slag filter effluents (Lee et al., 2010; Barca et al., 2014). Therefore, further investigations are required to ensure the applicability of these materials as fertilizer (Vohla et al., 2011).

Urine separation

It has been reported that 50–80% of P and 80–90% of N and K in domestic wastewater can be attributed to the urine, which itself corresponds to less than 1% of the total volume of the domestic wastewater flow (Maurer et al., 2003; Ronteltap et al., 2007; Ishii and Boyer, 2015; Zamora et al., 2017). On average typical urine of an adult contains 8180 mg/L of N and 670 mg/L of P (Maurer et al., 2003); hence, source separation of urine would greatly reduce the nutrient load to treatment plants. In addition, O'Neal and Boyer (2013) showed that P can be recovered more effectively at the point of generation than at a conventional centralized treatment plant, and such source separation might allow treatment of pharmaceuticals and hormones in higher concentrations given that the majority of such compounds are contained in the urine (Wilsenach and Van Loosdrecht, 2004; O'Neal and Boyer, 2013; Ishii and Boyer, 2015).

Studies have reported P recovery, and simultaneous N and P recovery, from source-separated urine using struvite precipitation. For example, some authors observed coprecipitation of struvite type compounds such as MgK(PO₄) and MgNa(PO₄) (Wilsenach et al., 2007; Ronteltap et al., 2010; Ishii and Boyer, 2015; Tian et al., 2016; Zamora et al., 2017). In particular, Udert et al. (2015) indicated P recovery by struvite precipitation. Furthermore, although N recovery generally requires substantial chemical addition (Ishii and Boyer, 2015), Udert et al. (2015) were able to produce a highly concentrated nutrient solution containing NH₄NO₃, and even inactivate pathogens, by coupling struvite precipitation with nitrification and distillation. In addition, Merino-Jimenez et al. (2017) showed that urine can be used as a fuel in microbial fuel cells (MFCs), after pre-precipitating struvite using MgSO₄, MgO, Mg(OH)₂, MgCl₂, or synthetic sea water as a cost-effective Mg source. Xu et al. (2011) recovered 77% and 98% of P and K simultaneously from pretreated ammonia-free urine, as per the method proposed by Başakçilardan-Kabakci et al. (2007), although Mg²⁺ and PO₄³⁻ were added externally.

Several factors may hinder the widespread applicability and ultimate efficiency of urine separation for nutrient recovery and other purposes. First, implementation would incur costs, energy demands, and practical issues related to (a) the onsite homeowner-managed storage and treatment of urine or (b) the installation and maintenance of separate conveyance networks (e.g., hydraulic or transport based) for urine. In addition, public acceptance of urine separation is a critical factor in its implementation, although a survey conducted by Ishii and Boyer (2016) indicated a high level of support within a university community for such source separation.

Struvite precipitation

Struvite is considered a good slow-release fertilizer and may achieve enhanced agronomic properties when combined with complex fertilizers (Bridger et al., 1962; Ryu et al., 2012; Talboys et al., 2016). Also known as magnesium ammonium phosphate, struvite has the advantages of minimum leaching, less-frequent need for application, and lack of fertilizer burn of the crops even at high application rates (Bridger et al., 1962; Münch and Barr, 2001). Several studies indicated that P recovery by struvite crystallization had advantages over other technologies in terms of purity, crystalline form, and dewatering characteristics of the product, the efficiency of P removal, the presence of Mg in struvite, the ability to simultaneously remove ammonia, and lower evaporative N losses compared to other N rich fertilizer (Parsons and Smith, 2008; Muster et al., 2013; Liu et al., 2013b; Kumar and Pal, 2015; Puchongkawarin et al., 2015).

Struvite is crystallized when the concentrations of PO₄³⁻, NH₄⁺, and Mg²⁺ exceed the struvite solubility index. Mavinic et al. (2007) indicate that a supersaturation ratio of at least 20 is required to achieve 80% P removal. Struvite has a minimum solubility around pH 9, and the optimum pH for struvite precipitation is in the range of pH 7–11. The specific optimal pH depends on nutrient composition, and low-P streams will require pH higher than 9.0 (Adnan et al., 2003, 2004). Mg²⁺ is identified as the limiting component in struvite precipitation from wastewater treatment side streams. Accordingly, MgCl₂ and Mg(OH)₂ are two Mg sources used in struvite precipitation, with Mg(OH)₂ being less expensive and beneficial in terms of pH adjustment but requiring a long reaction time due to its low solubility (Münch and Barr, 2001). When MgCl₂ is used as a Mg source, pH is adjusted by adding NaOH. Typically an Mg:P ratio of 1.3:1 is maintained (Katsuura, 1998; Dastur, 2001; Münch and Barr, 2001; Jaffer et al., 2002; Britton et al., 2005).

P recovery as struvite is a common separation step applied to many nutrient-rich streams such as source-separated urine, landfill leachate, swine wastewater, industrial effluent, anaerobic sludge digester supernatant, and IE regenerant solutions (Kumar and Pal, 2015). Among those applications, struvite precipitation from anaerobic sludge digester supernatant in enhanced biological phosphorus removal (EBPR) plants is particularly useful (De-Bashan and Bashan, 2004; Gassie et al., 2016), as it allows separation of phosphorus which would otherwise recirculate within the plant (Münch and Barr, 2001).

Several studies report P recovery as struvite combined with EBPR (Morse et al., 1998; Venkatesan et al., 2015) as a preferred approach for P recovery (Münch and Barr, 2001). Münch and Barr (2001) and Shu et al. (2006) showed that struvite precipitation in anaerobic digesters can achieve very low P concentrations in EBPR effluents. However, P removal to <5 mg/L is expensive, and at least 40 mg/L of P is required to achieve >80% P removal by struvite precipitation (Adnan et al., 2004; Britton et al., 2005). In addition, because the N concentration in an anaerobic sludge digestion side stream (700–800 mg/L) is typically much higher than the stoichiometric value required for struvite precipitation, only 6% of N is removed (Münch and Barr, 2001; Mehta et al., 2015).

Source separated urine is a nutrient-rich stream in which P is recovered by struvite precipitation. Zamora et al. (2017) demonstrated the first continuous operation of a pilot struvite plant fed by source-separated urine and achieved 85–99% P recovery during 12 months of continuous operation. The authors achieved over 90% purity, with impurities being small amounts of calcium and potassium precipitates and heavy metals at concentrations well below the allowable limits for fertilizers in The Netherlands. Although the optimum Mg:P ratio for struvite precipitation is 1.3:1 (Wilsenach et al., 2007; Xu et al., 2011), P recovery increases with increasing Mg (Liu et al., 2008); however, typical urine has a Mg:P ratio of only 0.1:1 (Xu et al., 2011). Although pH adjustment is not required because of the inherent ammonium/ammonia buffering capacity of urine (Ronteltap et al., 2007), addition of excess Mg²⁺ as MgCl₂ is required (Zamora et al., 2017), by the addition of RO concentrate (Tian et al., 2016), or seawater and brine (Liu et al., 2013a).

P can be recovered from source-separated urine in various struvite type compounds. Wilsenach et al. (2007) indicated that although MgK(PO₄) could be coprecipitated from source-separated urine, MgK(PO₄) volume index, in mg-P/L of settled precipitates, was low compared to that of precipitated MgNH₄PO₄. Başakçilardan-Kabakci et al. (2007) achieved 95% P removal by precipitating either ammonium or potassium struvite, but Xu et al. (2011) indicated that ammonium type struvite has a higher tendency to precipitate as it has a low solubility compared to sodium and potassium type struvite. Regardless of the form of struvite precipitated, nutrient recovery from source-separated urine requires chemical addition.

Even though struvite precipitation has been a recommended P recovery process, the cost of chemical addition for Mg supplement and pH adjustment is a major drawback. Jaffer et al. (2002) indicated that 97% of the chemical costs were attributed to NaOH for pH adjustment, and several studies have indicated that CO₂ stripping could lower the caustic requirement for pH adjustment (Battistoni et al., 1997; Fattah et al., 2008, 2010; Stolzenburg et al., 2015). Venkatesan et al. (2015) indicated that struvite precipitation from an anaerobic sludge digestion side stream is economical, and projected operation and management cost for the struvite plant would be as high as the capital cost due to the Mg(OH)₂ requirement.

Electrodialysis

Electrodialysis is a process in which anions (e.g., PO₄³⁻, SO₄²⁻, and Cl⁻) and cations (e.g., NH₄⁺, K⁺, Na⁺, Ca²⁺, Mg²⁺, and so on) are separated using anion and cation exchange membranes in the presence of an electric field. During this process, concentrated solutions of anions and cations can be obtained separately. This process is typically used in N and K recovery (Mehta et al., 2015), but is also used to concentrate P such that it can be subsequently recovered typically as struvite.

As mentioned in Struvite precipitation section, it may be important to concentrate P to reduce the chemical cost (Adnan et al., 2004; Britton et al., 2005) and increase recovery (Zhang et al., 2013), since low P streams may require pH adjustment for struvite crystallization. In addition, although Zhang et al. (2013) demonstrated 93% recovery from struvite reactor effluent with a low phosphate-P concentration of 20–30 mg/L, the authors indicated that struvite precipitation would not be cost efficient at such low P concentrations because optimum pH would be higher than pH 9.

Chemical precipitation

Chemical precipitation by addition of metal ions is popular in Europe and Scandinavian countries (Hultman and Löwén, 2001; Egle et al., 2015), which can achieve European discharge standards with low chemical consumption (Clark et al., 1997). Chemical precipitation is considered an efficient stable process (Ye et al., 2017), which can achieve 0.1 mg/L P effluent with high chemical doses, while chemical cost can be reduced by combining the process with advanced filtration technology to remove fine P precipitates (Tetra Tech, Inc., 2014). Fe³⁺, Al³⁺, and Ca²⁺ are the most common metal ions used in chemical precipitation (Baettens, 2001; Mulbry et al., 2005; Klimeski, 2007; Bertanza et al., 2013; Kim and Chung, 2014; Egle et al., 2015; Sengupta et al., 2015). Chemical precipitation is often used simply to remove P from wastewater, with precipitates removed together with sewage sludge. However, in P-recovery applications, this method was reported to increase the sludge volume up to 35% (Sengupta et al., 2015), which caused problems in handling and disposal (Antunes et al., 2018).

In general, lower P dissolution from metal phosphate compounds such as hydroxyapatite restricts the plant availability of P in the recovered sludge (De-Bashan and Bashan, 2004; Egle et al., 2015), and disadvantages of chemical precipitation include the need to adjust pH, the cost of reagents (e.g., coagulants), inhibitory effects on biological processes, increased sludge production (Haas et al., 2000), the accumulation of fixed solids at high Fe³⁺ application rates (Bertanza et al., 2013), and heavy metal leaching. To increase the solubility of metal phosphate compounds, postprocessing of the sludge either by wet-chemical extraction or wet-chemical leaching is required (Egle et al., 2015). Furthermore, De-Bashan and Bashan (2004) indicated that cost-effective methods such as the use of phosphate-solubilizing bacteria and phosphate-solubilizing fungi can aid in solubilizing hydroxyapatite (Whitelaw, 1999; Richardson, 2001).

Phostrip is one of the processes which recovers phosphates from the sludge settler supernatant. P is released from the sludge due to the anaerobic conditions maintained in the stripper and, subsequently, precipitated in a pH regulated reactor (pH = 9) as hydroxyapatite by adding slaked lime or NaOH (Kaschka and Weyrer, 1999). However, the process has the disadvantages of requiring additional sludge digestion and associated costs (Bertanza et al., 2013) and separation of hydroxyapatite precipitates from the sludge, to avoid problems associated with direct application (Zou and Wang, 2016).

Biological P recovery

P can be removed biologically by the EBPR process, and the process has been preferred over chemical precipitation when stringent P limits are not enforced (Münch and Barr, 2001; Nancharaiah et al., 2016). The EBPR process is based on the accumulation of P by phosphate-accumulating organisms (PAOs), which results in the enrichment of P in the sludge (Mulkerrins et al., 2004), removing as much as 80–90% of P (Morse et al., 1998). Wasting of such P-rich sludge can achieve high P-removal rates (Mino et al., 1998). Furthermore, the addition of an anoxic tank can also incorporate simultaneous nitrogen removal (Tchobanoglous et al., 2003). Zou and Wang (2016) indicated that denitrifying PAOs can accumulate up to 20% of their dry biomass weight in P, and EBPR can achieve P concentrations <1 mg/L at the appropriate conditions (Mulkerrins et al., 2004; Venkatesan et al., 2015). Alum is added to reduce P to <1 mg/L when insufficient biological P removal occurs. However, cost of chemicals for chemical precipitation may become prohibitively high (Stratful et al., 1999).

EBPR is a very sensitive biological process, which requires careful control of many operational parameters. Such parameters include pH, temperature, volatile fatty acid content, cation concentration, dissolved oxygen concentration, food-to-microorganism ratio, hydraulic retention time, solid retention time, and wastewater composition. In fact, Pitman (1991) indicated that the desired P removal may not be achieved when the anoxic chemical oxygen demand (COD):P ratio drops below 50 (Mulkerrins et al., 2004). Furthermore, Powell et al. (2008) indicate that EBPR is an expensive and operationally complex process. This conclusion implies that EBPR may be expensive and unstable for small treatment facilities, which typically experience sudden changes in flow rate and influent composition.

Even though EBPR can remove P in wasted sludge, P reuse involves direct land application or separation from concentrated streams. Although direct land application is effective, the product is bulky and may pose contamination risks. Therefore, P is often recovered from nutrient-rich anaerobic sludge digestion supernatant (Yuan et al., 2012), typically as struvite (Münch and Barr, 2001) as explained in Struvite precipitation section.

Algae harvesting

Nutrient recovery using algal biomass has drawn considerable attention due to current nutrient scarcity. Shilton et al. (2012) indicate that algal ponds and macrophytes can recover nutrients using only one tenth of the land required for recovery by terrestrial crops/pastures, or potentially even less considering potential luxury uptake. According to Powell et al. (2008) and Richmond (2004), P content in dry algal biomass can reach 3.4%, including such luxury uptake, which depends on light intensity, P concentration in the medium, and temperature (Powell et al., 2009; Cade-Menun and Paytan, 2010; Fanta et al., 2010).

In addition to growth in suspension, algae can be also grown in immobilized beads and algal turf scrubbers (Garbisu et al., 2000; Christenson and Sims, 2011), and this technique has been used in treating many types of wastewater, including dairy (Pizarro et al., 2002) and swine manure (Kebede-Westhead et al., 2006). Shilton et al. (2012) indicated that experimentation with offshore nutrient recovery using algae has also been proposed.

Although algae harvesting appears to be a viable option for nutrient recovery, research on the use of algal biomass as a fertilizer is still at an early stage (Shilton et al., 2012). Compared with the biomass obtained from EBPR, algal biomass would require a longer time for release of nutrients. Therefore, further processing of algal biomass, such as hydrothermal liquefaction (Jena et al., 2011), may be required to produce fertilizer. However, recent research indicated that direct application of dried algal biomass demonstrated similar performance to the application of commercial fertilizer, with heavy metal loading well below permitted levels (Mulbry et al., 2005). However, pathogens and micropollutants may be a concern.

Currently, regulations and public acceptance are barriers to the use of algal biomass as fertilizer (Solovchenko et al., 2016). In addition, many aspects of the use of algal biomass as fertilizer require further investigation (Solovchenko et al., 2016), including potential adverse effects of bald application of algal biomass, nutrient availability to crops, potential effects of allelochemicals and/or cyanotoxins (Berry et al., 2008), and pathogens and micropollutants (Mulbry et al., 2005).

Although the use of algal biomass as a direct plant nutrient is limited to date, its use as animal feed and a source for biofuel production has attracted attention (Madeira et al., 2017; Adeniyi et al., 2018; Barbera et al., 2018). Current research has produced promising results on the use of algae for livestock production, and both applications alleviate competition for the use of food crops. However, algae cultivation techniques require improvements to reduce cost (Madeira et al., 2017), and the cost of algal biofuel production is currently a principal limitation to large-scale implementations (Doshi et al., 2016). Interestingly, cultivation and harvesting, specifically concentration of the harvested algal biomass, are energy intensive and consume 25–70% of the energy produced (Su et al., 2017). Hence, although there is a large driving force for algal biofuel research (Reed Business Information Limited, 2018), energy costs remain a drawback, and the approach is still far from commercialization (Su et al., 2017).

Electrochemical P recovery

Electrochemical nutrient recovery methods can be divided into two categories: (1) processes which use sacrificial anodes, and (2) processes which use dimensionally stable anodes (DSAs). In general, processes which use sacrificial anodes are operated so as to dose Mg²⁺, Al³⁺, Fe²⁺, or Fe³⁺, to assist chemical precipitation (Chen, 2004; Hug and Udert, 2012; Kruk et al., 2014; Huang et al., 2017). Processes which use DSAs work by altering the water matrix, which in turn precipitates phosphate compounds (Kappel et al., 2013).

Among the processes using sacrificial anodes, Mg anodes are becoming quite popular because of their identification as a cost-effective alternative for chemically induced struvite precipitation (Hug and Udert, 2012). The authors indicated that the process is economical compared with the addition of fast dissolving chemicals such as MgCl₂ or MgSO₄, but more expensive than adding slowly-dissolving MgO. Electrochemical dosing of Mg²⁺ also improves struvite recovery due to continuous Mg dosing compared with one-time chemical addition. Ronteltap et al. (2010) indicated that continuous dosing can lower the supersaturation ratio compared to one-time chemical addition and promote the growth of large particles, which in turn improves recovery. Even though Hug and Udert (2012) did not experience a considerable pH shift, probably due to buffering in urine, Mariakakis et al. (2017) indicated a shift in pH when treating a centrate recovered in the dewatering of digested municipal wastewater treatment sludge. This result implies that electrochemical Mg dosing would also reduce or eliminate the chemical cost of caustic addition. However, the method will only work with influents containing sufficient excess ammonia for struvite precipitation.

As an alternative to chemical addition, Kappel et al. (2013) have shown a process using DSAs to shift the pH of a water matrix, which causes phosphate to precipitate with counter cations in the solution. The authors demonstrated over 90% phosphate recovery by shifting the pH above 9. Kappel et al. (2013) indicated that use of multivalent cation exchange membranes also allow migration of K⁺, Na⁺, Ca²⁺, and Mg²⁺ ions, which are typically present at much higher concentrations than H⁺ at neutral pH, and avoid cross-neutralization of the compartments. Phosphate recovery was proportional to the final pH, and the recovered precipitates consisted principally of amorphous calcium phosphate and amorphous calcium carbonates. Thus, the process did not require chemical dosing, reducing cost and complexity.

Nitrogen recovery

Nitrogen removal from wastewater streams is important in preventing eutrophication of water bodies. Therefore, conventional wastewater treatment processes are focused principally on removal of nitrogen from wastewater, most commonly by biological nitrification and denitrification. Newly developed processes, such as Anammox, can convert ammonia and nitrite to nitrogen gas directly under anaerobic conditions (Jetten et al., 1998). However, these processes do not recover nitrogen but return them to the atmosphere. Hence, recovery of reactive N from wastewater streams would be an energy-efficient alternative to the manufacture of fertilizer from atmospheric nitrogen.

IE/Adsorption

IE/adsorption is a widely studied process for ammonia removal from wastewater. Use of zeolites as adsorbents is very attractive in NH₄⁺ removal because it is efficient, economically competitive, and operationally simple (Koon and Kaufman, 2018) and can withstand shock loadings (Huang et al., 2010). Natural and modified zeolites and clinoptilolite are the most commonly used adsorbents for ammonium removal (Gupta et al., 2015; Sengupta et al., 2015). Typically zeolites have a porous structure made of Al³⁺, Si⁴⁺, and O, having exchangeable cations such as Na⁺, K⁺, Ca²⁺, Mg²⁺, and Ba²⁺, which have a high affinity toward NH₄⁺ (Dixon and Weed, 1989; Saltali et al., 2007). Natural zeolites and clinoptilolite were reported to have ammonium adsorption capacities ranging from 3.11 to 13.73 mg/g (Gupta et al., 2015).

The adsorption capacity of natural clinoptilolite can be enhanced by pretreatment with NaCl (Vassileva and Voikova, 2009). The authors indicated that adsorbent dosage, pH, temperature, and initial NH₄⁺ concentration affected NH₄⁺ removal efficacy in both natural and pretreated clinoptilolite. In fact, the ammonia/ammonium equilibrium is controlled by pH, and only ammonium ions can be removed from adsorption. Mohd et al. (2010) also indicated that zeolite synthesized from rice husk ash had a higher adsorption capacity (46.56 mg/g) compared with natural mordenite (15.13 mg/g).

Recent studies with new ammonium adsorbents suggest more efficient alternatives to zeolite adsorption. In particular, carbon nanotubes (Moradi and Zare, 2013), Romanian volcanic tuff (Marañón et al., 2006), and wheat straw (Ma et al., 2011) had adsorption capacities of 17.05, 19, and 148.7 mg/g, respectively. In addition, novel palygorskite nanocomposite exhibited an extremely high ammonium adsorption capacity of 237.6 mg/g (Wang et al., 2014) and achieved equilibrium in 12 min. The high ammonia absorption capacity was attributed to adsorption of NH₄⁺ to -COO⁻ groups in the polymeric network, from pH 4 to 8, as previously observed by Zheng et al. (2012). Similar NH₄⁺ absorption mechanism by -COO⁻ groups was observed by Wang et al. (2014). The authors also indicate that the adsorbent can be subsequently used as a slow-release fertilizer; however, adsorption capacity of the material was significantly decreased in the presence of competing cations.

Lin et al. (2014) indicated that zeolites can also be used to simultaneously remove N and P from high strength wastewater. P was removed by precipitation of thermodynamically favorable hydroxyapatite as Ca²⁺ is released during ammonium adsorption (Lin et al., 2014). The authors suggested that spent zeolites can then be used as a green fertilizer. Liberti et al. (1978, 1980) also demonstrated recovery of both N and P from spent NaCl regenerant solution. In particular, Liberti et al., (1978) recovered N and P as struvite, by adding Mg(OH)₂, whereas Liberti et al. (1980) used both struvite precipitation and absorption into H₂SO₄ after stripping NH₃ from the spent regenerant solution following pH shifting. The former team indicated that resin life, regeneration efficiency, and nutrient recovery need to be optimized through extensive pilot plant studies.

Even though adsorption is potentially effective and economical for ammonium separation, ion selectivity and the requirement for regeneration are drawbacks. Huang et al. (2010) observed that the adsorption capacity of zeolite was significantly impaired by the presence of cations Na⁺ > K⁺ > Ca²⁺ > Mg²⁺ and anions CO₃²⁻ > Cl⁻ > SO₄²⁻ > PO₃³⁻. Some studies indicate the possibility of direct application of spent zeolite adsorbents as fertilizer (Malekian et al., 2011; Smith and Smith, 2015; Sengupta et al., 2015). However, when reuse of the adsorbent is desired it must either be regenerated with concentrated brine, typically NaCl solution (Liberti et al., 1978, 1980; Demir et al., 2002; Li et al., 2011; Sengupta et al., 2015), or by thermal regeneration (USEPA, 1971; Demir et al., 2002). Some authors suggest air stripping of ammonia from regenerant solution when regenerant reuse is desired (USEPA, 1971; Liberti et al., 1980; Koon and Kaufman, 2018). However, few studies focus on disposal of spent regenerant or recovery of adsorbed ammonia, and such approaches are still under development (Vassileva and Voikova, 2009; Smith and Smith, 2015; Koon and Kaufman, 2018).

Electrodialysis

Electrodialysis has been used to concentrate and recover ammonia, more often than to recover phosphate (Mehta et al., 2015), and the same principle as described in Electrodialysis section has been used. In contrast with phosphate recovery applications, however, ammonia is concentrated in the cathodic side of the electrodialysis cell.

Typically the practice of ammonia recovery using electrodialysis involves subsequent stripping and absorption of the concentrated ammonia solution (Desloover et al., 2012; Ippersiel et al., 2012). Desloover et al. (2012) suggested that stripping and absorption could achieve nearly 100% efficiency, but the recovery rate of the electrodialysis process was limited by the concentrations of Na⁺ and K⁺, in the influent. The authors indicate that the addition of NaCl to the catholyte increased ammonium flux, by reducing the Na⁺ concentration gradient (Desloover et al., 2012). Furthermore, the generation of OH⁻ ion at the cathode eliminated the chemical requirement for pH elevation. While precipitate formation on the cathode can be problematic in long-term operation, the problem can be addressed by intermittent polarity reversal (Pikaar et al., 2011, 2013; Desloover et al., 2012).

Electrochemical

Ammonia has also been recovered electrochemically, similar to an electrodialysis process involving concentration of ammonia in the cathode compartment of an electrochemical cell. Ammonia was then recovered by stripping and absorption into an acid such as H₂SO₄ to produce ammonium salts (Luther et al., 2015). pH adjustment was not necessary due to OH⁻ generation at the cathode by water electrolysis. However, the authors indicated that buffering capacity of the influent is important to maintain NH₄⁺ current efficiency. Electromigration depends on the valence, concentration, diffusion coefficient of the ionic species, and the strength of the electrical field (Harnisch et al., 2009). In fact, the mobility of H⁺ is 5–6 times higher than NH₄⁺, K⁺, and Na⁺ (Rozendal et al., 2006), and hence NH₄⁺ transfer would be limited once the anode pH drops. Luther et al. (2015) also suggested that the H₂ produced can be used as a passive means to strip ammonia, to reduce air pumping costs for stripping. Absorption efficiency of 95% ± 5% was observed in their experiment.

A recent study indicated that recycling H₂ generated in the cathode compartment can further reduce the energy required for electrochemical nitrogen recovery from urine (Kuntke et al., 2017). The authors indicate that energy demand for the H₂ recycling electrochemical system (26.1 kJ/g-N) is lower compared with the best performing electrochemical systems reported (31 kJ/g-N) (Luther et al., 2015), achieved by lowering cell potential as a result of H₂ oxidation on the anode (Kuntke et al., 2017). However, the authors indicated that urine needs to be pretreated by adding MgCl₂, as described by Zamora et al. (2017) to prevent scaling inside the system (Kuntke et al., 2017). Drawbacks of the process include the requirement of additional electrolyzer to produce about 10% of H₂ to compensate for H₂ losses, use of expensive Pt electrodes, and possible deactivation of Pt catalyst by Cl⁻.

Recently, Muster and Jermakka (2018) demonstrated NH₃ removal and recovery using an electrochemically assisted ammonia recovery process, involving electro-oxidation and an electrochemically-assisted surface transfer mechanism, respectively. This novel process recovered ammonia from high strength NH₃ streams (900–560 mg/L) without the use of an IE membrane and with a theoretical energy requirement (29.3 kJ/g-N) comparable to H₂ recycling electrochemical system developed by Kuntke et al. (2017).

Bioelectrochemical

Bioelectrochemical systems work similar to electrochemical nutrient recovery systems, but can generate the energy required for the process when the influent contains sufficient organic matter, using a MFC (Wu and Modin, 2013). Kuntke et al.'s (2012) energy balance indicated that an MFC can generate ∼61% of the aeration energy for ammonium recovery from urine. In bioelectrochemical systems, biologically catalyzed oxidation of organic substrate takes place on the anode, and released electrons travel through an external resistor to reduce O₂ to OH⁻ at the cathode. Due to the reaction just mentioned, cations such as H₃O⁺, Na⁺, K⁺, Mg²⁺, Ca²⁺, and NH₄⁺ are transported to the cathode compartment through a cat IE membrane to maintain the charge neutrality. Because the air-cathode is continuously aerated, NH₄⁺ concentrated at the cathode is stripped with air as NH_3(g) at the high-pH condition in the cathode compartment. Ammonia in the leaving gas can be subsequently absorbed into an acid for recovery (Kuntke et al., 2012; Gildemyn et al., 2015).

The bioelectrochemical process was demonstrated on many types of wastewater, including diluted urine (Ieropoulos et al., 2012; Wu and Modin, 2013), undiluted urine (Kuntke et al., 2012), high strength wastewater (Kuntke et al., 2011), reject water from sludge treatment processes (Wu and Modin, 2013), synthetic digester liquor (Qin and He, 2014), and real landfill leachate (Qin et al., 2016). However, Luther et al. (2015) indicated that electrical current production is limited by microbial activity at the anode, the COD concentration, and pH (He et al., 2008; Yu et al., 2013; Liu et al., 2016; Li and Chen, 2018), and may cause irreversible reduction of MFC performance at pH ≤4 (Zhang et al., 2011). In addition, although the process produced some H₂, a potential fuel, and consumes less energy than electrochemical treatment alone (Kuntke et al., 2012; Wu and Modin, 2013), the process is sensitive to pH, influent toxicity, and carbon loading especially as regards readily-biodegradable substrate. Therefore, the process may not be stable for fluctuating loads, and electrochemical methods will have more stability and control in such applications (Luther et al., 2015).

Struvite precipitation

In addition to its use for phosphorous recovery, struvite precipitation can be carried out targeting nitrogen recovery. However, both Mg²⁺ and PO₄³⁻ need to be added because NH₄⁺ often exists in excess of the stoichiometric requirement for struvite precipitation, and the amount of P is insufficient for complete N removal (Yetilmezsoy and Sapci-Zengin, 2009; Escudero et al., 2015; Nancharaiah et al., 2016; Jia et al., 2017). Even though El Diwani et al. (2007) showed that bittern can be used as an alternative Mg²⁺ source, the use of seawater and its derivatives would reduce the product purity due to precipitation of hydroxyapatite, calcium phosphates, and calcite at high Ca/P-PO₄³⁻ ratios (Liu et al., 2013a; Lahav et al., 2013; Rubio-Rincón et al., 2014). In contrast, when pure Mg sources are used, the cost of chemical addition and the need for pH control have limited the operation of industrial scale struvite precipitation reactors for nitrogen recovery (Escudero et al., 2015). However, recent work has been carried out to improve the struvite precipitation process, and preliminary economic analysis indicates the technical and economic feasibility of the process (Jia et al., 2017).

Stripping and absorption

Air stripping of ammonia has been investigated in many studies (Bonmatı and Flotats, 2003; Lei et al., 2007; Gustin and Logar, 2010; Desloover et al., 2012; Ippersiel et al., 2012; Kuntke et al., 2012; Wu and Modin, 2013; Kelly and He, 2014), and Wu and Modin (2013), in particular, demonstrated recovery efficiencies as high as 94%. In many cases, ammonia is stripped from pH-elevated concentrated ammonia solutions (greater than 1,000 mg/L) and recovered into an acid solution (Bonmatı and Flotats, 2003; Lei et al., 2007; Desloover et al., 2012; Ippersiel et al., 2012; Kuntke et al., 2012; Wu and Modin, 2013). Most ammonia recovery methods, including electrochemical and bioelectrochemical technologies, use stripping and absorption into an acid as a final step in the recovery (Saracco and Genon, 1994; Desloover et al., 2012; Wu and Modin, 2013; Luther et al., 2015).

The kinetics of ammonia stripping have been found to be controlled by pH, temperature, and gas flow rate (Gustin and Marinsek-Logar, 2010; Liu et al., 2015). Among those parameters, pH has a dominant influence on stripping efficiency (Liu et al., 2015). In fact, Ippersiel et al. (2012) observed that only 0.22% of ammonia could be recovered from 21,000 mg/L concentrated ammonia at pH 8.3–8.6. Jiang et al. (2010) indicated that the optimum pH for ammonia stripping is above 9.5; however, Gustin and Logar (2010) observed that pH above 10 and an air-to-liquid ratio above 2000 did not substantially improve the recovery efficiency. Many studies on ammonia stripping have been conducted on high strength wastewater (Bonmatı and Flotats, 2003; Başakçilardan-Kabakci et al., 2007; Lei et al., 2007; Laureni et al., 2013). However, the process was successfully demonstrated on low ammonia concentrations as well (Culp et al., 1978). The process is typically carried out in a packed bed tower to improve mass transfer efficiency (Sengupta et al., 2015). Moreover, pH, temperature, gas flow rate, and the time for desorption are principle design parameters in ammonia desorption systems and must be optimized for the specific characteristics of the wastewater to be treated (Srinath and Loehr, 1974; Shpirt, 1981).

Trace concentrations of ammonia, and recarbonation, can cause pH drift and reduce stripping efficiency. In a study of ammonia stripping from landfill leachate after pH adjustment using Ca(OH)₂, it was observed that pH was lowered due to recarbonation of lime by atmospheric CO₂ (Cheung et al., 1997). Yoon et al. (2008) indicated that trace ammonia concentrations in the stripping gas might affect the removal efficiency. Trace NH₃ concentrations can be expected, when a gas is recirculated in between ammonia stripping and absorption columns.

The principal drawback of ammonia stripping has been reported to be the high energy requirement (Kuntke et al., 2012). However, Liu et al. (2015) indicate that the process is widely used due to its low cost, easy installation, and high recovery rate. The authors point out that to achieve 80% recovery, temperature and air flow rate can be increased without affecting process economics, but that pH elevation would increase the operating cost due to chemical requirements. The energy requirement for conventional air stripping and (NH₄)₂SO₄ production from urine has been reported at 9 kWh/kg N, in addition to embodied chemical energy for pH adjustment (Maurer et al., 2003; Desloover et al., 2012). This energy requirement may be compared to that of ammonia fertilizer production using Haber–Bosch process, ∼10 kWh/kg (Maurer et al., 2003; Luther et al., 2015). Aeration energy can be minimized using more efficient air stripping reactors such as the jet loop reactor (Deĝermenci et al., 2012) and the aerocyclone reactor (Quan et al., 2009).

Gas permeable membranes

The use of hydrophobic gas permeable membranes (GPMs) is a novel technology for ammonia recovery. Studies have demonstrated that on the order 95% of free ammonia can be recovered from both gaseous phases (Vanotti et al., 2009; Rothrock et al., 2013) and liquid phases (Vanotti and Szogi, 2010). Typically, NH₃ is recovered as (NH₄)₂SO₄ by recirculating H₂SO₄ in tubular membranes submerged in ammonia-containing wastewater or vice versa (Darestani et al., 2017). El-bourawi et al. (2007) and Garcia-González and Vanotti (2015) demonstrated over 90% ammonia removal and indicated that pH is a critical parameter and that the ideal pH is above 9. Darestani et al. (2017) showed that pH values of 10 or greater substantially accelerate ammonia removal, and recovery would drop to 57% at low pH and that this rate depends also on the NH₃ concentration (Garcia-González and Vanotti, 2015). Furthermore, low rate aeration of swine manure digester effluent with nitrification inhibitor can shift the pH to 8.3 and eliminate alkali chemical addition while achieving fivefold increase in the recovery rate compared with recovery without aeration (or alkali addition) (Dube et al., 2016). Recently, Vanotti et al. (2017) demonstrated simultaneous N (∼83%) and P (>90%) recovery using GPMs. Ammonia recovery using hollow fiber contactors is demonstrated in pilot and commercial scale; however, the technology is best used for ammonia removal down to 50 mg/L and number of modules and capital cost would substantially increase beyond that (Darestani et al., 2017).

Direct conversion to livestock feed and protein

Nitrogen in wastewater can be directly converted to protein-rich food sources through the use of heterotrophic microorganisms, analogous to algae harvesting discussed in Algae harvesting section (Matassa et al., 2015). This emerging technology short-circuits the nitrogen cycle, reducing N losses in the food production process. Biofloc technology is one example in which heterotrophic microorganisms assimilate waste material, including reactive nitrogen into biomass, which becomes a feed ingredient for fish (Widanarni et al., 2012). The technology has been successfully demonstrated in terms of feed requirement, without adverse effects on fish growth (Widanarni et al., 2012). Marine aquaponics is another example of direct conversion of N waste to plant proteins, where edible plants (for human consumption and animal fodder) such as halophytes are grown in recirculating aquaculture systems (Boxman et al., 2018).

Studies have identified single cell proteins (SCPs) as an alternative protein source for human populations, to alleviate food scarcity (Anupama and Ravindra, 2000; Matassa et al., 2015). Matassa et al. (2015) indicated that the majority of reactive N losses are associated with production of plant protein and proposed direct conversion of N recovered by physicochemical processes to edible biomass. Although benefits may be significant, SCP suffers from issues related to nucleic acids and toxins, and heavy metal contamination during growth on hydrocarbon media, and requires further study. Hence, strict sanitation and purification standards are enforced for quality control of SCP products (Anupama and Ravindra, 2000). In addition, the technology is still in its infancy, and production costs are high relative to conventional protein sources (Becker, 2007; Matassa et al., 2015).

Discussion

Among the currently developed nutrient recovery technologies, only few produce a fertilizer with minimal postprocessing. It is widely accepted that struvite is a good fertilizer, which can be directly applied without any postprocessing. However, processes such as IE and adsorption and electrodialysis produce only a concentrated solution from which nutrients would need to be recovered, for example, by precipitation with chemical addition as struvite or calcium phosphate. In addition, although chemical precipitation separates P from solution, P precipitates are typically wasted in a treatment plant along with the sludge, producing a bulky product. In contrast, electrochemical processes can recover P either as struvite or calcium phosphates, without increasing the sludge volume. In addition, when electrochemical processes are used to recover P from anaerobic sludge digester supernatant after biological P accumulation by EBPR, much purer product can be obtained. Furthermore, electrochemical processes can be used for simultaneous N and P recovery with minimal chemical use.

In general, all the recovery processes work efficiently when the nutrients are concentrated. In that context, urine separation can make the recovery more efficient. However, urine separation requires either onsite treatment, potentially problematic for homeowners, or a separate conveyance network for delivery to central urine treatment, a substantial investment. In addition, struvite precipitation in a urine conveyance network can be problematic if onsite pretreatment (i.e., phosphate precipitation) is not implemented. Public acceptance of onsite urine treatment is another concern that would need to be addressed before implementing nutrient recovery from source separated urine.

Phosphorus recovery

Although many P recovery processes are currently available for P recovery, the energy required may be very different and should be considered in process selection. According to Kappel et al. (2013), energy cost for P recovery using ∼23 mg/L nanofiltration concentrate was 2.4 kWh/kg-P at pH 9. This is substantially lower than the energy demand of 74.4 kWh/kg-P for electrocoagulation, based on a pilot scale study for P recovery (Smoczyński et al., 2017), 13.6 kWh/kg-P for P precipitation using FeSO₄, 4.7 kWh/kg-P for EBPR in wastewater treatment plants, 60.8 kWh/kg-P for thermal volume reduction of stabilized urine, 4.4 kWh/kg-P for struvite production, 7.5 kWh/kg-P for triple phosphate fertilizer production, and 5.6 kWh/kg-P for average P-fertilizer production in Europe (Maurer et al., 2003).

Overall, each nutrient recovery technique has advantages and disadvantages, and selection of a process for a particular application requires careful comparison.

Table 3 summarizes the pros and cons of P recovery technologies.

Table 3.

Comparison of P Recovery Processes

Process	Pros	Cons
Ion exchange	Can achieve high quality effluent <10 μg-P/L and <0.2 mg/L NH₄⁺-N). Can achieve >100x concentration No pH alteration required. No additional sludge generation.	Brine or NaOH required for media regeneration. 100% desorption of nutrients cannot be achieved during media regeneration. Leaching of other metal cations such as Al³⁺, Ca²⁺ affects purity of subsequent struvite precipitation. Biofouling of resin. Requires low suspended solid concentration in the influent to avoid clogging of the media. Continuous resin requirement due to reduction of regeneration capacity over time. Limited selectivity. Competition of other anions such as SO₄²⁻. Disposal of used media requires further treatment.
Reactive filtration	Can remove both orthophosphate and particulate phosphorus. Can achieve effluent P concentration of 0.011 mg/L depending on influent water.	Requires dosing chemicals such as FeCl₃. Recovery of nutrients from backwash requires further treatment.
Magnetic separation	Can achieve effluent P concentration of 0.01 mg/L depending on influent water. Process does not interfere with subsequent biological processes. Small footprint due to short required contact time.	Removal efficiency strongly dependent on adsorbent-to-P ratio. Regeneration requires NaOH solution. Cannot achieve high concentrations of P in the regeneration solution without impairing removal efficiency. Low P concentration in the regeneration solution makes subsequent recovery uneconomical.
Urine separation	50–80% of P and 80–90% of N can be separated from the bulk wastewater volume. Reduce nutrient load to treatment plants. Allow separation of pharmaceuticals from the bulk wastewater volume. High concentrations of nutrients (670 mg/L-P and 8180 mg/L-N) make recovery more efficient.	Onsite recovery is required unless separate collection network is installed. Possible health risks due to incomplete pathogen inactivation. Recovery requires chemical addition. N recovery requires pH adjustment. Public acceptance on onsite nutrient recovery needs further research.
Struvite precipitation	Considered as a good slow releasing fertilizer (minimum leaching, no fertilizer burn). Can be crystallized with lower impurities. Contains Mg as an additional nutrient	Requires chemical addition and pH adjustment. Cost of P removal from dilute solutions increases with chemical requirement.
Electrodialysis	Can achieve ∼8 times the P concentration. Can be used as a concentration step before struvite recovery to reduce chemical requirements.	Heavy metal accumulation in catholyte. Requires precipitation with chemical addition for recovery.
Chemical precipitation	Can be used to meet low P discharge limits. Easy to operate process. Can handle shock loadings.	Primarily used for P removal than recovery. Requires chemical dosing. Can affect preceding biological processes. Costs related to excess sludge generation and handling. Low pant availability of nutrients in chemically precipitated sludge. Heavy metal contamination.
Biological	Economical compared to chemical precipitation. PAOs can accumulate P up to 20% of their dry biomass weight. Can achieve 1 mg/L-P in effluent at appropriate conditions. Simultaneous P and N removal is possible. P can be efficiently recovered from anaerobic EBPR sludge digester supernatant.	Sensitive to influent water characteristics and operational parameters such as pH, temperature, VFA content, cation concentration, DO concentration, HRT, SRT. Operationally complex.
Algae harvesting	Land requirement is less compared with terrestrial crops/pastures. Can achieve P content up to 3.4% of dry biomass weight. Offshore implementation is possible. Heavy metal loading below permitted levels.	Luxury P uptake depends on light intensity, P concentration in the medium, and temperature. Slow nutrient release. Requires post-processing to improve nutrient release. Pathogens and micropollutants are a concern. Lack of information on potential effects on allelochemicals and/or cyanotoxins. Process is still under development.
Electrochemical	Chemical free process. Economical compared to chemical dosing in struvite precipitation. Continuous cation dosing compared to one-time chemical addition, hence better control of the process. Can use pH shifting for direct P recovery. Can be used as an alternative to chemical pH adjustment. H₂ generated at the cathode can be used as a potential fuel.	Coprecipitation of CaCO₃ in direct P recovery by pH shifting. Energy requirement is inversely proportional to wastewater conductivity.

EBPR, enhanced biological phosphorus removal; PAO, phosphate-accumulating organism; VFA, volatile fatty acid; DO, dissolved oxygen; HRT, hydraulic retention time; SRT, solids retention time.

Nitrogen recovery

Compared with P recovery, N recovery can be quite challenging due to the high solubility of NH₄⁺ salts, and only struvite can precipitate NH₄⁺ in waste streams. Other methods of separation involve concentration by IE/adsorption, electrodialysis, and electrochemical and bioelectrochemical techniques. Selection of a concentration process would depend on factors, including wastewater characteristics, target effluent concentration and recovery goals, concerns regarding chemical dosing and storage, and operational complexity. However, gas stripping and absorption into an acid is typically used as the end recovery technique in those concentration processes. Table 4 summarizes the pros and cons of N recovery techniques.

Table 4.

Comparison of N Recovery Processes

Process	Pros	Cons
Ion exchange/adsorption	Can meet very low N discharge limits. Operationally simple process. Can withstand shock loadings.	Mainly used for N removal rather than recovery. Limited adsorption/desorption capacity and resin life. Adsorption capacity impaired by competing ions. Requires chemicals and reagents for media regeneration. N recovery from spent regenerant requires chemicals or stripping and absorption into an acid solution.
Electrodialysis	N can be concentrated to reduce chemical and operational costs in post-recovery steps. OH⁻ generation on cathode eliminates chemical pH adjustment for post-recovery steps.	N recovery rate is limited by competing cations such as Na⁺ and K⁺. N recovery requires stripping and absorption into an acid solution.
Electrochemical	N can be concentrated to reduce chemical and operational costs in post-recovery steps. OH⁻ generation on cathode eliminates chemical pH adjustment for post-recovery steps. Produced H₂ can be used as a potential fuel and passive means for NH₃ stripping.	Reaction time limited by anode pH. N recovery requires stripping and absorption into an acid solution.
Bioelectrochemical	Can generate energy required for the process with sufficient organic matter. Concentrate N in cathode compartment for post-recovery. Eliminate chemical pH adjustment due to OH⁻ generation at air cathode. H₂ generated at the cathode can be used as a potential fuel.	Current production limited by microbial activity on anode. Sensitive to pH, influent toxicity, and substrate loading. N recovery requires stripping and absorption into an acid solution.
Struvite precipitation	Lower evaporative losses compared to other N rich fertilizer.	Both Mg²⁺ and PO₄³⁻ need to be added for N recovery. High operational costs due to chemical requirement.
Stripping and absorption	Direct recovery of N in fertilizer form is possible. Easy installation and simple operation. Energy requirement for NH₃ stripping and adsorption from urine is economical compared to NH₃ production from Haber–Bosch process.	Require pH adjustment. Typically used for high strength wastewater. Energy requirement for stripping is inversely proportional to influent NH₃ concentration. Not suitable for cold climates.
Gas permeable membranes	Direct recovery of N in fertilizer form is possible. pH shifting by low rate aeration in manure (a typical influent for this process) is sufficient to run the process.	Requires pH adjustment. Typically used for high strength wastewater (not economical <50 mg/L NH₃). High capital costs are expected when target effluent NH₃ is <1 mg/L.
Direct conversion to livestock feed and protein	Direct recovery of reactive N as feed or food Elimination of losses in conversion of reactive N to plant protein	Potential issues related to nucleic acids and toxins. Potential heavy metal contamination (depending on growth media) Requirement to meet extensive sanitation and purification standards. Current high production cost

Nutrient recovery technologies for onsite treatment applications

Onsite nutrient recovery is a distinct application which may entail different constraints in terms of process operation, maintenance, footprint, noise, odor, and the need for trained operators. Onsite treatment facilities may favor chemical-free technologies with a small footprint, which can be easily automated. However, most nutrient recovery technologies require continuous chemical dosing and associated chemical storage and handling, with no guarantee of the availability of trained personnel. For example, adsorption processes such as IE and magnetic microsorbents require chemical media regeneration, along with brine management involving nutrient recovery or brine storage and delivery. Similarly, electrodialysis produces a concentrated nutrient stream, which cannot be directly used as a fertilizer and requires chemicals for P precipitation. Likewise, urine separation, struvite precipitation, and chemical precipitation require chemical addition for P precipitation.

Even though biological processes can be operated without chemical dosing, they are operationally complex and sensitive to shock loadings, and in addition, separation of nutrients as fertilizer requires post-processing with chemical addition. These chemical requirements and operational complexities may hinder the adoption of many of the currently developed nutrient recovery technologies in onsite treatment systems. Electrochemical nutrient recovery, on the other hand, has advantages over other technologies when considered for onsite treatment. Advantages include (1) operation without continuous chemical dosing, (2) small footprint, (3) potential ease of automation, (4) ability to handle shock loadings, and (5) generation of a marketable product which can be directly used as a fertilizer.

Conclusions

Current nutrient recovery processes are focused principally on recovery of P or N separately, with process selection dependent on a wide array of factors, including influent wastewater characteristics, recovery goals, and operating constraints regarding the use of chemicals, the need for separate conveyance network, requirement for onsite treatment, and the need for trained operators. The following conclusions can be drawn based on the literature found in this review:

Struvite and electrochemical P precipitation can produce a fertilizer requiring minimal post-processing, although Mg salts, pH adjustment, and drying would be required for struvite precipitation and filtration and drying would be required for electrochemical precipitation;

Although many processes, including IE/adsorption, electrodialysis, electrochemical, and bioelectrochemical, can be used to concentrate N, only struvite precipitation and acid absorption following separation by gas stripping or gas permeable membranes can be used to recover N;

EBPR combined with chemical and electrochemical struvite precipitation and chemical precipitation alone are applicable for P recovery from centralized treatment plants, and struvite precipitation is often cited as a preferred process;

Electrochemical precipitation, chemical precipitation, and IE processes may have relatively low maintenance and chemical requirements and be adaptable to onsite and packaged treatment plant implementation;

Recovery of nutrients in algal biomass may be particularly useful for production of livestock feed and biofuel, and direct microbiological recovery as a protein source in food appears to warrant further investigation; and

Although reactive filtration can come close, only adsorption technologies have so far been demonstrated capable of meeting stringent indirect potable reuse requirements for effluent P concentrations of 10 μg/L-P.

Footnotes

Acknowledgment

The support of the Electric Power Research Institute (EPRI Award 00-10007028) for this work is gratefully acknowledged.

Author Disclosure Statement

No competing financial interests exist.

References

Adeniyi

O.M.

, Azimov

, and Burluka

(2018). Algae biofuel: Current status and future applications. Renew. Sustain. Energy Rev. 90, 316.

Adnan

, Dastur

, Mavinic

D.S.

, and Koch

F.A.

(2004). Preliminary investigation into factors affecting controlled struvite crystallization at the bench scale. J. Environ. Eng. Sci. 3, 195.

Adnan

, Mavinic

D.S.

, and Koch

F.A.

(2003). Pilot-scale study of phosphorus recovery through struvite crystallization–examining the process feasibility. J. Environ. Eng. Sci. 2, 315.

Antunes

, Jacob

M.V.

, Brodie

, and Schneider

P.A.

(2018). Isotherms, kinetics and mechanism analysis of phosphorus recovery from aqueous solution by calcium-rich biochar produced from biosolids via microwave pyrolysis. Environ. Chem. Eng. 6, 395.

Anupama, and Ravindra

(2000). Value-added food: Single cell protein. Biotechnol. Adv. 18, 459.

Awual

M.R.

, Jyo

, Ihara

, Seko

, Tamada

, and Lim

K.T.

(2011). Enhanced trace phosphate removal from water by zirconium(IV) loaded fibrous adsorbent. Water Res. 45, 4592.

Baettens

(2001). Enhanced Biological Phosphorous Removal: Modelling and Experimental Design. Thesis, Gent, Univ., Diss., 2000/2001, HBZ-ID: HT012934128. Ghent University, Belgium.

Barbera

, Bertucco

, and Kumar

(2018) Nutrients recovery and recycling in algae processing for biofuels production. Renew. Sustain. Energy Rev. 90, 28.

Barca

, Gérente

, Meyer

, Chazarenc

, and Andrès

(2012). Phosphate removal from synthetic and real wastewater using steel slags produced in Europe. Water Res. 6, 2.

10.

Barca

, Meyer

, Liira

, Drissen

, Comeau

, Andrès

, and Chazarenc

(2014). Steel slag filters to upgrade phosphorus removal in small wastewater treatment plants: Removal mechanisms and performance. Ecol. Eng. 68, 214.

11.

Başakçilardan-Kabakci

, İpekoğlu

A.N.

, and Talinli

(2007). Recovery of ammonia from human urine by stripping and absorption. Environ. Eng. Sci. 24, 615.

12.

Battistoni

, Fava

, Pavan

, Musacco

, and Cecchi

(1997). Phosphate removal in anaerobic liquors by struvite crystallization without addition of chemicals: Preliminary results. Water Res. 31, 2925.

13.

Becker

E.W.

(2007). Micro-algae as a source of protein. Biotechnol. Adv. 25, 207.

14.

Berry

J.P.

, Gantar

, Perez

M.H.

, Berry

, and Noriega

F.G.

(2008). Cyanobacterial toxins as allelochemicals with potential applications as algaecides, herbicides and insecticides. Mar. Drugs. 6, 117.

15.

Bertanza

, Pedrazzani

, Manili

, and Menoni

(2013). Bio-P release in the final clarifiers of a large WWTP with co-precipitation: Key factors and troubleshooting. Chem. Eng. J. 230, 195.

16.

Blaney

L.M.

, Cinar

, and SenGupta

A.K.

(2007). Hybrid anion exchanger for trace phosphate removal from water and wastewater. Water Res. 41, 1603.

17.

Bonmatı

, and Flotats

(2003). Air stripping of ammonia from pig slurry: Characterisation and feasibility as a pre- or post-treatment to mesophilic anaerobic digestion. Waste Manag. 23, 261.

18.

Boxman

S.E.

, Nystrom

, Ergas

S.J.

, Main

K.L.

, and Trotz

M.A.

(2018). Evaluation of water treatment capacity, nutrient cycling, and biomass production in a marine aquaponic system. Ecol. Eng. 120, 299.

19.

Bridger

G.L.

, Salutsky

M.L.

, and Starostka

R.W.

(1962). Micronutrient sources, metal ammonium phosphates as fertilizers. J. Agric. Food Chem. 10, 181.

20.

Britton

, Koch

F.A.

, Mavinic

D.S.

, Adnan

, Oldham

W.K.

, and Udala

(2005). Pilot-scale struvite recovery from anaerobic digester supernatant at an enhanced biological phosphorus removal wastewater treatment plant. J. Environ. Eng. Sci. 4, 265.

21.

Cade-Menun

B.J.

, and Paytan

(2010). Nutrient temperature and light stress alter phosphorus and carbon forms in culture-grown algae. Mar. Chem. 121, 27.

22.

Chen

(2004). Electrochemical technologies in wastewater treatment. Sep. Purif. Technol. 38, 11.

23.

Cheung

K.C.

, Chu

L.M.

, and Wong

M.H.

(1997). Ammonia stripping as a pretreatment for landfill leachate. Water. Air. Soil Pollut. 94, 209.

24.

Christenson

, and Sims

(2011). Production and harvesting of microalgae for wastewater treatment, biofuels, and bioproducts. Biotechnol. Adv. 29, 686.

25.

Clark

, Stephenson

, and Pearce

P.A.

(1997). Phosphorus removal by chemical precipitation in a biological aerated filter. Water Res. 31, 2557.

26.

Cordell

, Drangert

, and White

(2009). The story of phosphorus: Global food security and food for thought. Glob. Environ. Chang. 19, 292.

27.

Cordell

, Neset

T.S.

, and Prior

(2012). The phosphorus mass balance: Identifying ‘hotspots’ in the food system as a roadmap to phosphorus security. Curr. Opin. Biotechnol. 23, 839.

28.

Culp

R.L.

, Wesner

G.M.

, and Culp

G.L.

(1978). Handbook of Advanced Wastewater Treatment, Second Edition. New York: NY, Van Nostrand Reinhold.

29.

Darestani

, Haigh

, Couperthwaite

S.J.

, Millar

G.J.

, and Nghiem

L.D.

(2017). Hollow fibre membrane contactors for ammonia recovery: Current status and future developments. J. Environ. Chem. Eng. 5, 1349.

30.

Dastur

M.B.

(2001). Investigation into the factors affecting controlled struvite crystallization at the bench-scale; University of Brtish Colombia. Retrieved from https://open.library.ubc.ca/cIRcle/collections/ubctheses/831/items/1.0063895#downloadfiles (accessed January 16, 2018).

31.

de Vicente

, Merino-Martos

, Cruz-Pizarro

, and de Vicente

(2010). On the use of magnetic nano and microparticles for lake restoration. J. Hazard. Mater. 181, 375.

32.

De-Bashan

L.E.

, and Bashan

(2004). Recent advances in removing phosphorus from wastewater and its future use as fertilizer (1997–2003). Water Res. 38, 4222.

33.

Deĝermenci

, Ata

O.N.

, and Yildiz

(2012). Ammonia removal by air stripping in a semi-batch jet loop reactor. J. Ind. Eng. Chem. 18, 399.

34.

Demir

, Günay

, and Debik

(2002). Ammonium removal from aqueous solution by ion-exchange using packed bed natural zeolite. Water SA. 28, 329.

35.

Desloover

, Abate Woldeyohannis

, Verstraete

, Boon

, and Rabaey

(2012). Electrochemical resource recovery from digestate to prevent ammonia toxicity during anaerobic digestion. Environ. Sci. Technol. 46, 12209.

36.

Ding

, Wu

, Deng

, and Zhang

(2012). Adsorptive characteristics of perchlorate from aqueous solutions by MIEX resin. J. Colloid Interface Sci. 376, 224.

37.

Diwani

G. E

, l Rafie

S. E

, l Ibiari

N.N.

, El El-Aila

H.I.

(2007). Recovery of ammonia nitrogen from industrial wastewater treatment as struvite slow releasing fertilizer. Desalination. 214, 200.

38.

Dixon

J.B.

, and Weed

S.B.

(1989). Minerals in Soil Environments. Second Edition; Madison: WI, Soil Science Society of America.

39.

Doshi

, Pascoe

, Coglan

, and Rainey

T.J.

(2016). Economic and policy issues in the production of algae-based biofuels: A review. Renew. Sustain. Energy Rev. 64, 329.

40.

Drenkova-Tuhtan

, Mandel

, Paulus

, Meyer

, Hutter

, Gellermann

, Sextl

, et al. (2013). Phosphate recovery from wastewater using engineered superparamagnetic particles modified with layered double hydroxide ion exchangers. Water Res. 47, 5670.

41.

Drenkova-Tuhtan

, Schneider

, Franzreb

, Meyer

, Gellermann

, Sextl

, Mandel

, et al. (2017a). Pilot-scale removal and recovery of dissolved phosphate from secondary wastewater effluents with reusable ZnFeZr adsorbent @ Fe3O4/SiO2 particles with magnetic harvesting. Water Res. 109, 77.

42.

Drenkova-tuhtan

, Schneider

, Meyer

, Franzreb

, Gellermann

, Mandel

, and Steinmetz

(2017b) Polishing of secondary wastewater effluents through elimination and recovery of dissolved phosphorus with reusable magnetic microsorbents. WEF Nutr. Symp., 2017, 169.

43.

Dube

P.J.

, Vanotti

M.B.

, Szogi

A.A.

, and García-González

M.C.

(2016). Enhancing recovery of ammonia from swine manure anaerobic digester effluent using gas-permeable membrane technology. Waste Manag. 49, 372.

44.

Egle

, Rechberger

, and Zessner

(2015). Overview and description of technologies for recovering phosphorus from municipal wastewater. Resour. Conserv. Recycl. 105, 325.

45.

El-bourawi

M.S.

, Khayet

, Ma

, Ding

, Li

, and Zhang

(2007). Application of vacuum membrane distillation for ammonia removal. J. Memb. Sci. 301, 200.

46.

Escudero

, Blanco

, Lacalle

, and Pinto

(2015). Struvite precipitation for ammonium removal from anaerobically treated ef fl uents. J. Environ. Chem. Eng. 3, 413.

47.

Fanta

S.E.

, Hill

W.R.

, Smith

T.B.

, and Roberts

B.J.

(2010). Applying the light: Nutrient hypothesis to stream periphyton. Freshw. Biol. 55, 931.

48.

Fattah

K.P.

, Zhang

, Mavinic

D.S.

, and Koch

F.A.

(2008). Application of carbon dioxide stripping for struvite crystallization—I: Development of a carbon dioxide stripper model to predict CO2 removal and pH changes. J. Environ. Eng. Sci. 7, 345.

49.

Fattah

K.P.

, Zhang

, Mavinic

D.S.

, and Koch

F.A.

(2010). Use of carbon dioxide stripping for struvite crystallization to save caustic dosage: Performance at pilotscale operation. Can. J. Civ. Eng. 37, 1271.

50.

Florida Everglades Forever Act. (1994). Florida Everglades Forever Act. Florida State Legislature, Tallahassee.

51.

Galloway

J.N.

, and Cowling

E.B.

(2002). Reactive nitrogen and the world: 200 years of change. AMBIO A J. Hum. Environ. 31, 64.

52.

Garbisu

, Alkorta

, Llama

M.J.

, and Serra

J.L.

(2000). Immobilisation of living microalgae and their use for inorganic nitrogen and phosphorus removal from water. In Olguín

, G. Sanchez & E. Hernández Alvidrez, Eds., Environmentol Biotechnology and Cleaner Bioprocesses. London, Taylor & Francis, pp. 107–121.

53.

Garcia-González

M.C.

, and Vanotti

M.B.

(2015). Recovery of ammonia from swine manure using gas-permeable membranes: Effect of waste strength and pH. Waste Manag. 38, 455.

54.

Gassie

L.W.

, Englehardt

J.D.

, Wang

, Brinkman

, Garland

, Gardinali

, and Guo

(2016). Mineralizing urban net-zero water treatment: Phase II field results and design recommendations. Water Res. 105, 496.

55.

Gildemyn

, Luther

A.K.

, Andersen

S.J.

, Desloover

, and Rabaey

(2015). Electrochemically and bioelectrochemically induced ammonium recovery. J. Vis. Exp. 95, 52405.

56.

Gupta

V.K.

, Sadegh

, Yari

, Shahryari Ghoshekandi

, Maazinejad

, and Chahardori

(2015). Removal of ammonium ions from wastewater A short review in development of efficient methods. Glob. J. Environ. Sci. Manag. 1, 149.

57.

Gustafsson

J.P.

, Renman

, and Poll

(2008). Phosphate removal by mineral-based sorbents used in filters for small-scale wastewater treatment. Water Res. 42, 189.

58.

Gustin

, and Marinsek-Logar

(2010). Effect of pH, temperature and air flow rate on the continuous ammonia stripping of the anaerobic digestion effluent. Process Saf. Environ. Prot. 9, 61.

59.

Haas

D.W.

, De Wentzel

M.C.

, and Ekama

G.A.

(2000). The use of simultaneous chemical precipitation in modified activated sludge systems exhibiting biological excess phosphate removal Part 1: Literature review. Water SA. 26, 439.

60.

Harnisch

, Warmbier

, Schneider

, and Schröder

(2009). Modeling the ion transfer and polarization of ion exchange membranes in bioelectrochemical systems. Bioelectrochemistry, 75, 136.

61.

, Huang

, Manohar

A.K.

, and Mansfeld

(2008). Effect of electrolyte pH on the rate of the anodic and cathodic reactions in an air-cathode microbial fuel cell. Bioelectrochemistry. 74, 78.

62.

Hermassi

, Dosta

, Valderrama

, Licon

, Moreno

, Querol

, Batis

N.H.

, et al. (2018). Simultaneous ammonium and phosphate recovery and stabilization from urban sewage sludge anaerobic digestates using reactive sorbents. Sci. Total Environ. 630, 781.

63.

Hermassi

, Valderrama

, Moreno

, Font

, Querol

, Batis

, and Cortina

J.L.

(2016). Powdered Ca-activated zeolite for phosphate removal from treated waste-water. J. Chem. Technol. Biotechnol. 91, 1962.

64.

Huang

, Xiao

, Yan

, and Yang

(2010). Ammonium removal from aqueous solutions by using natural Chinese (Chende) zeolite as adsorbent. J. Hazard. Mater. 175, 247.

65.

Huang

, Zhang

, Zhao

, Zhang

, and Gao

(2017). Comparison investigation on phosphate recovery from sludge anaerobic supernatant using the electrocoagulation process and chemical precipitation. J. Clean. Prod. 141, 429.

66.

Hug

, and Udert

K.M.

(2012). Struvite precipitation from urine with electrochemical magnesium dosage. Water Res. 47, 289.

67.

Hultman

, and Löwén

(2001). Combined phosphorus removal and recovery. Polish-Swedish Semin., 11–18; Stockholm, Sweden.

68.

Ieropoulos

, Greenman

, and Melhuish

(2012). Urine utilisation by microbial fuel cells; energy fuel for the future. Phys. Chem. Chem. Phys. 14, 94.

69.

Ippersiel

, Mondor

, Lamarche

, Tremblay

, Dubreuil

, and Masse

(2012). Nitrogen potential recovery and concentration of ammonia from swine manure using electrodialysis coupled with air stripping. J. Environ. Manage. 95, S165.

70.

Ishii

S.K.L.

, and Boyer

T.H.

(2015). Life cycle comparison of centralized wastewater treatment and urine source separation with struvite precipitation: Focus on urine nutrient management. Water Res. 79, 88.

71.

Ishii

S.K.L.

, and Boyer

T.H.

(2016). Student support and perceptions of urine source separation in a university community. Water Res. 100, 146.

72.

Ishiwata

, Miura

, Hosomi

, Shimizu

, Ito

, and Yoda

(2010). Removal and recovery of phosphorus in wastewater by superconducting High Gradient Magnetic Separation with ferromagnetic adsorbent. Phys. C Supercond. Appl. 470, 1818.

73.

Ito

, Nishimura

, and Miura

(2009). Removal and recycle of phosphate from treated water of sewage plants with zirconium ferrite adsorbent by high gradient magnetic separation. J. Phys. Conf. Ser. 156, 012033.

74.

Jaffer

, Clark

T.A.

, Pearce

, and Parsons

S.A.

(2002). Potential phosphorus recovery by struvite formation. Water Res. 36, 1834.

75.

Jena

, Vaidyanathan

, Chinnasamy

, and Das

K.C.

(2011). Evaluation of microalgae cultivation using recovered aqueous co-product from thermochemical liquefaction of algal biomass. Bioresour. Technol. 102, 3380.

76.

Jetten

M.S.M.

, Strous

, Pas-Schoonen

K.T.

, Van De Schalk

, Dongen

U.G.J

. M., Van Graaf

A.A.

, Van De Logemann

, et al. (1998). The anaerobic oxidation of ammonium. FEMS Microbiol. Rev. 22, 421.

77.

Jia

, Zhang

, Krampe

, Muster

, Gao

, Zhu

, and Jin

(2017). Applying a chemical equilibrium model for optimizing struvite precipitation for ammonium recovery from anaerobic digester effluent. J. Clean. Prod. 147, 297.

78.

Jiang

, Zhang

, Zhao

, Frear

, and Chen

(2010). Chapter 8: Integrated Ammonia Recovery Technology in Conjunction with Dairy Anaerobic Digestion. 1. Final report of Climate Friendly Farming project conducted at Center of Sustaining Agriculture and Natural Resources at Washington State University (http://csanr.wsu.edu/program-areas/climate-friendly-farming/climate-friendly-farming-final-report/) (accessed March 14, 2018)

79.

Johir

M.A.H.

, George

, Vigneswaran

, Kandasamy

, and Grasmick

(2011). Removal and recovery of nutrients by ion exchange from high rate membrane bio-reactor (MBR) ef fl uent. DES. 275, 197.

80.

Jönsson

, Stintzing

A.R.

, Vinnerås

, and Salomon

(2004). Guidelines on the Use of Urine and Faeces in Crop Production. Stockholm, Sweden. Retrieved from www.ecosanres.org

81.

Kappel

, Yasadi

, Temmink

, Metz

S.J.

, Kemperman

A.J.B.

, Nijmeijer

, Zwijnenburg

, et al. (2013). Electrochemical phosphate recovery from nanofiltration concentrates. Sep. Purif. Technol. 120, 437.

82.

Kaschka

, and Weyrer

(1999). Phostrip handbook, Fourth Edi. Simbach: Phostrip Abwassertechnik GmbH.

83.

Katsuura

(1998). Phosphate recovery from sewage by granule forming process (full scale struvite recovery from a sewage works at Shimane Prefecture, Japan). In International Conference on Phosphorus Recovery From Sewage and Animal Waste. Warwick University, U.K.

84.

Kebede-Westhead

, Pizarro

, and Mulbry

W.W.

(2006). Treatment of swine manure effluent using freshwater algae: Production, nutrient recovery, and elemental composition of algal biomass at four effluent loading rates. J. Appl. Phycol. 18, 41.

85.

Kelly

P.T.

, and He

(2014). Nutrients removal and recovery in bioelectrochemical systems: A review. Bioresour. Technol. 153, 351.

86.

Kim

J.O.

, and Chung

(2014). Implementing chemical precipitation as a pretreatment for phosphorus removal in membrane bioreactor-based municipal wastewater treatment plants. KSCE J. Civil Eng. 18, 956.

87.

Klimeski

(2007). Phosphorus Recovery from Black Water by Chemical Precipitation. Göteborg, Sweden: Chalmers University of Technology.

88.

Koon

J.H.

, and Kaufman

W.J.

(2018). Ammonia removal from municipal wastewaters by ion exchange. Water Pollut. Control Fed. 47, 448.

89.

Kruk

D.J.

, Elektorowicz

, and Oleszkiewicz

J.A.

(2014). Struvite precipitation and phosphorus removal using magnesium sacrificial anode. Chemosphere. 101, 28.

90.

Kumar

, and Pal

(2015). Assessing the feasibility of N and P recovery by struvite precipitation from nutrient-rich wastewater: A review. Environ. Sci. Pollut. Res. 22, 17453.

91.

Kuntke

, Geleji

, Bruning

, Zeeman

, Hamelers

H.V.M.

, and Buisman

C.J.N.

(2011). Effects of ammonium concentration and charge exchange on ammonium recovery from high strength wastewater using a microbial fuel cell. Bioresour. Technol. 102, 4376.

92.

Kuntke

, Rodríguez Arredondo

, Widyakristi

, Heijne, A. Ter Sleutels

T.H.J

.A., Hamelers

H.V.M.

, and Buisman

C.J.N.

(2017). Hydrogen gas recycling for energy efficient ammonia recovery in electrochemical systems. Environ. Sci. Technol. 51, 3110.

93.

Kuntke

, Śmiech

K.M.

, Bruning

, Zeeman

, Saakes

, Sleutels

T.H.J

.A., Hamelers

H.V.M.

, et al. (2012). Ammonium recovery and energy production from urine by a microbial fuel cell. Water Res. 46, 2627.

94.

Kuzawa

, Jung

, Kiso

, Yamada

, Nagai

, and Lee

(2006). Phosphate removal and recovery with a synthetic hydrotalcite as an adsorbent. Chemosphere. 62, 45.

95.

Lahav

, Telzhensky

, Zewuhn

, Gendel

, Gerth

, Calmano

, and Birnhack

(2013). Struvite recovery from municipal-wastewater sludge centrifuge supernatant using seawater NF concentrate as a cheap Mg(II) source. Sep. Purif. Technol. 108, 103.

96.

Laureni

, Palatsi

, Llovera

, and Bonmatí

(2013). Influence of pig slurry characteristics on ammonia stripping efficiencies and quality of the recovered ammonium-sulfate solution. J. Chem. Technol. Biotechnol. 88, 1654.

97.

Lee

M.S.

, Drizo

, Rizzo

D.M.

, Druschel

, Hayden

, and Twohig

(2010). Evaluating the efficiency and temporal variation of pilot-scale constructed wetlands and steel slag phosphorus removing filters for treating dairy wastewater. Water Res. 44, 4077.

98.

Lei

, Sugiura

, Feng

, and Maekawa

(2007). Pretreatment of anaerobic digestion effluent with ammonia stripping and biogas purification. J. Hazard. Mater. 145, 391.

99.

, Zhu

, Ren

, Cao

, and Song

(2011). Application of modi fi ed zeolite for ammonium removal from drinking water. Desalination. 271, 295.

100.

, and Chen

(2018). Factors Affecting the Effectiveness of Bioelectrochemical System Applications: Data Synthesis and Meta-Analysis (Figure 1). Batteries. 4, 34.

101.

Liberti

, Boari

, and Passino

(1978). Phosphates and ammonia recovery from secondary effluents by selective ion exchange with production of a slow-release fertilizer. Water Res. 13, 65.

102.

Liberti

, Boari

, Petruzzelli

, and Passino

(1980). Nutrient removal and recovery from wastewater by ion exchange. Water Res. 15, 337.

103.

Lin

, Wan

, Lee

D.J.

, Lei

, and Liu

(2014). Ammonium assists orthophosphate removal from high-strength wastewaters by natural zeolite. Sep. Purif. Technol. 133, 351.

104.

Liu

, Giannis

, Zhang

, Chang

V.W.C.

, and Wang

J.Y.

(2013a). Characterization of induced struvite formation from source-separated urine using seawater and brine as magnesium sources. Chemosphere. 93, 2738.

105.

Liu

, Giannis

, Zhang

, Chang

, and Wang

(2015). Air stripping process for ammonia recovery from source-separated urine: Modeling and optimization. Chem. Technol. Biotechnol. 90, 2208.

106.

Liu

, Kumar

, Kwag

J.H.

, and Ra

(2013b). Magnesium ammonium phosphate formation, recovery and its application as valuable resources: A review. Chem. Technol. Biotechnol. 88, 181.

107.

Liu

, Qin

, Luo

, He

, and Qiao

(2016). Understanding Ammonium Transport in Bioelectrochemical Systems towards its Recovery. Sci. Rep. 6, 22547.

108.

Liu

Z.G.

, Zhao

Q.L.

, Wang

, Qiu

, Li

, and Wang

J.F.

(2008). Comparison between complete and partial recovery of N and P from stale human urine with MAP crystallization. Environ. Eng. Sci. 7, 223.

109.

Luther

A.K.

, Desloover

, Fennell

D.E.

, and Rabaey

(2015). Electrochemically driven extraction and recovery of ammonia from human urine. Water Res. 87, 367.

110.

, Li

, Yue

, Gao

, Li

, Xu

, and Zhong

(2011). Adsorption removal of ammonium and phosphate from water by fertilizer controlled release agent prepared from wheat straw. Chem. Eng. J. 171, 1209.

111.

Madeira

M.S.

, Cardoso

, Lopes

P.A.

, Coelho

, Afonso

, Bandarra

N.M.

, and Prates

J.A.M.

(2017). Microalgae as feed ingredients for livestock production and meat quality: A review. Livest. Sci. 205, 111.

112.

Malekian

, Abedi-Koupai

, and Eslamian

S.S.

(2011). Influences of clinoptilolite and surfactant-modified clinoptilolite zeolite on nitrate leaching and plant growth. J. Hazard. Mater. 185, 970.

113.

Marañón

, Ulmanu

, Fernández

, Anger

, and Castrillón

(2006). Removal of ammonium from aqueous solutions with volcanic tuff. J. Hazard. Mater. 137, 1402.

114.

Mariakakis

, Bilbao

, and Egner

(2017). Effect of Pilot Plant Configuration on the Electrochemical Precipitation for Phosphorus recovery with the ePhos Technology. WEF Nutr. Symp. 2017, 537.

115.

Martin

B.D.

, Parsons

S.A.

, and Jefferson

(2009). Removal and recovery of phosphate from municipal wastewaters using a polymeric anion exchanger bound with hydrated ferric oxide nanoparticles. Water Sci. Technol. 60, 2637.

116.

Matassa

, Batstone

D.J.

, Hülsen

, Schnoor

, and Verstraete

(2015). Can direct conversion of used nitrogen to new feed and protein help feed the world? Environ. Sci. Technol. 49, 5247.

117.

Maurer

, Schwegler

, and Larsen

T.A.

(2003). Nutrients in urine: Energetic aspects of removal and recovery. Water Sci. Technol. 48, 37.

118.

Mavinic

D.S.

, Koch

F.A.

, Huang

, and Lo

K.V.

(2007). Phosphorus recovery from anaerobic digester supernatants using a pilot-scale struvite crystallization process. Environ. Eng. Sci. 6, 561.

119.

Mehta

C.M.

, Khunjar

W.O.

, Nguyen

, Tait

, and Batstone

D.J.

(2015). Technologies to Recover Nutrients from Waste Streams: A Critical Review. Crit. Rev. Environ. Sci. Technol. 45, 385.

120.

Merino-Jimenez

, Celorrio

, Fermin

D.J.

, Greenman

, and Ieropoulos

(2017). Enhanced MFC power production and struvite recovery by the addition of sea salts to urine. Water Res. 109, 46.

121.

Merino-Martos

, Vicente, J. de Cruz-Pizarro

, and Vicente, I. de. (2011). Setting up High Gradient Magnetic Separation for combating eutrophication of inland waters. J. Hazard. Mater. 186, 2068.

122.

Mino

, Loosdrecht

M.C.M.

, and Van Heijnen

J.J.

(1998). Microbiology and biochemistry of the enhanced biological phosphate removal process. Water Res. 32, 3193.

123.

, Ko

, Liira

, and Kirsima

(2010). Phosphorus removal using Ca-rich hydrated oil shale ash as filter material–The effect of different phosphorus loadings and wastewater compositions. Water Res. 44, 5232.

124.

Mohd

, Kian

, Ibrahim

, Abdul

, and Ahmad

(2010). Kinetic and equilibrium studies of the removal of ammonium ions from aqueous solution by rice husk ash-synthesized zeolite Y and powdered and granulated forms of mordenite. J. Hazard. Mater. 174, 380.

125.

Moradi

, and Zare

(2013). Adsorption of Ammonium Ion by Multi-walled Carbon Nanotube: Kinetics and Thermodynamic Studies. Fullerenes Nanotub. Carbon Nanostructures. 21, 449.

126.

Morse

G.K.

, Brett

S.W.

, Guy

J.A.

, and Lester

J.N.U.

(1998). Review: Phosphorus removal and recovery technologies. Sci. Total Environ. 212, 69.

127.

Mulbry

, Westhead

E.K.

, Pizarro

, and Sikora

(2005). Recycling of manure nutrients: Use of algal biomass from dairy manure treatment as a slow release fertilizer. Bioresour. Technol. 96, 451.

128.

Mulkerrins

, Dobson

A.D.W.

, and Colleran

(2004). Parameters affecting biological phosphate removal from wastewaters. Environ. Int. 30, 249.

129.

Münch

E.V.

, and Barr

(2001). Controlled struvite crystallisation for removing phosphorus from anaerobic digester sidestreams. Water Res. 35, 151.

130.

Muster

T.H.

, and Jermakka

(2018). Electrochemically-assisted ammonia recovery from wastewater using a floating electrode. Water Sci. Technol. 7, 1804.

131.

Muster

T.H.

, Douglas

G.B.

, Sherman

, Seeber

, Wright

, and Güzükara

(2013). Towards effective phosphorus recycling from wastewater: Quantity and quality. Chemosphere. 91, 676.

132.

Nancharaiah

Y.V.

, Venkata Mohan

, and Lens

P.N.L.

(2016). Recent advances in nutrient removal and recovery in biological and bioelectrochemical systems. Bioresour. Technol. 215, 173.

133.

Newcombe

R.L.

, Hart

B.K.

, and Möller

(2006). Arsenic Removal from Water by Moving Bed Active Filtration. J. Environ. Eng. 132, 5.

134.

Newcombe

R.L.

, Rule

R.A.

, Hart

B.K.

, and Moller

(2008). Phosphorus Removal from Municipal Wastewater by Hydrous Ferric Oxide Reactive Filtration and Coupled Chemically Enhanced Secondary Treatment: Part I—Performance. Water Environ. Res. 80, 238.

135.

Nilsson

, Lakshmanan

, and Renman

(2013). Efficacy of reactive mineral-based sorbents for phosphate, bacteria, nitrogen and TOC removal e Column experiment in recirculation batch mode. Water Res. 47, 5165.

136.

O'Neal

J.A.

, and Boyer

T.H.

(2013). Phosphate recovery using hybrid anion exchange: Applications to source-separated urine and combined wastewater streams. Water Res. 47, 5003.

137.

Parsons

S.A.

, and Smith

J.A.

(2008). Phosphorus Removal and Recovery from Municipal Wastewaters. Elements. 4, 109.

138.

Pikaar

, Rozendal

R.A.

, Rabaey

, and Yuan

(2013). In-situ caustic generation from sewage: The impact of caustic strength and sewage composition. Water Res. 47, 5828.

139.

Pikaar

, Rozendal

R.A.

, Yuan

, and Rabaey

(2011). Electrochemical caustic generation from sewage. Electrochem. Commun. 13, 1202.

140.

Pitman

A.R.

(1991). Design Considerations for Nutrient Removal Activated Sludge Plants. Water Sci. Technol. 23, 781.

141.

Pizarro

, Mulbry

, and Kebede-Westhead

(2002). Nitrogen and phosphorus removal rates using small algal turfs grown with dairy manure. J. Appl. Phycol. 14, 469.

142.

Powell

, Shilton

, Chisti

, and Pratt

(2009). Towards a luxury uptake process via microalgae—Defining the polyphosphate dynamics. Water Res. 43, 4207.

143.

Powell

, Shilton

A.N.

, Pratt

, and Chisti

(2008). Factors Influencing Luxury Uptake of Phosphorus by Microalgae in Waste Stabilization Ponds. Environ. Sci. Technol. Lett. 42, 5958.

144.

Puchongkawarin

, Gomez-mont

, Stuckey

D.C.

, and Chachuat

(2015). Optimization-based methodology for the development of wastewater facilities for energy and nutrient recovery. Chemosphere. 140, 150.

145.

Qin

, and He

(2014). Self-Supplied Ammonium Bicarbonate Draw Solute for Achieving Wastewater Treatment and Recovery in a Microbial Electrolysis Cell-Forward Osmosis-Coupled System. Environ. Sci. Technol. Lett. 1, 437.

146.

Qin

, Molitor

, Brazil

, Novak

J.T.

, and He

(2016). Recovery of nitrogen and water from landfill leachate by a microbial electrolysis cell-forward osmosis system. Bioresour. Technol. 200, 485.

147.

Quan

, Wang

, Zhao

, and Xiang

(2009). Air stripping of ammonia in a water-sparged aerocyclone reactor. J. Hazard. Mater. 170, 983.

148.

Ragheb

S.M.

(2013). Phosphate removal from aqueous solution using slag and fly ash. HBRC J. 9, 270.

149.

Reed Business Information Limited. (2018). Companies advance algae biofuels programme. Focus Catal. 2018, 4.

150.

Renman

, and Renman

(2010). Long-term phosphate removal by the calcium-silicate material Polonite in wastewater filtration systems. Chemosphere, 79, 659.

151.

Renman

, Cucarella

, Mazurek

, Zaleski

, and Kopec

(2009). Effect of Polonite used for phosphorus removal from wastewater on soil properties and fertility of a mountain meadow. Environ. Pollut. 157, 2147.

152.

Richardson

A.E.

(2001). Prospects for using soil microorganisms to improve the acquisition of phosphorus by plants. Aust. J. Plant Physiol. 28, 897.

153.

Richmond

(2004). Handbook of Microalgal Culture: Biotechnology and Applied Phycology. Oxford, United Kingdom: Blackwell Science Ltd.

154.

Ronteltap

, Maurer

, and Gujer

(2007). Struvite precipitation thermodynamics in source-separated urine. Water Res. 41, 977.

155.

Ronteltap

, Maurer

, Hausherr

, and Gujer

(2010). Struvite precipitation from urine—Influencing factors on particle size. Water Res. 44, 2038.

156.

Rothrock

M.J.

, Szögi

A.A.

, and Vanotti

M.B.

(2013). Recovery of ammonia from poultry litter using flat gas permeable membranes. Waste Manag. 33, 1531.

157.

Rozendal

R.A.

, Hamelers

H.V.M.

, and Buisman

C.J.N.

(2006). Effects of Membrane Cation Transport on pH and Microbial Fuel. Environ. Sci. Technol. 40, 5206.

158.

Rubio-Rincón

F.J.

, Lopez-Vazquez

C.M.

, Ronteltap

, and Brdjanovic

(2014). Seawater for phosphorus recovery from urine. Desalination, 348, 49.

159.

Ryu

H.D.

, Lim

C.S.

, Kang

M.K.

, and Lee

S.I.

(2012). Evaluation of struvite obtained from semiconductor wastewater as a fertilizer in cultivating Chinese cabbage. J. Hazard. Mater. 221, 248.

160.

Saltali

, Sari

, and Aydin

(2007). Removal of ammonium ion from aqueous solution by natural Turkish (Yıldızeli) zeolite for environmental quality. J. Hazard. Mater. 141, 258.

161.

Saracco

, and Genon

(1994). High temperature ammonia stripping and recovery from process liquid wastes. J. Hazard. Mater. 37, 191.

162.

Sengupta

, and Pandit

(2011). Selective removal of phosphorus from wastewater combined with its recovery as a solid-phase fertilizer. Water Res. 45, 3318.

163.

Sengupta

, Nawaz

, and Beaudry

(2015). Nitrogen and Phosphorus Recovery from Wastewater. Curr. Pollut. Rep. 1, 155.

164.

Shilton

A.N.

, Elmetri

, Drizo

, Pratt

, Haverkamp

R.G.

, and Bilby

S.C.

(2006). Phosphorus removal by an ‘active’ slag filter–a decade of full scale experience. Water Res. 40, 113.

165.

Shilton

A.N.

, Powell

, and Guieysse

(2012). Plant based phosphorus recovery from wastewater via algae and macrophytes. Curr. Opin. Biotechnol. 23, 884.

166.

Shpirt

(1981). Role of hydrodynamic factors in ammonia desorption by diffused aeration. Water Res. 15, 739.

167.

Shu

, Schneider

, Jegatheesan

, and Johnson

(2006). An economic evaluation of phosphorus recovery as struvite from digester supernatant. Bioresour. Technol. 97, 2211.

168.

Smith

B.E.

(2002). Nitrogenase Reveals Its Inner Secrets. Science (80-.). 297, 1654.

169.

Smith

D.P.

, and Smith

N.T.

(2015). Anaerobic-ion exchange (AN-IX) process for local-scale nitrogen recovery from wastewater. Bioresour. Technol. 196, 324.

170.

Smoczyński

, Kalinowski

, Ratnaweera

, Kosobucka

, Trifescu

, and Pieczulis-Smoczyńska

(2017). Electrocoagulation of municipal wastewater—a pilot-scale test. Desalin. Water Treat. 72, 162.

171.

Solovchenko

, Verschoor

A.M.

, Jablonowski

N.D.

, and Nedbal

(2016). Phosphorus from wastewater to crops: An alternative path involving microalgae. Biotechnol. Adv. 34, 550.

172.

Srinath

E.G.

, and Loehr

R.C.

(1974). Ammonia desorption by diffused aeration. Water Pollut. Control Fed. 46, 1939.

173.

Stolzenburg

, Capdevielle

, Teychené

, and Biscans

(2015). Struvite precipitation with MgO as a precursor: Application to wastewater treatment. Chem. Eng. Sci. 133, 9.

174.

Stratful

, Brett

, Scrimshaw

M.B.

, and Lester

J.N.

(1999). Biological Phosphorus Removal, Its Role in Phosphorus Recycling. Environ. Technol. 20, 681.

175.

, Song

, Zhang

, Su

, Cheng

, and Chen

(2017). Progress of microalgae biofuel's commercialization. Renew. Sustain. Energy Rev. 74, 402.

176.

Sutton

P.M.

, Melcer

, Schraa

O.J.

, and Togna

A.P.

(2011). Treating municipal wastewater with the goal of resource recovery. Water Sci. Technol. 63, 25.

177.

Talboys

P.J.

, Heppell

, Roose

, Healey

J.R.

, Jones

D.L.

, and Withers

P.J.A.

(2016). Struvite: A slow-release fertiliser for sustainable phosphorus management?. Plant Soil. 401, 109.

178.

Tchobanoglous

, Abu-Orf

, Burton

F.L.

, Bowden

, Stensel

H.D.

, Tsuchihashi

, Pfrang

, et al. (2014). Wastewater Engineering: Treatment and Resource Recovery, 5th ed.; New York: McGraw-Hill.

179.

Tchobanoglous

, Burton

F.L.

, Stensel

H.D.

, and N/A Metcalf & Eddy, Inc. (2003). Wastewater Engineering: Treatment and Reuse. 4th ed.; New York: McGraw-Hill.

180.

Tetra Tech, Inc. (2014). Lake Champlain Phosphorus Removal: Technologies and Cost for Point Source Phosphorus Removal. Denver, CO: Tetra Tech, Inc.

181.

Tian

, Wang

, Guan

, Li

, Wang

, Li

, Yu

, et al. (2016). Reverse osmosis brine for phosphorus recovery from source separated urine. Chemosphere. 165, 202.

182.

Udert

K.M.

, Buckley

C.A.

, Wachter

, McArdell

C.S.

, Kohn

, Strande

, Zollig

, et al. (2015). Technologies for the treatment of source-separated urine in the eThekwini Municipality. Water SA,, 41, 212.

183.

USEPA. (1971). Wastewater Ammonia Removal by Ion Exchange. United States Environ. Prot. Agency, Water Pollut. Control Res. Ser. Richland, Washington.

184.

Vanotti

M.B.

, and Szogi

A.A.

(2010). Removal and Recovery of Ammonia from Liquid Manure Using Gas-Permeable Membranes. ASABE Meet. Present. Peper # 1008376.

185.

Vanotti

M.B.

, Dube

P.J.

, Szogi

A.A.

, and García-González

M.C.

(2017). Recovery of ammonia and phosphate minerals from swine wastewater using gas-permeable membranes. Water Res. 112, 137.

186.

Vanotti

M.B.

, Szogi

A.A.

, and Rothrock Jr

M.J.

(2009). Removal and Recovery of Ammonia From Liquid Swine Manure and Poultry Litter Using Gas Permeable Membranes. USDA. http://ramiran.uvlf.sk/ramiran2010/docs/Ramiran2010_0190_final.pdf (accessed March 31, 2017).

187.

Vassileva

, and Voikova

(2009). Investigation on natural and pretreated Bulgarian clinoptilolite for ammonium ions removal from aqueous solutions. J. Hazard. Mater. 170, 948.

188.

Venkatesan

A.K.

, Hamdan

A.-H.M.

, Chavez

V.M.

, Brown

J.D.

, and Halden

R.U.

(2015). Mass Balance Model for Sustainable Phosphorus Recovery in a US Wastewater Treatment Plant. J. Environ. Qual. 45, 84.

189.

Vohla

, Kõiv

, Bavor

H.J.

, Chazarenc

, and Mander, Ü. (2011). Filter materials for phosphorus removal from wastewater in treatment wetlands—A review. Ecol. Eng. 37, 70.

190.

Wang

, Lü

, Gao

, Xu

, Zhang

, Bai

, Liu

, et al. (2014). Highly efficient adsorption of ammonium onto palygorskite nanocomposite and evaluation of its recovery as a multifunctional slow-release fertilizer. Chem. Eng. J. 252, 404.

191.

Whitelaw

M.A.

(1999). Growth Promotion of Plants Inoculated with Phosphate-Solubilizing Fungi. Adv. Agron. 69, 99.

192.

Widanarni, Ekasari

, and Maryam

(2012). Evaluation of Biofloc Technology Application on Water Quality and Production Performance of Red Tilapia Oreochromis sp. Cultured at Different Stocking Densities. HAYATI J. Biosci., 19, 73.

193.

Williams

A.T.

, Zitomer

D.H.

, and Mayer

B.K.

(2015). Ion exchange-precipitation for nutrient recovery from dilute wastewater. Environ. Sci. Water Res. Technol. 1, 832.

194.

Wilsenach

J.A.

, and van Loosdrecht

M.C.M.

(2004). Effects of Separate Urine Collection on Advanced Nutrient Removal Processes. Environ. Sci. Technol. 38, 1208.

195.

Wilsenach

J.A.

, Schuurbiers

C.A.H.

, and van Loosdrecht

M.C.M.

(2007). Phosphate and potassium recovery from source separated urine through struvite precipitation. Water Res. 41, 458.

196.

, and Modin

(2013). Ammonium recovery from reject water combined with hydrogen production in a bioelectrochemical reactor. Bioresour. Technol. 146, 530.

197.

, Wang

, Liu

, and Qian

(2011). Simultaneous removal of phosphorus and potassium from synthetic urine through the precipitation of magnesium potassium phosphate hexahydrate. Chemosphere. 84, 207.

198.

, Hao

, Guo

, Liu

, Li

, Liu

, Zhang

, et al. (2017). Insight into chemical phosphate recovery from municipal wastewater. Sci. Total Environ. 576, 159.

199.

Yetilmezsoy

, and Sapci-Zengin

(2009). Recovery of ammonium nitrogen from the effluent of UASB treating poultry manure wastewater by MAP precipitation as a slow release fertilizer. J. Hazard. Mater. 166, 260.

200.

Yoon

, Lim

J.-H.

, and Chung

H.-K.

(2008). Ammonia Removal Model Based on the Equilibrium and Mass Transfer Principles. Bull. Korean Chem. Soc. 29, 7.

201.

You

, Valderramaa

, and Cortina

J.L.

(2017). Simultaneous recovery of ammonium and phosphate from simulated treated wastewater effluents by activated calcium and magnesium zeolites. Chem. Technol. Biotechnol.

202.

, Kaksonen

A.H.

, and Cord-ruwisch

(2013). Ammonia recycling enables sustainable operation of bioelectrochemical systems. Bioresour. Technol. 143, 25.

203.

Yuan

, Pratt

, and Batstone

D.J.

(2012). Phosphorus recovery from wastewater through microbial processes. Curr. Opin. Biotechnol. 23, 878.

204.

Zamora

, Georgieva

, Salcedo

, Elzinga

, Kuntke

, and Buisman

C.J.N.

(2017). Long-term operation of a pilot-scale reactor for phosphorus recovery as struvite from source-separated urine. J. Chem. Technol. Biotechnol. 92, 1035.

205.

Zhang

, Li

, Ding

, Xu

, and Ren

(2011). Influences of initial pH on performance and anodic microbes of fed-batch microbial fuel cells. Chem. Technol. Biotechnol. 1226.

206.

Zhang

, Desmidt

, Looveren, A. van Pinoy

, Meesschaert

, and van der Bruggen

(2013). Phosphate Separation and Recovery from Wastewater by Novel Electrodialysis. Environ. Sci. Technol. 47, 5888.

207.

Zheng

, Xie

, and Wang

(2012). Rapid and wide pH-independent ammonium-nitrogen removal using a composite hydrogel with three-dimensional networks. Chem. Eng. J. 179, 90.

208.

Zou

, and Wang

(2016). Phosphorus removal and recovery from domestic wastewater in a novel process of enhanced biological phosphorus removal coupled with crystallization. Bioresour. Technol. 211, 87.