Improving wastewater management using free nitrous acid (FNA)
Haoran Duana,b, Shuhong Gaoc, Xuan Lia, Nur Hafizah Ab Hamidb, Guangming Jiangd, Min Zhenga, Xue Baia, Philip L. Bonda, Xuanyu Lua,b, Mariella M. Chisletta, Shihu Hua, Liu Yeb, Zhiguo Yuana,*
aAdvanced Water Management Centre, The University of Queensland, St Lucia, QLD 4072, Australia
bSchool of Chemical Engineering, The University of Queensland, St. Lucia, QLD 4072, Australia
cInstitute for Environmental Genomics, Department of Microbiology and Plant Biology, University of Oklahoma , Norman , Oklahoma 73019 , United States.
dSchool of Civil, Mining and Environmental Engineering, University of Wollongong, Wollongong, NSW 2522, Australia
*Corresponding author: Tel.: +61 7 3365 4726; fax: +61 7 3365 4726; E-mail address: [email protected] (Z. Yuan).
1 Table of content
Abstract ...... 2 1. Introduction ...... 4 2. Inhibitory and biocidal effects of FNA...... 6
2.1 The chemical property of FNA ...... 6 2.2 Inhibitory and biocidal effects of FNA on microbes ...... 6 2.3 Proposed inhibitory and bactericidal mechanisms of FNA on microbes ...... 11
3. The use of FNA in sewer system ...... 14 4. The use of FNA in wastewater treatment system ...... 19
4.1 Achieving energy and carbon efficient nitrogen removal in mainstream wastewater treatment ...... 19 4.2 Enhancing sludge reduction and energy recovery in sludge management ...... 25 4.3 Production of FNA from wastewater ...... 29 4.4 The FNA-based applications are feasible and economically viable in wastewater treatment systems ...... 30
5. The use of FNA in membrane and algae systems ...... 34 6. Knowledge gaps and opportunities ...... 36
6.1 Fundamental knowledge on the chemistry of FNA ...... 36 6.2 The impact of FNA sludge treatment on biological phosphorus removal ...... 37 6.3 Constraints of lab-scale studies call for pilot- and full-scale trials ...... 38 6.4 Alternatives for acid addition ...... 38 6.5 Intrinsic limitations and opportunities...... 39
7. Conclusions ...... 40 References ...... 42
List of Abbreviations
Abstract
Free nitrous acid (FNA), the protonated form of nitrite, has historically been an unwanted substance in wastewater systems due to its inhibition on a wide range of microorganisms. However, in recent years, advanced understanding of FNA inhibitory and biocidal effects on microorganisms has led to the development of a series of FNA-based applications that improve wastewater management practices. FNA has been used in sewer systems to control sewer corrosion and odor; in wastewater treatment to achieve carbon and energy efficient nitrogen
2 removal; in sludge management to improve the sludge reduction and energy recovery; in membrane systems to address membrane fouling; and in wastewater algae systems to facilitate algae harvesting. This paper aims to comprehensively and critically review the current status of FNA-based applications in improving wastewater management. The underlying mechanisms of FNA inhibitory and biocidal effects are also reviewed and discussed. Knowledge gaps and current limitations of the FNA-based applications are identified; and perspectives on the development of FNA-based applications are discussed. We conclude that the FNA-based technologies have great potential for enhancing the performance of wastewater systems; however, further development and demonstration at larger scales are still required for their wider applications.
Keywords: Free nitrous acid; sewer; wastewater treatment; sludge management; membrane fouling; algae;
3 1. Introduction
Wastewater management, including wastewater collection, transportation, treatment and associated sludge management, is a most critical feature in modern societies, as it provides essential human health and environment protection. There has been extensive research and development in this area in recent times, however, there are still many challenges and improvements to be made.
Sewer systems receive and transport wastewater via underground pipelines to wastewater treatment plants (WWTPs) for pollutant removal before environmental discharge. However, sewer systems are largely concrete constuctions and are under serious threat of corrosion, which costs billions of dollars annually and has been identified as a main cause of global sewer deterioration (Pikaar et al. 2014).
In WWTPs, there are three generally accepted objectives guiding its operation: satisfactory pollutant removal, sludge reduction and resource & energy efficiency. A few obstacles in achieving the goals are: 1) conventional nitrogen removal process is carbon and energy demanding. Nitrification requires significant aeration energy and denitrification consumes a large amount of carbon source. In most conventional WWTPs, aeration takes up approximately 50-60% of all energy consumption (Gu et al. 2017); 2); waste activated sludge (WAS) is poorly degradable, resulting in high sludge handling cost, which represents 20–60% of the overall costs for wastewater treatment, and low biogas production (Foladori et al. 2010). In addition, there are many other challenges within wastewater management, such as membrane fouling impeding the development of membrane-based wastewater treatment technologies (Meng et al. 2009), lack of cost-effective algae harvesting approach that retards the application of algae technology for wastewater treatment (McGarry and Tongkasame 1971, Milledge et al. 2013).
Extensive research endeavours have been taken to address those challenges in wastewater management. In recent years, Free Nitrous Acid (FNA), the protonated form of nitrite, has shown great potential in addressing some of the above-mentioned challenges.
FNA was previously regarded mainly as an unfavourable inhibitor in wastewater management. As a compulsory intermediate of nitrification and denitrification, nitrite frequently accumulates during biological wastewater treatment. In nitrification, many environmental conditions, e.g., low dissolved oxygen (DO), short sludge retention time (SRT), high temperature, could lead to nitrite accumulation, that can be equivalent of up to 100% of the ammonium oxidized. The
4 conditions favouring the nitrite accumulation during nitrification were reviewed elsewhere (Cao et al. 2017, Ge et al. 2015, Ma et al. 2016, Peng and Zhu 2006). In denitrification, the presence of easily biodegradable organic carbon (e.g., volatile fatty acids), high C/N ratio and high pH could lead to nitrite accumulation (Cao et al. 2013, Ge et al. 2012, Glass and Silverstein 1998, Oh and Silverstein 1999). It is also reported that as high as 80% of the nitrite reduced from nitrate could accumulate stably in wastewater treatment (Du et al. 2015, Ma et al. 2017a). FNA has been a concern as is known to inhibit a wide range of microorganisms in wastewater management, including nitrogen removal bacteria (nitrifiers, denitrifier and anammox), phosphorus removal bacteria and methanogenic archaea, thus deteriorating the performance of WWTPs. Thus, much effort has been placed to understand and mitigate these inhibitions caused by FNA (Zhou et al. 2011). Readers can refer to Zhou et al. (2011) for an overview of FNA- caused inhibition of key microorganisms in wastewater treatment processes.
Recently there has been improved knowledge of the inhibitory and biocidal effects of FNA, and in contrast to FNA being a problem, there are instances where FNA can be used beneficially to improve wastewater management. The use of FNA has attracted attention for various purposes, these include reducing sewer odor and corrosion, enhancing nitrogen removal in wastewater treatment, to aid sludge reduction, for fouling control in membrane- based systems and for improving algae harvesting. Despite the obvious enhancements to wastewater management, and the rapidly increasing number of publications, there has not been an attempt to critically review all the FNA-based technologies, which is required to support further technology development and applications. Therefore, this paper aims to review the current status of the uses of FNA in wastewater management and, to provide perspectives for the ongoing development of FNA-based technologies.
5 2. Inhibitory and biocidal effects of FNA
2.1 The chemical property of FNA
FNA, the protonated form of nitrite ( ), is a weak monobasic acid (pKa=3.16, 25°C), which is mostly stable at pH ≥ 5 (da Silva et al. 2006). FNA is thought to decompose to form + nitrogen dioxide (NO2), nitrosonium cation (NO ), dinitrogen trioxide (N2O3) and nitric oxide radical (NO•) in acidic aqueous solutions (Oldreive and Rice-Evans 2001, Takahama and Hirota 2012):
- + NO2 + H ⇔ HONO (pKa=3.3) (1)
+ + + HONO + H ⇔ H2NO2 ⇔ H2O+ NO (2)
2HONO ⇔ N2O3 + H2O (3)
N2O3 ⇔ NO• + NO2 (4)
The decomposed products of FNA, and a variety of other reactive nitrogen intermediates (RNIs) that derive from their interactions with various organics in the presence of oxygen, such as peroxynitrite (ONOO-), NO-derived metal nitrosyl complexes, and S-nitrosothiols (RSNO) could have cytotoxic properties to bacterial cells (Duncan et al. 1995, Dykhuizen et al. 1996, Fang 1997, Jiang et al. 2011a, Klebanoff 1993, Phillips et al. 2004, Yoon et al. 2006) (Takahama and Hirota 2012), as will be discussed in details in section 2.3.
2.2 Inhibitory and biocidal effects of FNA on microbes
The antimicrobial properties of nitrite have been known for a long time. It has been used as a food preservative for centuries to inhibit the growth of food spoilage bacteria (Pierson and Smoot 1982). Nitrite is also well documented in medical research. For example, it is reported that salivary nitrite could serve as a useful host defence mechanism against swallowed pathogens via the formation of bactericidal compounds in the stomach (Benjamin et al. 1994). In an in vitro study, Rao et al. (2006) showed that the normal stomach acidity in combination with physiological concentrations of nitrite can kill Clostridium difficile spores and that this was due to the action of FNA rather than nitrite.
There is evidence accumulating in recent years demonstrating that FNA is the true metabolic inhibitor behind the often observed “nitrite inhibition” to a broad range of microbes at parts per
6 billion (ppb) levels. It is also seen to be a strong bactericidal agent at parts per million (ppm) levels under both aerobic and anaerobic conditions in applications of wastewater management (Jiang et al. 2011b, Vadivelu et al. 2006b, Wang et al. 2013a, Zhou et al. 2008). In this regard, FNA can be used as a bacteriostatic or bactericidal agent to control the growth and activities of microbes important to wastewater management. The bacteriostatic and bactericidal effects of FNA have been characterised on a range of microorganisms (Table 1).
7 Table 1. A summary of the bacteriostatic and bactericidal effects of FNA detected in microorganisms relevant to wastewater treatment
Types of microbes Culture/System/Organism FNA (mg N/L) Remarks (bacteriostatic or bactericidal) References Bacteriostatic effects of FNA Nitrite oxidizing Activated sludge treating municipal 0.22-2.8 Threshold for inhibition Anthonisen et al. bacteria (NOB) wastewater (1976) NOB Enriched Nitrobacter culture (73%) 0.011 Threshold for anabolic process inhibition Vadivelu et al. (2006b) 0.023 100% inhibition of biosynthesis 0.05 No inhibition on catabolism Ammonia Nitrosomonas europaea 1.72 50% inhibition on ammonia monooxygenase activity Stein and Arp oxidizing bacteria (1998) (AOB) AOB AOB enrichment 0.16 20%-25% inhibition Fux et al. (2003) AOB Enriched Nitrosomonas culture (82%) 0.10 Inhibition threshold for anabolism Vadivelu et al. (2006a) 0.40 100% inhibition on anabolism 0.50-0.63 50% inhibition on catabolism AOB SHARON process 0.21 50% inhibition Van Hulle Stijn et al. (2007) AOB Enriched AOB culture 0.0013 50% inhibition Jiménez et al. (2012) AOB and NOB Urban landfill leachate 0.004-0.68 Stable nitration occurred since NOB was completely inhibited Sun et al. (2013) Denitrifier Activated sludge receiving high- 0.039 Threshold for inhibition Abeling and strength ammonium wastewater Seyfried (1992) Denitrifier Pure culture: Pseudomonas fluorescens 0.066 Inhibition threshold for cell growth, but not for nitrate and nitrite Almeida et al. reduction, and organic carbon utilisation (1995) Denitrifier Activated sludge 0.02 Threshold for inhibition Glass et al. (1997) Denitrifier Activated sludge 0.01-0.025 40% inhibition on nitrate reduction Ma et al. (2010) 0.2 100% inhibition on nitrate reduction
8 Denitrifier Pseudomonas aeruginosa PAO1 0.1 Suppressed denitrification activity, and decreased transcript Gao et al. (2016a) levels of most denitrification genes while there was increased NO reductase activity to detoxify the NO derived from FNA. Denitrifier Pseudomonas aeruginosa PAO1 0.1 - 0.2 Temporary inhibitory effect on growth while respiratory Gao et al. (2015) inhibition was not detected. Denitrifier Pseudomonas aeruginosa PAO1 1.0 Caused respiratory inhibition. Gao et al. (2015) Denitrifier Pseudomonas aeruginosa PAO1 5.0 Caused complete cell killing and likely cell lysis Gao et al. (2015) Polyphosphate- Enriched Accumulibacter culture (86%) 0.02 100% inhibition on P-uptake, 40% inhibition on denitrification Zhou et al. (2007) accumulating organisms (PAOs)
PAOs Denitrifying P-removal culture 0.0007-0.001 50% inhibition on N2O reduction Zhou et al. (2008) PAOs Denitrifying P-removal culture; 0.01 50% inhibition on P-uptake Accumulibacter accounts for 40% 0.037 100% inhibition on P-uptake 0.02 60% inhibition on glycogen production Zhou et al. (2010) 0.02-0.07 40% inhibition on PHA degradation PAOs Enriched Accumulibacter culture (90%) 0.0005 50% inhibition on anabolism Pijuan et al. (2010) 0.006 100% inhibition on anabolism 0.002-0.01 50%-60% inhibition on catabolism PAOs FNA adapted PAOs 0.01 100% growth inhibition (Zhou et al. 2012) Glycogen- Enriched Competibacter culture (90%) 0.0015 50% inhibition on cell growth Ye et al. (2010) accumulating organisms (GAOs) PAOs and GAOs EBPR system 0.0022 A stronger inhibitory effect of FNA on the anaerobic metabolism Ye et al. (2013) of PAOs than GAOs. (Maximum) 0.01-0.025 Nitrate reduction inhibition Sludge Biological nutrient removal sludge Ma et al. (2010) 0.2 Threshold for complete nitrate reduction inhibition Sulfate-reducing Desulfovibrio vulgaris Hildenborough Activities of growth, respiration, and ATP generation were Gao et al. (2019), bacteria (SRB) 0.001-0.008 inhibited. Abundance of nitrite reductase and nitrosative stress Gao et al. (2016b) related proteins increased. Abundance of proteins related to the
9 sulfate reduction and lactate oxidation activities were initially reduced, and then recovered. Bactericidal effects of FNA Sewer biofilm Lab scale sewer biofilm system 0.2 Viable cell percentage dropped from 80% to 5-15% after 6-24 h Jiang et al. (2011b) Sewer biofilm Lab scale sewer biofilm system fed 0.26 Reduced the average sulfide production by >80% when dosed Jiang et al. (2011a) with real wastewater with FNA for 12h every 5 days 0.09 Inhibition threshold for the methane generation after 6h exposure Sewer biofilm Anaerobic wastewater system 0.2-0.3 Caused 90% biofilm inactivation with an exposure time longer Jiang and Yuan than 6h (2013) ≥0.2 Caused a 2-log (99%) biofilm inactivation with hydrogen peroxide at 30 mg/L and exposure time more than 6h Sewer biofilm Sewers systems in field trials 0.26 Gaseous hydrogen sulfide production reduced by over 95% Jiang et al. (2013) immediately after applying FNA for 6 or 24 h Sewer sediment Sediment from real gavity sewer 0.14 Sulfide production reduced by approximately 50% after a 12 h Zheng et al (2017) exposure to FNA Secondary sludge Lab scale bioreactors 2.02 Viable cells decreased to 20% and biodegradability of secondary Pijuan et al. (2012) sludge enhanced after 48 h exposure AOB and NOB Lab scale bioreactors 0.24 – 2.0 FNA is substantially more biocidal to NOB than to AOB (Jiang et al. 2018, Wang et al. 2014b) Waste activated Sludge collected from wastewater 2.13 Methane production from activated sludge was enhanced six Wang et al. sludge (WAS) treatment plant times after 24 h FNA pre-treatment compared to sludge without (2013a) the pre-treatment WAS Sequencing batch reactors 2.0 Achieved 27% reduction in sludge production without affecting Wang et al. treatment performance and sludge properties (2013b) WAS Sludge collected from wastewater 0.52-1.11 Biochemical methane potential of the FNA pre-treated WAS was Wang et al. treatment plant increased by 17-26% and hydrolysis was improved by 20-25% (2014a) when a combined FNA and heat pretreatment was used Algae Lab cultured microalgae 2.19 Lipid extraction was enhanced with FNA pre-teatment Bai et al. (2014)
10 FNA is reportedly more inhibitory on the anabolism than on the catabolism of wastewater microorganisms (Table 1). For instance, FNA caused inhibition of the anabolic processes of Nitrobacter (NOB) at approximately 0.011 mg N/L and completely stopped biomass synthesis at a concentration of approximately 0.023 mg N/L, while at concentrations of up to 0.05 mg N/L FNA did not show any inhibitory effect on the catabolic processes (Vadivelu et al. 2006b). The inhibitory effect on the anabolic processes of PAOs was also much stronger than that on the catabolic processes. About 50% inhibition on all anabolic processes of PAOs occurred at approximately 0.5 µg N/L while 50%-60% inhibition on catabolic process occurred at 2-10 µg N/L (Pijuan et al. 2010). Thus it is apparent that FNA is having more immediate effects on growth related activities in comparison to substrate utilisation based activities such as energetic transformations.
Various aspects of the biocidal effects of FNA have been studied (Table 1). It is seen that FNA at ppm levels is strongly biocidal to wastewater microorganisms, including those in anaerobic sewer biofilms (Jiang et al. 2011b) and in activated sludge (Pijuan et al. 2012, Wu et al. 2018). In a study of nitrite addition to a sewer with simultaneous acidification of sewage to pH 6, a substantially reduced sulfide- and methane-production occurred immediately, then these activities slowly recovered. These altered activites suggest the inactivation and regrowth of sulfate reducing bacteria (SRB) and methanogens (Jiang et al. 2010), and imply the biocidal effects of FNA. The biocidal effects of FNA on biofilm microorganisms has been confirmed by detecting decreased viability in biofilm cells after FNA treatment (Jiang et al. 2011b). After 6–24 h treatment at FNA levels of 0.2-0.3 mgN/L, the viable fraction of microbes decreased significantly from approximately 80% prior to treatment to 5–15%. The biocidal effect was strongly dependent on the FNA concentration rather than the nitrite concentration or the pH level individually. Details of the application of FNA in the sewer is presented in Section 3.
Significant biocidal effects of FNA are also observed for activated sludge microorganisms. After 24 - 48 h contact with FNA at 0.5-2.0 mg-N/L, the viable fraction of microbes in activated sludge declined to 20-50% from an initial fraction of 80% in the absence of added FNA (Pijuan et al. 2012). In that study the denitrification activities of the activated sludge were completely lost after the 48 h contact. Following the contact period FNA addition was removed and the denitrification rates did not recover over 3 days, indicating permanent damage was caused by the FNA contact (Pijuan et al. 2012). It should be noted that sensitivities of different microorganisms to the biocidal effect of FNA may vary significantly. AOB are found more
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resistant to the biocidal effects of FNA than NOB (Table 1). This characteristic can be used to selectively remove NOB from the activated sludge in nitrogen removal systems (as discussed in Section 4). The strong biocidal effects of FNA could also be used to improve the biodegradability of secondary sludge (Section 4). In addition, the biocidal and acidic nature of FNA could be applied for biofouling and scaling control of membrane, and for facilitating the harvesting of algae (Section 5).
2.3 Proposed inhibitory and bactericidal mechanisms of FNA on microbes
It is believed that the antimicrobial mechanism of FNA is multi-targeted (Gao et al. 2016a, Gao et al. 2015, Gao et al. 2019, Gao et al. 2016b). FNA generally causes nitrosative stress to bacterial cells in co-action with various effective molecules (Fang 1997, Kim et al. 1999). FNA itself and the RNIs derived directly from FNA, or from its reactions with certain cellular components as described above, may lead to various damage to enzymes, cellular membranes, cell walls and nucleic acids (Gao et al. 2014, O'Leary and Solberg 1976).
FNA alters the intracellular pH
FNA is able to diffuse across cell membranes, where it may accumulate inside the cells in its anionic form due to the near neutral intracellular pH. This release of protons within the cells would decrease the intracellular pH and interfere with the trans-membrane proton motive force (pmf) required for ATP synthesis (Anthonisen et al. 1976, Hinze and Holzer 1985, Prakasam and Loehr 1972). To counteract the acidification effect of FNA, cells could pump out protons through an energy-requiring plasma H+-ATPase, which would uncouple energy generation from growth (Brul and Coote 1999) and potentially be bactericidal. The intracellular pH of Salmonella enterica Serovar Typhimurium and Pseudomonas aeruginosa PAO1 are observed to decrease upon the addition of FNA (Gao et al. 2016a, Mühlig et al. 2014). However, other studies may not support this view. For example, there is no significant uncoupling effect detected in Paracoccus denitrificans upon FNA exposure (Alefounder et al. 1983). Although the growth was inhibited severely, the energy generation capacity of Nitobacter was not affected (Vadivelu et al. 2006b). These findings implicate that this mechanism of bactericidal effect of FNA requires further investigation, and it may not be a general mechanism across the range of microorganisms.
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FNA acts as a protonophore to uncouple the proton motive force and inhibit intracellular ATP generation
As mentioned above FNA may passively pass through bacterial cytoplasmic membranes, which would increase the proton permeability of membranes and lead to dissipation of the transmembrane electrochemical pmf. This collapse of pmf would inhibit the ATP synthesis, affecting various exchange reactions catalysed by ATPase, and be energetically damaging to the cell. Decreased bacterial growth and even cell death could be caused by the expenditure of energy required to maintain a pmf (Almeida et al. 1995, Anthonisen et al. 1976, Sijbesma et al. 1996). The intracellular ATP generation of Pseudomonas, Desulfovibrio and Accumulibacter were consistently seen to decrease with increasing levels of FNA (Gao et al. 2016a, Gao et al. 2016b, Zhou et al. 2010, Zhou et al. 2007). As well, the addition of FNA is also found to inhibit ATP synthesis in Escherichia coli O157:H7 cells and in the yeast Saccharomyces cerevisiae (Hinze and Holzer 1985, O'Leary and Solberg 1976).
FNA inactivates microbial metabolic activities
FNA is reported to interact with enzymes that contain thiols, heme groups, iron sulfur clusters, phenolic or aromatic amino acid residues, tyrosyl radicals, and amines. This would lead to altered activities of these enzymes (Fang 1997). For instance, sulfhydryl (SH)-containing enzymes are key participants in the tricarboxylic acid (TCA) cycle. The activities of those enzymes could be suppressed by FNA, which would have negative impact on the energy generation process of cells (O'Leary and Solberg 1976, Park 1993). FNA could also inhibit the genes/proteins expression involved in anaerobic respiration, carbon utilization, and energy production of P. aeruginosa and D. vulgaris (Gao et al. 2016a, Gao et al. 2015, Gao et al. 2019, Gao et al. 2016b). In P. aeruginosa, FNA was seen to down-regulate the expression of genes involved in denitrification, and carbon utilization. This included genes coding for nitrate
reductase, nitrite reductase, nitric oxide reductase, and N2O reductase and the altered expression would affect the energy conservation abilities. In addition to the down regulation of
the genes, the activity of the N2O reductase could be directly altered by FNA. The reductase contains two metal centres, a binuclear copper centre, CuA, for electron transfer, and another tetranuclear copper-sulfide centre, CuZ, at the active site. FNA could bind to the active sites of these copper-containing enzymes and cause competitive inhibition of N2O reduction (Zhou et al. 2008). In D. vulgaris, FNA was seen to inhibit gene and protein expression involved in
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sulfate reduction, energy generation, and lactate oxidation (e.g. pyruvate ferredoxin reductase). Additionally, FNA is seen to disrupt the expression of genes participating in DNA replication, transcription, and translation for both P. aeruginosa and D. vulgaris (Gao et al. 2016a, Gao et al. 2015, Gao et al. 2019, Gao et al. 2016b).
Glyceraldehyde-3-phosphate dehydrogenase, an enzyme involved in both glycolysis and gluconeogenesis is found to be inactivated by FNA (Hinze and Holzer 1986, O'Leary and Solberg 1976), This finding coincides with that of Zhou et al. (2010) where both glycogen generation (gluconeogenesis) and degradation (glycolysis) were severely inhibited under -3 anoxic conditions with FNA at 10.0×10 mg HNO2-N/L. In contrast, in P. aeruginosa and D. vulgaris, the down regulation of glyceraldehyde-3-phosphate dehydrogenase was not detected (Gao et al. 2016a, Gao et al. 2015, Gao et al. 2019, Gao et al. 2016b).
Other damage caused by FNA
As an extremely reactive molecule, FNA is capable of interacting with a broad range of substrates and inducing mutagenic effects. For example, FNA can cause oxidative deamination of the amine (NH2) group of adenine or cytosine (Malling 2004). It can also convert adenine to hypoxanthine, cytosine to uracil and guanine to xanthine. By altering DNA bases directly this could generate an altered or “miscoding” form, which could have detrimental or lethal affects (Malling 2004).
RNIs cause inhibitory and biocidal effects to bacterial cells
Under aerobic conditions, the reactive species, ONOO- can be formed from the reaction of NO with superoxide or through nitrosyl ion reaction with oxygen (Maraj et al. 1995). As a reactive
nitrogen species, ONOO- can be transformed to NO2• and OH• radicals, that would contribute to oxidative stress. ONOO- can cause microbial cell death through the oxidation of protein thiols, iron-sulfur centres, DNA, and membrane phospholipids, as well as through the - inhibition of the electron transport chain (Zaki et al. 2005). ONOO and NO2 are seen to mediate lipid peroxidation, which results in cell membrane damage (Halliwell et al. 1992, Rubbo et al. 1994) and can also non-specifically oxidize proteins at a variety of sites (Ischiropoulos and Al- Mehdi 1995). Peroxynitrite is proven to be a strong nitrosating intermediate and can cause altered function of certain proteins (Yoon et al. 2006) as well as DNA mutation through base - conversions (e.g. G:CA:T transitions). Changes may also involve ONOO and NO2, since
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both of them can cause oxidative damage to DNA, resulting in termination sites, strand breaks, and a variety of other DNA alterations (Juedes and Wogan 1996). It is seen that bacteria deficient in DNA recombinational repair are hypersusceptible to inhibition or killing by S- nitrosotiols and peroxynitrite, this further indicating the detrimental interactions between RNIs, DNA and DNA repair systems.
NO reacts directly with haem and metal centres of proteins, forming nitrosyl complexes (Fang 2004). Molecular targets of NO encompass haem and nonhaem iron cofactors, iron-sulfur clusters, and other redox metal sites, all forming metal-nitrosyl complexes (Fang 2004). Therefore, NO can cause the inactivation of metal-containing enzymes such as glyceraldehyde- 3-phosphate dehydrogenase, guanylyl cyclase, catalase, and cytochrome systems (Murad 1994). (Hinze and Holzer 1986). For instance, FNA is seen to promote iron depletion inside bacterial cells through the formation of the iron-nitrosyl complexes or cause other iron related effects by direct release of iron from metalloenzymes (Drapier et al. 1991). Moreover, the observed production of iron-nitrosyl complexes and inactivation of enzymes containing Fe-S clusters (e.g. aconitase, NADH dehydrogenase, succinate dehydrogenase), is found to be triggered by RNIs other than NO itself. This suggests that NO-related inactivation of iron-sulfur clusters might be an indirect result of reduced iron availability (Drapier et al. 1991, Hentze and Kühn 1996). In addition, NO can interact with proteins with reduced thiol groups in the presence of electron acceptors, forming S-nitrosothiols and thereby disabling the enzymes (Gow et al. 1997). S-nitrosothiols are organic compounds or functional groups containing a nitroso group attached to the sulfur atom of a thiol and may cause biocidal effects. For example, the reaction of NO with cysteine sulfhydryl can result in the S-nitrosylation and disulphide-bond formation, which will inactivate certain enzymes and is seen to cause detrimental effect to bacterial cells (Nathan and Shiloh 2000). N-nitroso compounds are easily formed by interaction of a secondary amino compound with a nitrosating agent such as NO+ (Lijinsky 1999). It is seen that nitrosamine can be formed by the reaction of NO+ with an amine, thereby modifying proteins and potentially altering function (Kulshrestha et al. 2010). On top of these, the interactions between NO and tyrosyl radicals accounts for the inhibition of ribonucleotide reductase by RNIs, which would limit the availability of precursors for the synthesis and repair of DNA (Lepoivre et al. 1991). During investigation of the disruption of P. aeruginosa biofilms it was observed that bacteriophage genes were activated with NO exposure. It was speculated that the NO induced phage could mediate the cell lysis and death (Platt et al. 2008).
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There are various inhibitory and biocidal effects proposed for these RNIs. The extent of the on the microorganisms would depend on the physiological conditions within the cell and the propensity for the RNI formation from FNA.
3. The use of FNA in sewer systems
Hydrogen sulfide (H2S), produced by sulfate-reducing bacteria (SRB) in sewers under anaerobic conditions, is an important source of odor and health hazards (Jiang et al. 2017, Jiang et al. 2015a, WERF 2007). The transfer of hydrogen sulfide from sewage to the sewer atmosphere will occur and in the aerobic conditions can be oxidized by sulfide-oxidizing microorganisms (SOM) into sulfuric acid (H2SO4). The acid then reacts with the concrete to cause corrosion, creating major challenges to infrastructure management through reduced service life and high costs for sewer remediation (Jiang et al. 2016, Li et al. 2017, US EPA 1992).
Different types of biocides like caustic and molybdate were employed to prevent sulfide generation by inactivating SRB in sewer biofilms (Predicala et al. 2008, Zhang et al. 2009). Caustic compounds like sodium hydroxide are among the commonly used biocidal agent for sulfide control in sewers (Ganigue et al. 2011). The effectiveness of caustic shock, i.e.
temporarily increasing wastewater pH to 10-11, is limited, with H2S levels being reduced by 40-50% (Gutierrez et al. 2009, O'Gorman et al. 2011, Tomar and Abdullah 1994). Another SRB inhibitor, molybdate, is not reported in sewer applications although it is demonstrated for sulfide control in oil fields and in anaerobic digesters (Kjellerup et al. 2005, Tanaka and Lee 1997).
The strong biocidal effect of FNA on sewer biofilms implies that the simultaneous dosage of nitrite and acid can achieve rapid inactivation of microbes in sewer biofilms, making it possible to achieve sulfide and methane control through short, intermittent dosages. In comparison to the conventional biocides, FNA is being applied only recently for sulfide control in sewers. A long-term laboratory study was carried out to optimize key operational parameters for FNA dosing, i.e. FNA concentration, exposure time, and dosing interval, for an intermittent dosing strategy (Jiang et al. 2011a). It was found that FNA concentration at 0.26 mg-N/L or above, with an exposure time of 12 h or longer, could suppress sulfide and methane production by sewer biofilms. Furthermore, the inactivation kinetic parameters of sewer biofilm by FNA was evaluated with a 2-fraction Bayesian model (Jiang and Yuan 2014). The study determined the
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most cost-effective dosage (FNA concentration times exposure time) as 1.4 mgN∙h/L for the 80% control efficiency of hydrogen sulfide.
After lab-scale studies, the long-term effectiveness of FNA dosing for control of sulfide
production, with and without simultaneous H2O2 dosing, was evaluated in a full-scale rising main sewer located in the Gold Coast, Australia (Jiang et al. 2013). In addition to confirming the effectiveness of intermittent dosing of FNA or FNA+ H2O2 in controlling sulfide production in rising main sewers, no biofilm adaptation and resistance to FNA was observed through the 6-month trial. Instead, successive dosing achieved better control efficiency which was thought to be due to repetitive weakening of the biofilms. Furthermore, the general cost of chemical dosing for sulfide control in sewers is 0.04 to 0.18 $/m3. This is through use of the typical chemicals that include ferrous or ferric chloride, nitrate, oxygen, magnesium hydroxide and sodium hydroxide (Ganigue et al. 2011). In comparison to current chemical dosing strategies, 3 the cost of intermittent FNA or FNA+H2O2 dosing in sewers is much less at 0.01 – 0.03 $/m (Table 2). Other full-scale evaluations conducted in the USA and Australia also confirm the intermittent FNA dosing as a cost-effective and environmentally-friendly strategy for sulfide and methane control in sewers (Table 2).
Table 2. A summary of results and costs of intermittent dosage of FNA alone or with H2O2 dosing in sewers
Chemicals Dosage Dosing Reduction ratio Cost References (mgN∙h /L) Frequency ($/m3 sewage) FNA 3.12 12 hours/4.5 Sulfide: 80% 0.03 (Jiang et al. 2011a) days Methane: completely compressed FNA 1.54 6-8 Sulfide:88%- 0.02-0.03 (Leemon et al. 2015, hours/week 99% Nguyen and Marano 2016) FNA+H2O2 2.08+480 8 hours/10 Sulfide: about 0.01 (Jiang et al. 2013) mg/L H2O2 days 90% FNA 2.08 8 hours/10 Sulfide: about 0.01 (Jiang et al. 2013) days 80%
Following the formation and partition of H2S into the sewer air, microbial sulfide oxidation processes play critical roles in the corrosion of concrete sewers (Cayford et al. 2012, Jiang et al. 2014, Jiang et al. 2015b). To inhibit the growth and the activity of SOM on the concrete surface, various surface treatment methods are widely applied. This includes the blending of biocides or inhibitors into the concrete (Negishi et al. 2005); elevating surface pH of concrete
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by application of magnesium hydroxide (Davis et al. 1998) and antimicrobial coatings such as silver-loaded zeolite and polymers (Haile and Nakhla 2010, Lebrero et al. 2011). Among all the approaches, the environmental footprint due to chemicals application, the cost and the difficulty of surface coating cause further concerns regarding expenditure and safety.
Nitrite has been widely used in concrete to prevent rebar corrosion caused by chloride attack and accelerating the hydration process (Justnes and Nygaard 1995, Okeniyi et al. 2013). Recently, FNA, formed by spraying nitrite on acidic corrosion surface of concrete, is used as a biocide to inhibit the metabolism of acidophilic SOM. After spraying nitrite on corroding
concrete surface, the H2S uptake rates of the corroding concrete were reduced by 84%-92% and no obvious recovery of the uptake rates was detected within 12 months of the nitrite spray (Sun et al. 2015). Compared with conventional approaches, FNA spraying is more cost- effective and environmentally-friendly to mitigate the sewer corrosion as nitrite is cheap and the residual nitrite is biodegradable in wastewater. Aside from mitigation/rehabilitation of the corrosion in existing sewers, in a recent study, FNA achieved a long-lasting effect on inhibiting the growth and activity of acidophilic SOM on the concrete surface by blending calcium nitrite into concrete (Li et al. 2019). With the nitrite admixed concrete, around 30% reduction in sulfide uptake rate was observed in comparison to the concrete without the nitrite admixture, during 18 months of exposure in a real sewer system (Li et al 2019). This was accompanied by the same level of reduction in concrete loss (due to corrosion) measured when the admixture was present.
Compared with the mitigation approach of spraying nitrite on the acidic concrete surfaces, the admixed nitrite inside concrete achieves a long-lasting effect with minimum operational requirements. However, the application of FNA for corrosion control on concrete is an emerging area with limited understanding. For FNA spraying on corroding concrete surfaces, further investigations are needed to optimize the concentration, delineate the influence of flooding events and to evaluate the effectiveness in real sewers. For nitrite-admixed concrete, the impact of admixture concentrations, the potential biofilm adaption and long-term effect throughout the service life of sewers are still unclear. Further tests are indispensable to fully understand the FNA impact and mitigation on concrete corrosion processes.
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4. The use of FNA in wastewater treatment systems
As afore-mentioned, three main objectives generally guiding the operations of wastewater treatment are: pollutant removal, sludge reduction and energy efficiency/recovery. Over the recent years, FNA has been fruitfully applied in wastewater treatment systems to achieve those objectives directly or indirectly, mainly by attaining mainstream energy and carbon efficient nitrogen removal and improving sludge management. It is also important to note that the proposed applications of FNA in wastewater treatment systems are practically feasible and economically attractive (Wang et al. 2016c).
4.1 Achieving energy and carbon efficient nitrogen removal in mainstream wastewater treatment
Removing pollutants, including carbon, nitrogen and phosphorus, and their recovery as renewable resources, is one of the primary objectives of wastewater treatment systems. + - - Conventionally, nitrogen is removed biologically by nitrification (NH4 →NO2 →NO3 ) and - - denitrification (NO3 → NO2 →N2, with organic carbon as the electron donor) in WWTPs. + - Biological nitrogen removal via the nitrite shunt (NH4 →NO2 →N2) or by combining Partial + - + - Nitritation (NH4 →NO2 ) and Anammox (NH4 +NO2 →N2) (PN/A, also called deammonification) processes are two novel energy and carbon efficient nitrogen removal technologies (Cao et al. 2017, Kartal et al. 2010, Peng and Zhu 2006). The novel processes bring benefits through COD usage, lowered oxygen requirement and less excess sludge production. Compared with conventional process, the nitrite shunt process reduces the COD requirement for denitrification, oxygen requirement for nitrification, and sludge production by 40%, 25% and 33-55% respectively (Antileo et al. 2013, Turk and Mavinic 1986). The PN/A process could further reduce the COD and oxygen demand, and excess sludge production, by nearly 100%, 60% and 80%, respectively (Jetten et al. 1997, Wett 2007). The key to the nitrite - shunt or partial nitritation is the selective suppression of nitrite oxidizing bacteria (NOB; NO2 - + - →NO3 ) while maintaining the ammonium oxidizing bacteria (AOB; NH4 →NO2 ), which can be achieved by using FNA. FNA has been known to be more inhibitory to the growth of NOB than to AOB, which plays a major role in the washout of NOB in systems treating ammonium- rich wastewater (e.g., sidestream, landfill leachate) (Zheng et al. 2018).
+ Unlike ammonium-rich wastewater, the low ammonium concentration (30–70 mg NH4 -N/L) in mainstream wastewater does not allow sufficient levels of FNA to washout NOB in
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mainstream wastewater treatment. To this regards, Wang et al. (2014b) firstly proposed the FNA sludge treatment approach to selectively suppress mainstream NOB by sidestream treatment of activated sludge using FNA produced from anaerobic digestion liquor (the FNA production will be discussed in section 4.3). In this approach, a certain portion (20-30%) of mainstream activated sludge was transferred into a side-stream treatment unit on a daily basis, for exposure to high level of FNA to inactivate the NOB contained in the activated sludge. By adopting a treatment ratio of 22% (i.e. 22% of the sludge treated daily), an FNA concentration - of 1.35 mgN/L (NO2 = 550 mgN/L, pH=6.0) and treatment time of 24 hours, Wang et al. (2014b) demonstrated that mainstream nitrite shunt can be quickly established within two - - - weeks, achieving a nitrite accumulation ratio (NAR=NO2 /(NO2 +NO3 )) of 81.5 ± 0.1%. NAR is a direct indicator of the level of selective NOB suppression, which can reflect the attainment of energy and carbon efficient nitrogen removal processes. The FNA sludge treatment approach was also successfully adopted in a mainstream PN/A system obtaining stable partial nitritation with NAR of 60 - 80% (Wang et al. 2016a). The FNA sludge treatment approach has also been tested in other long-term experiments under varying conditions and different reactor configurations, achieving stable mainstream NOB suppressions with NAR as high as 90%, which further confirmed the effectiveness of this approach (Table 3).
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Table 3 Long-term demonstrations of FNA sludge treatment approach with different treatment conditions in mainstream wastewater treatment
Wastewater Wastewater Wastewater SRT HRT Reactor FNA Nitrite pH Treatment Treatment DO AOB NOB NAR Reference + type NH4 COD (d) (h) configuration (mgN/L) (mgN/L) ratio Time (h) (mgO2/L) activities activities (%) (mgN/L) (mgCOD/L) (%) (%) Synthetic 50 300 15 24 SBR 1.35 550 6 0.22 24 1.5-2.0 77.00 10 82 (Wang et al. 2014b) Synthetic 57 0 8 13.2 SBR 1.82 750 6 0.25 24 2.5-3 88.5 50.4 60 (Wang et al. 2016a) Synthetic 57 0 8 13.2 SBR 1.82 750 6 0.25 24 0.3-0.8 101.4 15.2 80 (Wang et al. 2016a) Synthetic 57 0 8 13.2 SBR 1.82 750 6 0.123 24 0.3-0.8 100.1 15.7 80 (Wang et al. 2016a) Synthetic 50 300 15 24 SBR 1.35 550 6 0.22 24 1.5-2 95.2 27.9 56 (Duan et al. 2018) Synthetic 50 300 15 24 SBR 1.35 550 6 0.31 24 1.5-2 74.6 23 72 (Duan et al. 2018) Synthetic 50 300 15 24 SBR 1.35 550 6 0.38 24 1.5-2 40 9.8 75 (Duan et al. 2018) Synthetic 50 300 15 24 SBR 4.23 550 5.5 0.31 24 1.5-2 82.5 6.6 90 (Duan et al. 2018) Synthetic 50 150 30 3.75 Two-sludge 1.1 450 6 0.063 24 2-2.5 95.1 59.1 29 (Wang et al. 2017a) system Synthetic 50 150 30 3.75 Two-sludge 1.9 780 6 0.063 24 2-2.5 91.8 45.1 68 (Wang et al. 2017a) system Real 60.2±7.5 257.1±7.4 15 4 Plug flow 0.25 - 1 100-400 6 0.10 6 ~20 (Ma et al. 2017b) (An/O) Real 60.2±7.5 257.1±7.4 15 4 Plug flow 1 400 6 0.30 6 68 99 ~17 (Ma et al. 2017b) (An/O) Real 50.3±4.9 365.2±40.5 15 10 Plug flow 1.2 550 6 0.30 18 0.5 55.3 2.1 77 (Jiang et al. 2019) (A2O) Real 50.3±4.9 365.2±40.5 15 10 Plug flow 1.2 550 6 0.30 18 1.5 84 10.8 78 (Jiang et al. 2019) (A2O)
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The key factor underpinning the FNA sludge treatment approach is the fact that AOB are more resilient to the treatment and therefore are less or even not inactivated. Batch tests with FNA treatment consistently suggest selective inactivation of NOB activities against AOB (Jiang et al. 2019, Jiang et al. 2018, Ma et al. 2017b, Wang et al. 2016a, Wang et al. 2014b). The mechanism of the sludge treatment approach can be interpreted as manipulating SRTs of AOB and NOB in the bioreactor. Although the overall SRT in the bioreactor is maintained by sludge wastage, the selective inactivations by FNA treatment forms an additional pressure on the growth of microorganisms, leading to distinct reductions of real SRTs for AOB and NOB. As a consequence of not significantly affected SRTAOB and substantially reduced SRTNOB, NOB were washed out while ammonium oxidation capability was retained (Duan et al. 2018).
Figure 1 Effects of FNA treatment conditions on AOB and NOB activities (Jiang et al. 2019, Jiang et al. 2018, Ma et al. 2017b, Wang et al. 2016a, Wang et al. 2014b).
The treatment conditions of the approach, namely the treatment ratio, FNA concentration and treatment time, can significantly affect the effectiveness of FNA treatment on NOB suppression. A large amount of batch tests were carried out to investigate the FNA treatment conditions, as visualised in Figure 1. Relative activities of AOB and NOB indicate the extent of suppressions by varying FNA treatment conditions. While the sensitivities to FNA treatment vary substantially among studies, NOB were always more sensitive than AOB to any FNA treatment conditions. In general, the effectiveness of NOB suppressions improves with the increment of FNA concentrations to 1.35 mgN/L. Further increase of FNA concentrations did not significantly improve the NOB suppression. At relatively low FNA concentrations (<1 mgN/L), the effectiveness of NOB suppression is strongly correlated to FNA treatment time (0-30 h).
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At higher FNA concentrations (>1 mgN/L), increasing treatment time beyond 12 hours does not significantly enhance the efficiency. Similar correlations between AOB suppressions with FNA treatment conditions were observed, nevertheless in a less sensitive way than that with NOB. In addition, the activated sludge concentrations might influence NOB suppression. It was recently revealed that FNA inactivation of NOB became less effective with higher sludge concentrations, at relatively low FNA concentrations (<0.77 mgN/L) (Wang et al. 2019). It should also be noted that the starvation of activated sludge could facilitate the NOB resistance to FNA treatment. Both pre-starvation and post-starvation were shown, in batch tests, to deteriorate the effectiveness of selective NOB suppression by FNA treatment (Nan et al. 2019).
Although batch tests provided useful information on FNA treatment conditions, cautions should be taken when applying batch tests result in long-term reactor operation designs. This is because batch tests only reflect the short-term consequences of FNA treatment. Long-term effects, such as microorganism adaptations are, however, not demonstrated. For example, based on batch test results, Ma et al. (2017b) selected one FNA treatment condition (FNA = 0.25 mgN/L, treatment ratio = 0.1 and treatment time = 6h) for application in the long-term study, which resulted in only fleeting success, establishing mainstream nitrite shunt with NAR as high as 75%. While the selected FNA treatment condition produced promising results in batch test (107.8±4.8% AOB activities and 35.2±1.7% NOB activities), it induced quick adaptation of NOB to FNA treatment over the long-term reactor operation, leading to the eventual failure of the NOB suppression. Therefore, long-term evaluation of FNA treatment conditions is necessary. Limited number of long-term studies (Table 2) have suggested that treatment ratio should be relatively high (≥25%) during start-up to allow maximal inactivation of NOB, and can be relaxed at steady operation (as low as 6.3%) (Duan et al. 2018, Wang et al. 2017a, Wang et al. 2016a). The FNA concentration should be relatively high (>1.35 mgN/L) to prevent quick NOB adaptation. More long-term evaluations of FNA treatment conditions should be carried out in the future to guide the selection of FNA treatment conditions for real application.
It has been increasingly realised that NOB adaptation poses a critical challenge to exisiting NOB suppression approaches, including the FNA sludge treatment approach. NOB resistances to FNA treatment have been observed in long-term studies, due to NOB microbial community shifts, enriching certain strains of either Candidatus Nitrotoga or Nitrospira, which are likely more resistant to FNA treatment (Duan et al. 2019b, Duan et al. 2019c, Ma et al. 2017b). NOB
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adaptation is likely a universal problem for NOB suppression approaches. NOB resistances to low DO, free ammonia (FA) inhibition, have also been reported (Cao et al. 2018, Liu and Wang 2013, Turk and Mavinic 1989, Wong-Chong and Loehr 1978, Zhang et al. 2018). It is likely that not all NOB strains are equally susceptible to the same suppression, and consequently, the less susceptible strains could outcompete others and emerge as the dominating NOB population, leading to suppression failure. Therefore, it would be extremely challenging to suppress NOB growth using a single factor.
The combination of the FNA approach with others could address the universally challenging NOB adaptation issue, as demonstrated in a recent study (Duan et al. 2019b). In this work, the sludge treatment was alternated between using FNA and FA, both of which can be produced from AD liquor and are selectively inhibitory to NOB (Wang et al. 2017b). During over 650 days reactor operation, NOB adaptation to both treatment approaches was observed but the adaptation was successfully overcome by deploying the alternate treatment strategy. Stable mainstream nitrite shunt with NAR over 95% was achieved and maintained for over a 10- month demonstration period. It was revealed that Nitrospira and Nitrobacter, the key NOB populations in most WWTPs, could adapt to the FNA and FA treatment, respectively, but do not adapt to the alternation. This approach effectively addressed the NOB adaptation challenge and thus ensuring a more stable NOB suppression. The stable suppression of NOB is one of the major barriers for the scale-up of mainstream nitrite shunt or PN/A (Lotti et al. 2015). The FNA based approach showed very promising results in overcoming the NOB adaptation issue, making a significant step forward in achieving reliable mainstream nitrite shunt or PN/A process.
Apart from achieving the nitrite shunt or PN/A, the FNA sludge treatment brings additional benefits to mainstream nitrogen removal process, such as enhancing denitrification, reducing sludge production and abating N2O emissions. FNA could breakdown cell membrane and extracellular polymeric substances (EPS) in the activated sludge, releasing internal carbon source contained in activated sludge. It was demonstrated that, the internal carbon source production increased by 50% in a denitrification reactor fed with FNA treated activated sludge (T=24 h, FNA= 2.04 mgN/L). As a result, the FNA treatment of excess sludge along with its subsequent recirculation enhanced denitrification efficiency by up to 76%, and reduced sludge production by 28% and 87.5% in a nitrification/denitrification reactor and in a denitrification reactor, respectively (Ma et al. 2015, Wang et al. 2013b). In addition, the emission of N2O, an
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important greenhouse gas from nitrogen removal process, was found to reduce after receiving FNA sludge treatment (Wang et al. 2016b). The decrease could be attributed to two possible mechanisms. Firstly, the increased internal carbon facilitated denitrification (Ma et al. 2015),
which is a sink for N2O emission (Duan et al. 2017). Secondly, FNA treatment could lead to a
shift of the stimulation threshold of nitrite on N2O emission by inhibiting the enzymes relevant
to nitrifier denitrification (Wang et al. 2016b). However, other studies reported elevated N2O emission after FNA treatment during nitrogen removal process (Duan et al. 2018, Duan et al.
2019b, Wang et al. 2014b). The effects of FNA treatment on N2O emission remains to be further investigated.
4.2 Enhancing sludge reduction and energy recovery in sludge management
The activated sludge process in WWTPs generates a large amount of excess sludge, i.e., waste activated sludge (WAS) which contains bacteria, EPS, recalcitrant organics, and inorganics (Tchobanoglous et al. 2003). Due to the health concerns and to reduce sludge disposal cost, sludge stabilization is practiced in most WWTPs. Among all sludge stabilization technologies, anaerobic digestion (AD) and aerobic digestion (AeD) are the two most used ones. The biodegradability of WAS is relatively low, which curbs the performance of AD or AeD (Carrère et al. 2010); therefore lots of pre-treatment technologies have been developed to enhance the biodegradability of WAS (Wang et al. 2017c).
FNA-based pre-treatment has been successfully shown to enhance sludge degradation and methane production in AD. It was firstly demonstrated in a Biochemical Methane Potential (BMP) test that pre-treatment of WAS with FNA at concentrations from 1.78- 2.13 mgN/L for 24 hours could substantially facilitate sludge solubilisation, and increase the subsequent methane production in AD by around 30% (Wang et al. 2013a). Model analysis suggested the improvement in methane production in AD was due to the sharp increase of hydrolysis rate −1 (from 0.16 to 0.25 d , 50% enhancement) and methane potential (from 201 to 255 L CH4/kg VS added, 27% enhancement). Intensive studies were later carried out to evaluate the FNA pre-treatment with WAS and primary sludge (PS) (Zahedi et al. 2016, 2017a, b, Zahedi et al. 2018). Results of some typical studies are presented in Table 4. Improved solubilization and digestibility of WAS were consistently observed after FNA pre-treatments, contributing to the increased methane production in BMP tests. However, FNA pre-treatments with PS resulted in limited release of readily biodegradable matters and compromised methane productions.
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Accordingly, the FNA pre-treatment with the mixture of PS + WAS produced mixed results (Table 4). The contrasting observations with WAS and PS can be explained by the mechanism of FNA pre-treatment with sludge. FNA pre-treatment is effective in disintegrating EPS and cell membrane (Wu et al. 2018, Zhang et al. 2015). The distinct portion of biomass in WAS (high) and in primary sludge (low) leads to the different effectiveness of FNA pre-treatment.
Similar to the FNA treatment for sewer and NOB suppression, the FNA concentration and treatment time are critical operational parameters in the FNA pre-treatment process for WAS. FNA concentrations in the range of 0.36 -2.13 mgN/L and treatment time in the range of 0 - 24 hours both were positively correlated to enhancements of WAS solubilization, and increments of methane production (Ma et al. 2015, Wang et al. 2013a). Although batch studies have indicated a clear relationship between the treatment conditions and the WAS solubilization, methane production in AD, long-term studies were still required to verify the real impact of treatment conditions on the AD performance. In all batch studies, pH was adjusted after the pre-treatment to remove residual FNA inhibition, before entering the serum bottles for anaerobic digestion. However, the residual FNA inhibition under different pre-treatment conditions could affect the AD operation and therefore needs to be evaluated. Also, due to the inherent limitations of batch tests, it is difficult if not impossible to infer conclusions as to the corresponding effects in full scale continuous flow reactors (Hilton Barry and Oleszkiewicz Jan 1988). Therefore, effects of FNA pre-treatment conditions needed to be systematically evaluated in continuous AD reactor operations.
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Table 4. Typical studies on the effects of FNA pre-treatments in different conditions with WAS, PS and WAS+PS. *increments are shown.
FNA Nitrite pH Treatment Sludge sCOD SKN k (d-1) Specific methane Proteins Fraction of Reference concentrations (mgN/L) time type (mgCOD/gVS) (mgN/gVS) production (mg/gVS) biocidal (mgN/L) (h) (mL/gVS) cells (%) 0 0 N/A N/A WAS 26.1 2.3 0.16 ± 0.01 201 ± 5 3.4 ± 2.6 (Wang et al. 2013a) 0.36 50 5.5 24 83.7 10.1 0.19 ± 0.02 218 ± 6 23.2 ± 2.7 0.71 100 5.5 24 109.1 11.4 0.20 ± 0.02 230 ± 6 23.3 ± 2.5 1.07 150 5.5 24 142.4 11.9 0.21 ± 0.02 228 ± 7 23.7 ± 3.3 1.42 200 5.5 24 144.9 12.4 0.25 ± 0.02 236 ± 8 25.7 ± 4.1 1.78 250 5.5 24 162.5 14.3 0.25 ± 0.02 246 ± 8 30.0 ± 3.1 2.13 300 5.5 24 165.0 14.6 0.18 ± 0.02 255 ± 7 33.6 ± 2.7 0 0 5.6- 24 PS 78.7 ± 1.0* 0.25 ± 0.01 415 ± 12 (Zhang et 5.8 al. 2016) 0.77 100 5.5 24 90.0 ± 2.1* 0.14 ± 0.01 400 ± 10 1.54 200 5.5 24 108.5 ± 0.9* 0.12 ± 0.01 403 ± 12 2.31 300 5.5 24 104.0 ± 0.8* 0.1 ± 0.01 412 ± 20 3.08 400 5.5 24 100.2 ± 2.1* 0.14 ± 0.02 359 ± 13 3.85 500 5.5 24 103.9 ± 0.9* 0.12 ± 0.02 345 ± 20 0 0 5.5 N/A PS + WAS 308.5 7.9 388 ± 26 16.5 ± 2.0 (Zahedi et
al. 2016) 2.49 350 5.5 2 429.1 15.9 449 ± 16 71.1 ± 4.7 2.49 350 5.5 5 501.4 17.3 486 ± 31 92.5 ± 3.1
2.49 350 5.5 9 434.8 19.1 388 ± 15 68.3 ± 3.9 3.55 500 5.5 2 454.6 17.2 395 ± 50 84.7 ± 3.1
3.55 500 5.5 5 499.3 19.0 402 ± 20 92.0 ± 1.7 3.55 500 5.5 9 487.9 19.6 310 ± 11 60.3 ± 2.1
4.62 650 5.5 2 495.7 16.4 427 ± 8 88.2 ± 3.9 4.62 650 5.5 5 456.7 19.7 369 ± 11 95.0 ± 3.4
4.62 650 5.5 9 465.2 20.1 355 ± 12 53.6 ± 3.1
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Consequently, continuous AD reactors were operated with full-scale WAS to systematically assess the impacts of FNA pre-treatment on AD performances (Wei et al. 2018, Zhang et al. 2019). Two identical lab-scale AD reactors were operated in Wei et al. (2018) study, receiving FNA pre-treated and untreated WAS, respectively. Due to the enhanced WAS biodegradability by the FNA pre-treatment (FNA = 1.8 mgN/L, treatment time = 24h), VS destruction of the experimental AD was enhanced by 17 ± 1% (from 29.2 ± 0.9% to 34.2 ± 1.1%) and the methane production increased accordingly, by 16 ± 1%. The capability of FNA in disintegrating EPS and cell membranes brings additional benefits for the AD process. The dewatered solid content of effluent sludge increased from 12.4 ± 0.4% to 14.1 ± 0.4% (indicator of dewaterability). A 2.1 ± 0.2 log improvement in pathogen reduction was also attained in the long-term study. Moreover, as FNA pretreatment could accelerate the WAS hydrolysis rate during digestion, its potential on the digestion intensification was further explored (Zhang et al. 2019). Two parallel lab-scale reactors were operated at different HRTs. It was demonstrated that the HRT of AD could be reduced to as low as 7.5 days, by applying the FNA pre-treatment at a higher FNA - concentration of 6.1 mgN/L (NO2 = 250 mgN/L, pH=5.0) and treatment time of 24 hours (Zhang et al. 2019). Even compared with the control AD reactor with higher HRT (15 days), the experimental AD with FNA pre-treatment could achieve substantially higher VS destruction and specific methane production (more than 30%) under much shorter HRT (12 days and 7.5 days), promising to substantially increase AD capacities. It was also demonstrated in the continuous reactor studies that, pH adjustment after FNA treatment is not required. Acidification of AD is unlikely, as enhanced VS destruction led to increased alkalinity (Zhang et al. 2019).
In a progressive manner, the effectiveness of FNA pre-treatment on AD performance was recently demonstrated at pilot-scale (Meng et al 2019). Two pilot-scale AD reactors were continuously operated for one year at a full-scale WWTP, one receiving full-scale WAS (control) and another receiving WAS pre-treated for 24 h at an FNA concentration of 4.9-6.1 mgN/L (nitrite = 250 mgN/L, pH = 5.0, T = 22–30 °C). Results confirmed the enhancive effect of FNA pre-treatment on methane production (37±1%) and VS destruction (47±1%) at pilot- scale. Equally importantly, the pilot study, for the first time, showed that FNA pre-treatment significantly reduced the viscosity of WAS, by 58±7% (infinite shear rate), likely due to the solubilization of the WAS. The viscosity of the digested sludge was also reduced by 76±4% (infinite shear rate), caused by both enhanced VS destruction and inherent changes of the
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digested sludge. The reduction in viscosity is important as it allows the digester to be operated at a higher solids concentration, thus further increasing the capacity of anaerobic digester.
Apart from enhancing AD performance, due to capabilities of promoting WAS solubilization and digestibility, the FNA pre-treatment was reported efficient in enhancing Short Chain Fatty Acid (SCFA) productions (by 3 to 4 times) during anaerobic fermentation, too (Li et al. 2016, Okoye et al. 2018, Zhao et al. 2015). Although the FNA pre-treatment with WAS could also enhance AeD performance, some limitations render this approach unsustainable in AeD system. FNA pre-treatment (FNA = 2.02 mgN/L, Time = 24 h) enhanced the VS destruction of WAS greatly by up to 56% during AeD batch studies, as a result of enhanced digestibility (Pijuan et al. 2012, Wang and Yuan 2015). However, unlike in WWTPs with AD, a wastewater stream containing a high ammonium or nitrite concentration is not available in WWTPs with AeD, making the approach unsustainable in such a system. Moreover, the requirement of an additional pre-treatment unit is hindering the application of the FNA approach in AeD, considering AeD is popular for its low requirement for capital and operational management.
A novel operational approach was recently proposed and shown to enhance the AeD performance by in situ production and treatment of FNA during AeD (Duan et al. 2019a). This was achieved by inducing self-sustained nitrite accumulation via a single spike of nitrite to aerobic digesters operated at a natively low pH (<5.5), which forms a relatively high level of FNA. The accumulated nitrite, along with the low pH, will maintain the presence of FNA in the digester, which will keep up the suppression on NOB. The resulting consistent presence of FNA in the range of 0.1-1.0 mgN/L in the aerobic digestors enhanced the VS destruction by 35.0 - 38.4%, nitrogen removal by 58.5 - 70.8% and pathogen reduction by approximately 1 log. Compared with other sludge pre-treatment technologies, this strategy demonstrated that FNA production and treatment can be carried out in situ in the digester to enhance AeD performance. Since the nitrite production is self-supporting, no additional ongoing costs are incurred. The application of the proposed strategy could significantly benefit WWTPs by reducing sludge production volume, removing nitrogen, and improving pathogen removal and final sludge stability.
4.3 Production of FNA from wastewater
With many applications of FNA, it is important to note that nitrite can be produced on-site in WWTPs from AD liquor, which often contains high-concentration of ammonium (e.g., 0.8-1.2
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+ g NH4 −N/L). Commonly limited by the available alkalinity, around 50% of ammonium in AD liquor can be oxidized. By using online pH controller, nitritation process can sustain more than 90% conversion of the ammonium to nitrite (conversion to FNA requires a small amount of acid) in SBRs (Law et al. 2015). The nitritation was achieved through alternately inhibitive effects of FA and FNA on NOB reaction as well as pH variation due to AOB reaction in each SBR cycle. Extension of a two-step nitrification model further established the influent/operation space for the maintenance of the nitritation (effluent nitrite accumulation over 90% at steady state), i.e., minimal ammonium concentration of 54 mM (750 mg + − + NH4 −N/L) with an HCO3 /NH4 molar ratio range of 1.0−1.3 in influent under operational conditions of DO > 2.0 mg/L, SRT = 10 days, and T = 25 °C (Zheng et al. 2018). Low DO and short SRT are usually considered important to suppress NOB in a nitritation process. However, attributed to in situ FNA inhibition, it is strongly recommended to apply sufficient oxygen supply and retain adequate AOB biomass in the FNA/nitrite production SBR. Moreover, FNA can be produced similarly from other high strength wastewater streams, such as urine and landfill leachate (Zheng et al 2017).
4.4 The FNA-based applications are feasible and economically viable in wastewater treatment systems
The use of FNA in wastewater treatment is pragmatic and remunerative for WWTPs due to the resulting benefits and in situ production of FNA. For different sizes or configurations of WWTPs, FNA can be applied with flexibility to reach optimal outcomes.
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Figure 2. Flexible implementations of FNA-based technologies in different WWTPs.
In small-sized WWTPs, AeD is normally applied for its low capital cost, suitability for small amount of excess sludge, and simple operation and management (LeBlanc et al. 2009). In this scenario, FNA could be applied as Option A, illustrated in Figure 2. Sustained nitrite/FNA accumulation can be achieved via a single spike of nitrite to aerobic digester operated at a natively low pH (<5.5). Sustained FNA accumulation in the aerobic digester is capable of enhancing VS destruction, nitrogen removal and pathogen removal, and increasing stability of digested sludge (i.e. reduced SOUR). Improved VS destruction could directly reduce the cost of water utilities on sludge disposal. The improved pathogen removal performance demonstrated the process could potentially overcome one of the major disadvantages of aerobic sludge digestion. The improved digested sludge quality increased the possibilities for biosolids reuse (e.g., land application), which, again, could save cost on sludge handling. Furthermore, the strategy improves nitrogen removal. The supernatant after dewatering of aerobic digested sludge is normally returned back to the mainstream wastewater treatment process and may constitute up to 20–30% of the overall WWTP nitrogen load. Enhanced nitrogen removal with the proposed strategy during aerobic digestion could potentially reduce 10-15% nitrogen loading to the mainstream reactor, ensuring a better effluent quality. It is important that the proposed operation strategy incurs very little additional costs and is applicable to any existing
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aerobic digesters. It maintains the major advantage (i.e., simple operation and low cost) of conventional aerobic digesters while achieving substantial enhancements to conventional aerobic digesters.
In large-sized WWTPs, anaerobic digestion is commonly practiced. Most of those WWTPs recycle the AD liquor back to the head of the plant, while increasingly more plants have alternatively implemented a separate sidestream unit to treat the liquor (Janus and van der Roest 1997). In the case that the sidestream unit is not available, nitrite can be added externally for the FNA treatment, as illustrated in Option B (Figure 2). It should be noted the associated cost of chemical addition is relatively small, compared with the resulted economic benefits, and thus is cost-effective (Wei et al. 2018). FNA pre-treatment of WAS benefits WWTPs by
improving AD performance in the aspects of VS destruction, CH4 production, pathogen removal and sludge dewaterability. Similar to FNA assisted AeD, enhanced VS destruction, sludge dewaterability and pathogen removal directly and indirectly reduce the cost of water utilities on final sludge disposal. Furthermore, more energy could be recovered from wastewater via the increased CH4 production, which could contribute to energy neutrality of WWTPs operations.
WWTPs with AD and sidestream treatment units are highly desirable for the application of the FNA-based processes. By utilizing nitrite in the waste stream, the FNA-based technology set a good example for circular economy. Depending on specific needs, conservative, moderate or aggressive strategies can be considered. The conservative option (Option C) is taking one step further from Option B, by using the nitrite produced from AD liquor. Option C shares all the benefits of Option B, while further eliminates the cost and nitrogen loading of externally added nitrite.
The moderate strategy (Option D) is to implement sludge treatment to simultaneously achieve mainstream nitrogen removal via nitrite shunt, and enhanced AD performance. In addition to the benefits from enhanced AD performance discussed for Option C, the mainstream nitrite shunt provides extra advantages to WWTPs. For many WWTPs, receiving wastewater with insufficient COD restricted the conventional Biological Nutrient Removal (BNR) performance and resulted in a relatively high total nitrogen in the effluent (Peng et al. 2004). To meet the increasing regulatory nitrogen discharge standards, external carbon source (e.g. methanol) needs to be supplemented, which is a significant operational cost for WWTPs (Gu et al. 2017).
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Besides, aeration for nitrification is one of the most energy consuming processes and generates considerable cost in WWTPs (Longo et al. 2016). Compared with traditional BNR process, the mainstream nitrite shunt reduces carbon source consumption by 40%, aeration oxygen demand for nitrification by 25% and sludge production by 30 50% (Antileo et al. 2013, Turk and
Mavinic 1986, Turk and Mavinic 1989). The nitrite ∼shunt process achieved by the FNA- facilitated approach is therefore highly beneficial for WWTPs.
A more aggressive approach is to attain mainstream PN/A and enhanced AD at the same time (Option E). Compared to mainstream nitrite shunt, mainstream PN/A is an even more energy and carbon efficient nitrogen removal process that reduces the COD requirement by nearly 100%, ease the aeration oxygen demand by 60% and reduce sludge production by 80% (Jetten et al. 1997, Wett 2007). By decoupling carbon and nitrogen removal, the potential elimination of COD demand provides unique opportunities for energy positive nitrogen removal (Cao et al. 2017). Not only allowing carbon-efficient nitrogen removal, the decreased aeration energy consumption and less excess sludge production also significantly contribute to operational cost reduction, which is also greatly favourable for WWTPs. However, the difficulties of NOB suppression should be acknowledged. While the concepts of mainstream nitrite shunt or PN/A have been proposed for decades (Hellinga et al. 1998, Jetten et al. 1997) and a large number of lab-scale studies have been reported, its full-scale application is scarce. There are many obstacles for the implementation that have been reviewed elsewhere (Cao et al. 2017, Ma et al. 2016). The more aggressive Option E would pose more risks than Option D. Firstly, the slow growth rate and sludge yield of anammox bacteria requires special operations to retain the anammox bacteria in mainstream wastewater treatment conditions (Ma et al. 2016); Secondly, unstable mainstream NOB suppressions due to various reasons, e.g., operation failures, bacterial adaptations could jeopardize the process (Ma et al. 2017b). In the scenarios that mainstream NOB suppression fails, Option D is still capable of removing nitrogen by conventional nitrification and denitrification while Option E will not be able to do so.
Economic analyses have been carried out for Options B, C, D & E (Duan et al. 2018, Wang et al. 2016a, Wang et al. 2013a, Wei et al. 2018). While the FNA approaches incur extra operational costs due to chemical addition (nitrite and acid for Options B, and acid for Options C-E) and capital costs for a sidestream nitritation unit (for WWTPs without sidestream treatment) and sludge treatment unit, the approaches are overall economically favourable due to significant savings in aeration energy, methanol addition and sludge handling.
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5. The use of FNA in membrane and algae systems
Attempts were recently made to use FNA as i) a cleaning reagent in recovering the membrane filtration performance and ii) a facilitator in harvesting wastewater grown algae.
Membrane-based technology has been widely applied in wastewater treatment for water reclamation. Specifically, microfiltration (MF) and ultrafiltration (UF) membranes are commonly implemented in membrane bioreactors, followed by further purification by nanofiltration (NF) and reverse osmosis (RO) for water reclamation (Bartels et al. 2005, Garcia et al. 2013). Forward osmosis (FO) is another membrane-based treatment technology that has gained interests for its ability to pre-concentrate raw wastewater for biogas production (Ansari et al. 2016, Zhang et al. 2014). These membrane-based technologies apply hydraulic pressure or osmotic pressure to drive water molecules to pass across the membrane which results in the formation of fouling layer on the surface and inside the pores of the membrane (Lee et al. 2010, Mi and Elimelech 2010). The fouling formation reduces the water flux of the membrane and consequently deteriorates the performance of the membrane-based process (Alsvik and Hägg 2013, Wang et al. 2007). Membrane fouling is categorised into biofouling, inorganic fouling (scaling), organic fouling and colloids fouling, which may co-deposit depending on the characteristics of the receiving wastewater (Gkotsis et al. 2014, Wang et al. 2014c). Other than physical cleanings, fouled membranes are commonly treated by chemical cleanings, including acid, alkaline, and oxidants cleanings (Filloux et al. 2015, Le-Clech et al. 2006). Specifically, acid removes the inorganic matters, and alkali eliminates the organic foulants (Filloux et al.
2015, Le-Clech et al. 2006). Oxidants (e.g. H2O2 and NaClO) remove biological and organic matters (e.g., EPS) (Le-Clech et al. 2006, Yoon et al. 2013). However, the large volume of cleaning reagents could result in significant operational costs and environmental issues for disposal (Filloux et al. 2015, Wang et al. 2014c). This led to the exploration of alternative ‘green’ cleaning reagents. FNA becomes an attractive solution due to: i) the biocidal nature of FNA (Jiang et al. 2011b, Jiang and Yuan 2013); ii) the acidic nature of FNA (Filloux et al. 2015); and iii) the capability of FNA in breaking down EPS (Wang et al. 2014c, Zhang et al. 2015).
Studies have been conducted in evaluating the potential of FNA as a cleaning reagent, in which, the efficiencies are depending on: (i) the concentration of FNA; (ii) exposure duration of FNA; and (iii) propensity of the fouling layer (Ab Hamid et al. 2018, Filloux et al. 2015). Filloux et
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al. (2015) investigated the efficiency of FNA in treating fouled RO membranes, which suffered biofouling and scaling, from the full-scale water recycling and seawater desalination plants (Filloux et al. 2015). The FNA cleaning test was conducted at various pH levels (from pH 2 – - 6) at a nitrite concentration of 50 mgNO2 -N/L. The performance was compared with other cleaning reagents, i.e., NaOH, HCl, and citric acid. The FNA cleaning (pH = 3, and nitrite 50 - mgNO2 -N/L) was more efficient than NaOH (pH 11.0) to remove organic foulants. FNA cleaning resulted in high biomass inactivation (92–95%) and polysaccharides removals (68– 90%) after 24 hours of exposure. In terms of scale removal, FNA cleaning is as efficient as the - commonly used descaling agent (HCl and citric acid). The FNA cleaning (50 mgNO2 -N/L and pH at 2.0 and 3.0) could dissolve ≥ 30 g/m2 of calcium, leaving 1% of calcium quantified on the membrane surface after the cleaning. It is important to note that FNA can simultaneously achieve both biofouling removal and descaling in a one-step cleaning strategy (Filloux et al., 2015). In another study, FNA cleaning was applied to treat FO membranes that suffered biofouling, due to direct filtration in pre-concentrating raw wastewater (Ab Hamid et al. 2018). It was shown the 1-hour FNA exposure inactivated 43% of the biomass, which increased to 90% and 96%, respectively, when the treatment time was increased to 8 hours and after 24 hours (Ab Hamid et al. 2018).
Although FNA shows promising results as a cleaning reagent, the cleaning potential and optimal treatment conditions still remain to be further evaluated. The FNA cleaning only recovered 74% of water flux performance of a fouled FO membrane after ten cycles of cleaning (1-hour cleaning for each cycle), whilst the water flux was recovered to >90% by other cleaning reagents (e.g. NaOCl, EDTA) (Ab Hamid et al. 2018, Holloway et al. 2007, Valladares Linares et al. 2012). The formation of dense and compact fouling layer on the membrane can limit the FNA penetration into the fouling layer (Ab Hamid et al. 2018, Filloux et al. 2015, Mi and Elimelech 2010). The effective exposure duration of FNA cleaning needs to be correlated to the characteristics of the fouling layer. Further optimization of FNA cleaning in terms of frequency and duration is necessary to improve the cleaning performance.
Application of FNA has also been beneficial for algae-based wastewater treatment systems in terms of algae harvesting. Conventional wastewater treatment commonly finishes at a secondary treatment with an apparently clear effluent that can meet the discharge limits. However, secondary effluent generally still contains non-negligible loads of inorganic nitrogen and phosphorus which may cause eutrophication and long-term considerations. Algae
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technology applied in wastewater treatment has been considered since 1950s (Oswald et al. 1957). The idea has been focusing on utilising these photosynthetic microorganisms to remove inorganic nitrogen and phosphorus and mitigate CO2 as one of the major greenhouse gases. Algae also has the ability to accumulate lipid droplets inside the biomass which can then be used as a feedstock for biodiesel production. This field has been attracting research and industrial interests around the world.
The recovery of energy and other products from algae biomass is considered a necessary step to make algae technology economically viable. Pre-treatment is widely investigated by researchers and industry to boost algae valorisation. FNA was used in several studies as a novel algae pre-treatment method for boosting both product (crude lipids and triacyl glycerides via extraction) (Bai et al. 2014, Bai et al. 2015) as well as energy (biogas via anaerobic digestion) recovery (Astals et al. 2015, Bai et al. 2016). The authors reported a lipid yield 2.4-fold higher
for algae treated with FNA (up to 2.19 mg HNO2-N/L) and a 51% increase in methane yield treated with FNA (2.31 mg HNO2-N/L) comparing to untreated biomass. The FNA effect on algae is explained as an oxidative stress that disrupts cell walls and even intracellular membranes which can act as barriers to both product extraction as well as anaerobic biodegradability of encapsulated organics. However, more research needs to be conducted to fully address the mechanism of the FNA effect on algae biomass and the process.
6. Knowledge gaps and opportunities
6.1 Fundamental knowledge on the chemistry of FNA
Fundamental understanding of chemical reactions between FNA and large molecules is insufficient and remains to be investigated. While studies have observed the FNA pre-treatment could disintegrate EPS and cell membranes (Wu et al. 2018, Zhang et al. 2015), the chemical mechanism for such reactions remains indistinct. It is important to understand how FNA changes the chemical structure and functional groups of the organic molecules in the activated sludge. The knowledge will provide a mechanistic explanation for the destruction of volatile solids including EPS and cell membranes to more biodegradable organics. Those more biodegradable organics could facilitate the hydrolysis, acidification and methane production in AD following the FNA pre-treatment. The destruction of complex organics can also provide insights into floc disruption and consequent lysis of the cell envelope, which in turn likely increases dewaterability. This mechanistic knowledge can contribute towards enhancing the
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efficiency of the established AD pre-treatment process by providing information on reaction stages, intermediate products and the associated timing of the reaction stages.
Furthermore, knowledge on the chemical property of FNA could provide unique opportunity for removing hardly biodegradable pollutants in wastewater. Conventional WWTPs relies on biological and physical processes to remove pollutants, such as nitrogen, phosphate, organics in wastewater. However, it is not feasible to remove hardly biodegradable and hydrophilic (low sorption affinity) micropollutants such as some pharmaceuticals, pesticides, phosphorus flame retardants, sweeteners, or corrosion inhibitor, commonly present in wastewater (Margot et al.
2015). Chemical reactions, such as UV/H2O2, ozone oxidation have been shown effective in removing a wide range of micropollutants. A recent study showed that FNA can chemically transform the sulfonamide antibiotics sulfamethoxazole (SMX) into several intermediates (Sun et al. 2019). The understanding of the chemical property of FNA could therefore lead to the potential application in removing micropollutants in wastewater, as additional benefits of the FNA based technology.
6.2 The impact of FNA sludge treatment on biological phosphorus removal
During BNR process, phosphorus is commonly removed simultaneously with nitrogen by enhanced biological phosphorus removal (EBPR) (Wu et al. 2014). While the FNA sludge treatment approach is effective in achieving energy and carbon efficient nitrogen removal (mainstream nitrite shunt or PN), the treatment might affect the simultaneous phosphorus removal. Firstly, the FNA in wastewater has been recognized as a strong inhibitor on biological phosphorus removal (Pijuan et al. 2010, Zhou et al. 2010, Zhou et al. 2011, Zhou et al. 2007). For example, Candidatus Acumulibacter is the most widely known phosphate accumulating organisms (PAOs), which contribute to the EBPR process. FNA concentration of approximately 0.5 ugN/L could cause 50% inhibition on all anabolic processes of Candidatus Acumulibacter (P release, PHA synthesis and glycogen degradation), while 100% inhibition occurred at FNA concentrations of approximately 6.0 ugN/L. These concentrations of FNA could present in mainstream nitrite shunt or PN system with nitrite accumulation (Pijuan et al. 2010). Secondly, the high FNA concentration (>1 mgN/L) in sidestream FNA treatment unit is likely biocidal and could inactivate PAOs in the treatment unit. Thirdly, FNA has a significantly stronger inhibitory effect on the aerobic metabolism of PAOs than that of Glycogen Accumulating Organisms (GAOs) (Table 1). In an EBPR system, GAOs often
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proliferate under the same conditions as PAOs and competing for substrate against PAOs. However, GAO metabolism does not include anaerobic phosphorus (P) release and subsequent aerobic (and/or anoxic) phosphorus uptake like PAO, therefore not contributing to the EBPR performance (Saunders et al. 2003). The different sensitivities of PAOs and GAOs to FNA inhibitions potentially provides a selective advantage for GAOs to outcompete PAOs in EBPR systems, which could lead to deterioration and, in extreme cases, failure of the EBPR process. However, the impact of the FNA sludge treatment approach on the EBPR performance has not been evaluated. More studies are required to assess the impacts of the FNA approach on the EBPR performance.
6.3 Constraints of lab-scale studies call for pilot- and full-scale trials
Apart from the use of FNA in sewer systems, other FNA studies were all conducted in laboratories (with the exception of Meng et al. (2019) at pilot-scale). Some constraints of the lab-scale studies should be realised. Most lab-scale mainstream NOB suppression studies used synthetic wastewater to assess the effectiveness of the FNA sludge treatment approach. However, by using synthetic wastewater, the impact of varying composition in the real wastewater could not be assessed. The change of nitrogen loading could influence the stability of NOB suppressions (Mutlu et al. 2013). Furthermore, real wastewater contains a wide range of microorganisms, which continuously inoculate the bioreactor. This could affect the microbial community in the reactor, especially the NOB community, which cannot be evaluated with the use of synthetic wastewater(Duan et al. 2019c). In lab-scale anaerobic digestors operation with FNA pre-treated WAS, while real WAS was used, the small capacity of the pre-treatment and anaerobic digestors may not be representative for larger scale applications. The hydraulic dynamics in a larger scale FNA pre-treatment unit and anaerobic digestors could be substantially different, which could affect the mass transfer and the effectiveness of the FNA pre-treatment. Furthermore, economic analyses for FNA-based technologies in wastewater treatment and sludge management were performed based on laboratory results. Full-scale or pilot-scale tests are still required to fully reveal the feasibility and potential of the FNA-based technologies.
6.4 Alternatives for acid addition
An important feature of the FNA-based approach is the sustainability by using nitrite produced from waste stream in WWTPs. However, using FNA still requires external acid addition. Apart
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from the extra operational cost incurred by the acid addition, it could potentially affect alkalinity balance in wastewater. In mainstream PN/A where alkalinity could not be supplemented by denitrification, additional acid addition could limit the progress of nitrification by consuming alkalinity. The potential consequences need to be evaluated in larger scale and with real wastewater streams.
Alternatives for the acid addition should be explored. Duan et al. (2019a) demonstrated that the FNA pre-treatment could be carried out in an AeD by nitrite accumulation under low pH conditions, as the ammonium oxidation consumes alkalinity. In the FNA treatment unit for either WAS (or WAS +PS) pre-treatment or NOB inactivation, AOB present in the WAS could utilize the ammonium that exist in the sidestream PN effluent. Therefore, if aeration can be provided, the acidification process could be carried out spontaneously for FNA treatment. This is a possible line for future research to reduce the external acid addition and maintain the internal balance of alkalinity in wastewater treatment.
6.5 Intrinsic limitations and opportunities
Some limitations are intrinsic to the FNA-based technology, and the potential solutions could lie in the combinations with other methods. For example, while FNA has been applied to inactivate NOB to achieve carbon and energy efficient nitrogen removal in mainstream wastewater treatment, NOB could develop resistance to the FNA treatment. The development of adaptation to suppression is theoretically inevitable, as would occur with any other suppression approaches. In this regard, FNA-based treatment was alternated with FA-based treatment to address the identified adaptation issue (Duan et al. 2019b). Its combination with other strategies should also be investigated. Furthermore, the FNA pretreatment assisted AD is not expected to improve the quality of the digestate to meet the strict biosolids class A standards, due to the limited improvement to pathogen removal. Therefore, studies have explored the possibility of combining FNA pretreatment with heat pretreatment to enhance the pathogen removal (Wang et al. 2014a). Similarly, the FNA based approach could achieve approximately 1-log (90%) inactivation of the anaerobic biofilm in the sewer system while further increasing the inactivation efficiency seems not cost-effective. Research demonstrated the combination of
H2O2 with FNA greatly enhanced the inactivation of microorganisms to about 2-log (99%) of microbial inactivation (Jiang and Yuan 2013).
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7. Conclusions
The knowledge of inhibitory and biocidal effects of FNA on microbes, and the corresponding mechanisms were improved substantially in the last decade. The advancement of such knowledge has led to the development of a series of FNA-based applications enhancing wastewater management. In sewer systems, FNA has been successfully used to control sewer corrosion and odor as a cost-effective solution. In wastewater treatment systems, in situ produced FNA can be applied to enhance the system in many aspects. FNA sludge treatment selectively washout NOB to achieve carbon and energy efficient nitrogen removal. The FNA pre-treatment of WAS can significantly improve the sludge reduction and energy recovery in AD. In AeD, sustained FNA accumulation can be achieved in situ to increase sludge reduction and nitrogen removal. Furthermore, in membrane-based systems, FNA has been identified as a promising and green reagent that simultaneously address the scaling and organic fouling of membranes. FNA could also facilitate the harvesting of wastewater algae, in an algae system.
While significant progresses have been made in recent years, some limitations of FNA-based applications need to be realized. The fundamental understanding on the chemical reaction of FNA with organic and inorganic compounds are still lacking, restricting the expansion of FNA applications to more areas. Also, the FNA-based technology might potentially affect the biological phosphorus removal in wastewater treatment, which however has not been investigated to date. While nitrite could be produced from wastewater, acid addition is often required, which could result in extra operational costs and affect the alkalinity balance. It should also be noted that most of the FNA based technologies (except for sewers) have been conducted in laboratory environment, larger scale studies should be carried out to fully verify the feasibility of those technologies.
Further studies on the chemical reactions of FNA with organic and inorganic compounds are required. Possibilities of achieving simultaneous carbon and energy efficient nitrogen and phosphorus removal should be investigated. Larger scale trials and optimizations are required to verify the full potential of FNA-based technologies. Alternative approach to acid addition is promising and should be explored. More importantly, the future of FNA application should also be looking at the combination with other method to overcome the intrinsic limits. For example, FNA + FA to overcome NOB adaptation, FNA + heat to achieve cleaner digestate
(pathogen), FNA + H2O2 to achieve maximal microorganism inactivation.
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Acknowledgements
The authors acknowledge the financial support provided by the Australian Research Council through Linkage Projects LP150101337 and LP130100361, and Discovery Project DP120102832 and DP180103369. Prof. Zhiguo Yuan is a recipient of the ARC Australian Laureate Fellowship via FL170100086. Dr. Haoran Duan acknowledges the postdoctoral fellowship support of The University of Queensland.
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