Approaching easy water disinfection for all: Can in situ electrochlorination outperform conventional chlorination under realistic conditions?

Approaching easy water disinfection for all: Can in situ electrochlorination outperform conventional chlorination under realistic conditions? Aksana Atrashkevicha,b, Absar Aluma,c, Robert Stirlinga,b, Morteza Abbaszadegana,c, Sergi GarciaSeguraa,b,* aSchool of Sustainable Engineering and the Built Environment, Arizona State University, Tempe, AZ 85287-3005, USA bNanosystems Engineering Research Center for Nanotechnology-Enabled Water Treatment, Tempe, AZ 85287-3005, USA cWater and Environmental Technology Center, Arizona State University, Tempe, AZ 85281, USA Article submitted to be published in Water Research https://doi.org/10.1016/j.watres.2023.121014 Corresponding author: *e-mail: [email protected] (Dr. Sergi Garcia-Segura) © 2024. This manuscript version is made available under the CC-BY-NC-ND 4.0 license https://creativecommons.org/licenses/by-nc-nd/4.0/ Abstract 1 Electrochlorination has gained research interest for its potential application as decentralized water 2 treatment. A number of studies have displayed promising efficiency for water disinfection. However, a 3 comprehensive comparison of in situ electrodisinfection to existing disinfection techniques, particularly 4 under realistic water composition and flow rates, still needs additional research efforts. The aim of this 5 study is to evaluate in situ electrochlorination while comparing the treatment with conventional chemical 6 chlorination for point-of-entry decentralized disinfection at household level. An electrochemical flow cell 7 reactor was operated in a single pass mode considering water flows and water consumptions for a household 8 of four family members. Disinfection efficiency assessment of both electrochemical and chemical 9 chlorination were conducted using bacterial and viral surrogate, E. coli and MS2 bacteriophage. 10 Furthermore, a techno-economic analysis was conducted, using levelized cost of water, to compare two 11 electrochemical chlorination scenarios (i.e., electrical grid energy use, and solar panel powered system) and 12 benchmarked against the baseline treatment of chemical chlorination. The findings revealed significantly 13 increased inactivation efficiency of in situ electrochlorination over conventional chlorination (p-value < 14 0.05). The synergetic impact of radicals and chlorine, and contribution of high chlorine concentration at 15 acidic pH near anode surface were identified as key factors that could enhance disinfection performance of 16 in situ electrochlorination. . The techno-economic analysis demonstrated that electrochemical treatment, 17 when operated using renewable energy sources, is not only a more environmentally sustainable approach, 18 but also emerges as a more economically feasible solution for decentralized water treatment application. 19 The results highlight that in situ electrochlorination is a more advanced alternative to decentralized water 20 chlorination. However, further fundamental research on products and by-products formation under various 21 water matrices is required. 22 23 24 Key words: Electrochlorination, electrodisinfection, E. coli, Bacteriophages, electrochemical 25 water treatment, water reuse, active chlorine species 26 1. Introduction 27 Access to clean water and sanitation is identified as one of the sustainable development goals of 28 United Nations (UN, 2023). However, the complete achievement of this goal in certain regions by the 2030 29 deadline is in jeopardy due to the different challenges (dos Santos et al., 2023). Approximately 26% of the 30 global population lacks access to safely managed drinking water services, and 46% lacks access to safely 31 managed sanitation (UN, 2023). Unsatisfactory water sanitation practices lead to an increasing number of 32 waterborne disease outbreaks due to presence of microbial pathogens. The economic burden linked to the 33 implementation and maintenance of centralized water treatment and distribution networks is one of the 34 major barriers to ensure equitable water access for all, especially in economically disadvantaged areas. In 35 this frame, low power consumption technologies may arise as the most optimal tool to ensure effective 36 water disinfection in decentralized settings. 37 Electrochemical water treatment systems have garnered profound interest due to their 38 effectiveness in removing recalcitrant pollutants and pathogens, along with their safety and operational 39 convenience (Garcia-Segura et al., 2018; Hand and Cusick, 2021; Zuo et al., 2023). Microbial inactivation 40 is attributed to the generation of disinfecting agents during electrochemical treatment. Electrochemically41 driven water treatment offers promising prospects for decentralized treatment in rural regions (Bhattacharya 42 et al., 2021; Otter et al., 2019). Furthermore, energy efficiency and self-sufficiency can be achieved when 43 electrified technologies are operated using local renewable energy sources (e.g., household solar panels). 44 Therefore, electrochemical treatment can be operated off the water and electrical grid. In situ 45 electrochlorination is an indirect electrochemical treatment process that generates chlorine species from 46 ubiquitous chloride anions (i.e., present in almost any type of water source) for local water disinfection. 47 Chloride anions (Cl-) can be oxidized on the anode surface following reaction (1). Electrogenerated chlorine 48 (Cl2(aq)) reacts with water yielding hypochlorous acid (HOCl) (2). The speciation of HOCl/OClspecies is 49 defined by the solution pH driven by the acid-base equilibria (3) with a 𝑝𝐾𝑎= 7.5 at 25 ˚C (Scialdone et 50 al., 2021). Both species HOCl and OClhave high disinfectant capacity and can effectively inactivate 51 microorganisms (Overholt et al., 2018). Indeed, conventional chemical chlorination disinfection systems 52 rely on the activity of HOCl/OClfor water disinfection. 53 2Cl−→Cl2+ 2e− (1) Cl2(aq)+ H2O → HOCl + H++ Cl− (2) HOCl ⇋ H++ OCl− 𝑝𝐾𝑎= 7.5 at 25 ˚C (3) Water disinfection represents the final step within the conventional centralized water treatment 54 processes. Various treatment agents are typically employed to achieve primary water disinfection including 55 active chlorine, combined chlorines, chlorine dioxide, ozone, and ultraviolet (UV) radiation (Crittenden et 56 al., 2012). Among these methods, only chlorine-based treatment methods offer long-lasting water 57 disinfection within water distribution networks and equipment (Martínez-Huitle and Brillas, 2008). 58 However, centralized chlorination methods have limitations associated with handling of hazardous 59 chemicals and halogenated disinfection by-products (DBPs) formation. In this frame, in situ 60 electrochlorination appears as a user-friendly approach that partially removes the limitations of 61 transportation and storage of hazardous chemicals. Other advantages of in situ electrochlorination such as 62 straightforward process automation and the possibility of utilizing renewable energy sources make in situ 63 electrochlorination highly suitable for decentralized water disinfection (Otter et al., 2019). Moreover, 64 integration of electrochlorination into water reuse systems may provide a desirable safety net by 65 implementing water disinfection to enable safe water recycling, therefore reducing the demand for 66 freshwater resources (Huang et al., 2016; Otter et al., 2020). However, in situ electrochlorination has a 67 similar risk of organic DBPs formation as chlorination treatment. Currently, 9 organic DBPs are regulated, 68 including 5 haloacetic acids (HAAs) with the total maximum of 0.060 mg L-1 and 4 trihalomethanes 69 (THMs) with the limit of 0.080 mg L-1 (EPA, 2009). The generation of organic DBPs is dependent on water 70 characteristics, including concentrations of disinfectant and organics, contact time, temperature, pH, and 71 bromide content (Crittenden et al., 2012; Sedlak and von Gunten, 2011; Zhang et al., 2013). Water matrices 72 have a huge impact on DBPs formation and decreasing the level of its precursors might be a vital strategy 73 to reduce its generation (Yin et al., 2020). Inorganic DBPs such as toxic oxyanions (i.e., ClO3-, ClO4-) might 74 be formed simultaneously during in situ electrochlorination. The World Health Organization (WHO) set up 75 guideline values for ClO3and ClO4of 700 μg L-1 and 70 μg L-1, respectively (WHO, 2016a, 2016b). 76 Nevertheless, anode material and charge transfer play a huge role in the production of toxic oxyanions 77 during electrochemical treatment (Bagastyo et al., 2011; Cano et al., 2011; Hao et al., 2022; Scialdone et 78 al., 2021). It has been demonstrated that insignificant amount of toxic oxyanions might be formed if using 79 Ti/RuO2 anode material if low charge transfer applied (Yoon et al., 2015). 80 Previous research has demonstrated higher effectiveness of in situ electrochlorination over 81 conventional treatment for water disinfection under laboratory conditions (Diao et al., 2004; Ghernaout et 82 al., 2019; Li et al., 2004). However, a detailed disinfection comparison (i.e., controlled pH, contact time, 83 disinfectant dose) of in situ electrochlorination with conventional chlorination under operating settings 84 applicable for household-level systems while using freshwater composition (i.e., low salt content) has 85 received limited attention. Consequently, there are questions that remains unanswered regarding the 86 advantages of the in situ electrochlorination over conventional chlorination under operating conditions that 87 closely resemble real-world scenarios. This research aims to address this knowledge gap by examining the 88 proof of concept of in situ electrochlorination in order to provide a comprehensive assessment of its 89 disinfection effectiveness and economic feasibility as a decentralized treatment. 90 To bridge the gap between laboratory-scale and realistic conditions (Garcia-Segura et al., 2020), 91 we selected the water and operating parameters for the electrochemical flow cell that aligned with 92 residential use at household-level. The daily water consumption volumes for a household of a family of 93 four is estimated to be 432 L, therefore being this daily volume selected as a realistic water consumption 94 volumes figure to be considered when exploring treatment applicability (Crouch et al., Jacobs, and Speight 95 2021; Willis et al. 2011). Viability of the system was assessed using freshwater conductivity with chloride 96 content of 50 mg L-1, which is lower than drinking water maximum guideline (i.e., 250 mg L-1) and within 97 average (i.e., 20 - 150 mg L-1) (DES, 2010; EPA, 1979; WHO, 2003) concentrations. Furthermore, the 98 water matrix included other most commonly occurring ions in freshwater matrices such as Ca2+, Mg2+, SO4299 , Na+ (Olichwer et al., 2013; Wetzel, 2001). In order to comply with the WHO requirements for chlorine100 based water disinfection at household level, we determined generated dose of active chlorine species 101 ensuring it remains below 2 mg L-1 (Ocasio and Sedlak, 2022; WHO, 2017). This approach aimed to assess 102 the system’s performance while using realistic water consumption and adhered to drinking water quality 103 parameters. 104 Escherichia coli (E. coli) bacteria and MS2 bacteriophage were selected as surrogates of bacterial 105 and viral waterborne pathogens. E. coli is the most commonly used bacteria to assess water treatment 106 processes due to its easy detection method and value as a standard indicator of fecal contamination. 107 Bacteriophage MS2 has been used as a viral surrogate to evaluate inactivation effectiveness against viruses 108 due to its easy detectability, morphological resemblance, and similarity in resistance to chlorine exhibited 109 by various pathogenic human viruses (Alum et al., 2022, 2011; Yavarmanesh et al., 2013). 110 We assessed the disinfection performance of in situ electrochlorination by conducting a 111 comparative analysis with conventional chlorination and disinfection in electrochlorinated waters at 112 household-level settings, while ensuring consistent water composition and pH conditions. Moreover, 113 microbial inactivation in absence of chloride content was investigated. The enhanced disinfection 114 performance of in situ electrochlorination elucidated possible additional factors that contributed to 115 microbial inactivation besides active chlorine in the bulk solution. Furthermore, a techno-economic analysis 116 was conducted to assess the competitiveness of electrochlorination technologies in comparison to 117 conventional chemical chlorination. 118 119 Materials and methods 120 2.1 Chemicals and Reagents 121 Experiments were conducted using two types of water matrices A and B with composition and 122 physical-chemical characteristic summarized in Table 1. All the chemicals used in this study were obtained 123 from Sigma-Millipore (St. Louis, MO) and were managed and stored following safety data sheets protocols. 124 Representative water samples were prepared using magnesium sulfate heptahydrate (MgSO4 ∙ 7H2O, 125 ≥98%), calcium chloride dihydrate (CaCl2 ∙ 2H2O, ≥99%), sodium chloride (NaCl, ≥99%), calcium sulfate 126 dihydrate (CaSO4 ∙ 2H2O, ≥99%), and sodium sulfate (Na2SO4, ≥99%). For the conventional chlorination 127 experiments hypochlorite (NaOCl, 10-15% available chlorine) and 0.1 M hydrochloric acid (HCl, 37%) 128 solution were used to reach residual chlorine concentration of 1.75±0.05 mg L-1 and water pH of 7.0±0.2, 129 respectively. Sodium thiosulfate (Na2S2O3, ≥99%) solution with concentration of 0.1 M were used for 130 neutralizing the residual disinfectant/ chlorine at contact time of interest. To keep stable pH and osmotic 131 pressure of diluted samples identical to biological fluids during microbiological assay, 0.5 M phosphate 132 buffer saline (PBS) at pH 7.4 was prepared using potassium chloride (KCl, ≥99%), sodium chloride (NaCl, 133 ≥99%), sodium phosphate dibasic (Na2HPO4, ≥99%) and potassium phosphate monobasic (KH2PO4, 134 ≥99%). Tryptic soy broth (TSB) (cat # 211825 BD Bacto, Sparks MD) prepared according to the 135 manufacturer instructions was used for the growth and daily maintenance of E. coli culture. The spiking of 136 microbial stocks (E. coli or MS2) in test waters resulted in the introduction of trace amounts of TSB causing 137 elevated total organic carbon concentration (TOC) shown in Table 1. Microbial analyses including pour 138 plate method (E. coli) and double-layer plaque assay (MS2) were performed using Tryptic soy agar (TSA) 139 (cat # 22091 Sigma Millipore, St. Louis, MO) prepared according to the manufacturer instruction. Free 140 chlorine was analyzed using N,N diethyl-p-phenylene diamine (DPD) reagent supplied by HACH for 141 colorimetric method quantification (APHA et al., 2017). All solutions were prepared with ultrapure water 142 provided by an ultrapure water system Elga Water with resistivity >18.2 MΩ cm at 20 °C. 143 144 Table 1 145 Influent water characteristics and ionic composition. 146 Water Type (Water matrices) Initial pH Conductivity (μS cm-1) Cations (mg L-1) Anions (mg L-1) TOC (mg L-1) Ca2+ Mg2+ Na+ ClSO42E. coli MS2 A (with Cl-) 6.0±0.2 300±20 20 9.7 9.4 50 38 1.36 1.26 B (without Cl-) 20 9.7 23 0 134 147 2.2 E. coli and MS2 stock preparation and assays 148 The pure cultures of all bacterial and bacteriophage strains were obtained from the American Type 149 Culture Collection (ATCC, Manassas, VA) (S3). All the pure cultures were prepared and maintained 150 according to the ATCC instruction. Bacterial strains E. coli (ATCC 25922) were incubated in TSB for 18151 24 hours at 37 ˚C. Afterwards, the culture was separated from TSB by centrifuging at 4000 rpm for 10 152 minutes. The supernatant was discarded, and bacterial cell pellet was resuspended with water matrix 153 containing ionic composition required for the experiment of interest (i.e., either matrix A or B, Table 1), 154 creating 10 mL volume stock solution. Culturable counts for E. coli in samples were analyzed using the 155 pour plate method in TSA. Plates were incubated in inverted position at 37 ˚C and the colony forming units 156 (CFUs) were counted after 18-24 hours. Sterile plastic bags were used during incubation to prevent false 157 positive results. 158 For bacteriophage MS2 assays, the host bacteria E. coli (ATCC 15597) in TSB was incubated for 159 3 hours at 37 ˚C to achieve log phase culture. One mL of host bacterial culture and one mL of test water 160 sample were added and mixed in TSA tempered at 48 ˚C. Mixture was poured onto a TSA plate and allowed 161 to solidify for 30 minutes. Solidified plates were incubated in inverted position at 37 ˚C and plaque forming 162 units (PFUs) were counted after 18-24 hours. 163 164 2.3 Analytical instruments 165 The pH and conductivity were measured with a Thermo Scientific Orion Star A221 pH-meter and 166 A322 conductivity meter, respectively. The free chlorine (Cl2/HOCl/OCl-) concentrations were measured 167 utilizing DPD colorimetric method (APHA et al., 2017). Absorbance of 530 nm wavelength was detected 168 using a HACH DR 6000 UV-vis spectrophotometer. The total organic carbon analyzer TOC-LCSH FA 169 CN200 Shimadzu was used to measure the TOC concentrations in the filtered water matrices spiked either 170 with bacteria or bacteriophage cultures. The water samples before TOC analysis were filtered through 0.2 171 μm syringe filter. Reproducible TOC values with an accuracy of ±1% were determined by injecting 211 μL 172 aliquots to the TOC analyzer. The statistical paired t-tests analysis of experimental results was conducted 173 308 Fig. 3. (a) Schematic. (b) Initial and final concentrations when influent B spiked with E. coli. 309 310 In earlier works, scanning electron microscopy (SEM) of E. coli after electrochlorination using a 311 Ti rod coated with RuO2 and TiO2 as an anode material demonstrated more pronounced cellular damage 312 compared to conventional chlorination and to electrolysis in Na2SO4 solution. After in situ 313 electrochlorination cell wall had more explicit roughness and winkles, indicating the presence of stronger 314 oxidizing species than active chlorine (Diao et al., 2004; Li et al., 2004). In another study, it has been 315 reported that formation of chloride-based radicals, particularly the formation of ·OCl, is plausible during 316 electrochlorination using Ti/RuO2-IrO2 anode (Luna-Trujillo et al., 2020). Thus, the elevated disinfection 317 efficiency during in situ electrochlorination might be potentially attributed to formation of reactive chlorine 318 radicals on the Ti/RuO2 anode surface since chloride anions are present. Nevertheless, synergetic impact 319 from both radicals and chlorine species cannot be excluded. It is important to highlight that the generation 320 of radical species may vary significantly depending on electrocatalytic material characteristics, water 321 composition, and operating parameters. Further fundamental research on radical formation while using 322 anodes with low overpotential for OER are needed to verify their presence and understand contribution to 323 disinfection processes. 324 The influence of the high local concentration of active chlorine species under low pH within the 325 diffusion layer on the anode surface on the microbial inactivation has been overlooked in the literature, 326 while this might be a more contributing factor to enhanced disinfection performance during 327 LogN0 LogNt 0 1 2 3 4 5 6 7 LogNt, (CFU mL-1) LogN0LogNt Influent B spiked with E. coli or MS2 or ≈ 1.43 seconds Influent B SO42SO42SO42- (a) (b) electrochlorination while using Ti/RuO2. It is crucial to note that numerous studies have observed 328 significantly enhanced E. coli inactivation during chemical chlorination under lower pH conditions 329 (Bhandari et al., 2010; Park et al., 2004; Wang L et al., 2007). Furthermore, high local dosage of chlorine 330 near the anode surface can greatly facilitate disinfection since inactivation efficiency is also a function of 331 concentration (Crittenden et al., 2012). Previous studies have demonstrated that diffusion layer during 332 electrochemical oxidation can extend several hundreds of micrometers with acidic pH (Fuladpanjeh‐ 333 Hojaghan et al., 2019; Obata et al., 2020). Hence, we can suggest that the enhanced inactivation observed 334 during in situ electrochlorination might be potentially attributed to the influence of high chlorine 335 concentration under low pH within the diffusion layer at the anode surface. However, the characteristics of 336 diffusion layer also exhibit substantial variations depending on hydraulic parameters inside the reactor and 337 operating current. 338 The E. coli inactivation results revealed the statistically validated superiority of water disinfection 339 when implementing in situ electrochlorination over disinfection in chemically and electrochemically 340 chlorinated waters. No E. coli inactivation was detected when substituting the Clanions with SO42-, 341 indicating absence of contribution to disinfection from ROS standing alone. Therefore, we propose that 342 higher efficiency of E. coli inactivation observed during in situ electrochlorination might be attributed to 343 two potential factors i) formation of reactive chlorine radicals (RCR) on the anode Ti/RuO2 surface and/or 344 synergetic impact from both radical and chlorine species; ii) the influence of the localized high 345 concentration of free chlorine (i.e., Cl2(aq), HOCl) within the diffusion layer on the anode surface under 346 acidic localized pH. 347 348 3.2 MS2 inactivation during in situ electrochlorination 349 Besides enteric bacteria, we employed the MS2 bacteriophage as a surrogate microorganism to 350 simulate viral water contamination and conduct a more descriptive comparison of chemical and 351 electrochemical water disinfection. EPA provides inactivation credit (inactivation efficiency) data 352 applicable for various disinfectants and corresponding dosages and contact time (Ct value), primary for 353 viruses, Cryptosporidium, and Giardia. While Cryptosporidium, and Giardia might achieve several log of 354 removal through pre-treatment techniques such as filtration due to their relatively large sizes, the challenge 355 arises with nano-scale viruses using pore sizes larger than those employed in ultrafiltration (EPA, 2020). 356 Consequently, the evolution of viral inactivation assumes pivotal importance and provides more 357 representative assessment of in situ electrochlorination. 358 The conventional chlorination of MS2 demonstrated an inactivation efficiency of 3.2 log reduction, 359 decreasing from an initial concentration from 6.5 log to 3.3 log within a contact time of 60-second in a 360 storage reservoir (Fig. 4a). The concentration decrease followed a linear slope in the first 30 seconds of 361 chlorination. However, a plateau was observed in concentration change after 30-second time, suggesting 362 the presence of phage aggregations (Ghernaout, 2017). Similar concentration change patterns, including a 363 slope and tailing, were observed during disinfection in electrochlorinated water. The disinfection efficiency 364 of MS2 in electrochlorinated water resulted in 4.3 log inactivation, reducing the initial concentration from 365 6.5 log to 2.3 log within the same contact time (Fig. 4b). Statistical analysis using a paired t-test indicated 366 no statistically significant difference with a p-value of 0.075 in the final log concentrations between these 367 two inactivation efficiencies. The inactivation of MS2 yielded a less pronounced slope in concentration 368 reduction compared to E. coli, confirming relatively higher level of resistance to chlorination process 369 (Huang et al., 2016). These findings are consistent with the results obtained for E. coli, suggesting that apart 370 from active chlorine there were no other disinfecting agents present after electrochlorination that 371 contributed to disinfection. 372 373 Fig. 4. (a) Concentration of MS2 during convention chlorination; (b) concentration of MS2 when spiked in 374 effluent A after electrochloirnation; (d) concentration of MS2 when spiked in influent A before 375 electrochlorination. 376 377 After a single pass of 1.43 seconds through the cell of matrix A and 1 minute of contact time, the 378 inactivation efficiency of MS2 was 5.8 log, resulting in a decrease from 6.2 log initial concentration to 0.4 379 log (Fig. 4c). Unlike E. coli, which exhibited a drastic concentration change of 2 log in a single pass of 1.43 380 seconds, MS2 resulted in 0.5 log reduction (Fig. 4c). Moreover, a similar slope in all three types of MS2 381 disinfection was observed during the first 30 seconds of contact time (Fig. 4), not showing noticeable 382 difference in disinfection kinetics. However, continuous MS2 concentration decrease in storage reservoir 383 was detected after in situ electrochloirnation, leading to statistically significant difference between the final 384 concentrations. The p-value of 0.013 was determined when comparing final log concentrations after in situ 385 electrochlorination (Fig. 4c) with conventional chlorination (Fig. 4a), and 0.003 when comparing it with 386 disinfection in electrochlorinated water (Fig 4b). The higher final inactivation efficiency observed after a 387 single pass and 1 minute of contact time in the storage reservoir might be attributed to more effective 388 damage or inactivation primary of the outer layer of MS2 aggregates. 389 Active chlorine has been found to cause damage to both the genomeand protein-mediated 390 functions within MS2 (Rattanakul and Oguma, 2017; Wigginton et al., 2012). Where viral genome was a 391 0 5 10 15 20 25 30 35 40 45 50 55 60 0 1 2 3 4 5 6 7 LogNt (PFU mL-1) Time (seconds) 0 5 10 15 20 25 30 35 40 45 50 55 60 0 1 2 3 4 5 6 7 LogNt (PFU mL-1) Time (seconds) 0 5 10 15 20 25 30 35 40 45 50 55 60 0 1 2 3 4 5 6 7 LogNt (PFU mL-1) Time (seconds) (a) (b) (c) RA primary target of chlorination process. Conversely, the action of radical species has been suggested to 392 damage the viral capsid (Rattanakul and Oguma, 2017). The continuous linear decrease in MS2 393 concentration (Fig. 4c) might indicate that the bacteriophage capsids in the outer layer of aggregates were 394 preliminary damaged by radical species. This enhanced further penetration of available chlorine, leading to 395 subsequent inactivation. However, it is also not excluded that the damage or complete inactivation was 396 caused by the high concentration of active chlorine under low pH in the vicinity of the anode surface. 397 Despite the tendency of bacteriophage MS2 to form aggregations at pH lower its isoelectric point of 3.9, 398 which might inhibit complete inactivation (Brennecke and Kohn, 2009; Langlet et al., 2007; Overby,’ et 399 al., 1966), several studies have demonstrated more efficient MS2 inactivation by active chlorine at lower 400 pH (Cai et al., 2016; Tang et al., 2021). Therefore, both scenarios are possible and could contribute to more 401 successful inactivation during in situ electrochloirnation, including the formation of reactive chlorine 402 radical species and/or synergetic impact of radicals and chlorine to enhanced disinfection; influence of local 403 elevated chlorine dose under low pH near the anode surface. 404 The statistical analysis indicated that there was no significant difference in terms of initial chlorine 405 dose and residual chlorine concentrations after 60 seconds of contact time (p-value > 0.05) among all three 406 types of treatment. These findings suggest that factors other than active chlorine species at pH 7 might be 407 involved in more efficient disinfection during in situ electrochloirnation. After a single pass and 1 minute 408 of contact time in the storage reservoir (Fig. 5c), an active chlorine dose greater than 1.2 mg L-1 was 409 determined. This remaining dose of residual chlorine ensures continuous water disinfection. 410 411 Fig. 5. Residual chlorine (a) during convention chlorination; (b) in electrochlorinated water when effluent 412 A spiked with MS2; (c) right after (RA) single pass and over time in storage reservoir when influent A 413 spiked with MS2. 414 415 To confirm that ROS are not responsible for increased MS2 inactivation efficiency observed during 416 in situ electrochloirnation, we conducted electrochemical treatment of water without chloride anions present 417 (Table 1). In situ electrochemical treatment of matrix B demonstrated no MS2 inactivation. Statistical 418 analysis, when comparing the initial and final log concentrations of MS2, yielded a p-value of 1, indicating 419 no significant difference. These results shows that chloride content is crucial for achieving water 420 disinfection during in situ electrochloirnation while using Ti/RuO2 as an anode under realistic operating 421 conditions. Nevertheless, such diminutive chloride concentration as 50 mg L-1 in matrix A, which is 5 times 422 lower than maximum guideline value of 250 mg L-1 for drinking water, was still sufficient to provide enough 423 active chlorine for primary and secondary water disinfection. 424 0 5 10 15 20 25 30 35 40 45 50 55 60 0.0 0.5 1.0 1.5 2.0 Residual chlorine (mg [CL2] mL-1) Time (seconds) 0 5 10 15 20 25 30 35 40 45 50 55 60 0.0 0.5 1.0 1.5 2.0 Residual chlorine (mg [Cl2] L-1) Time (seconds) (a) (b) (c) RA 0 5 10 15 20 25 30 35 40 45 50 55 60 0.0 0.5 1.0 1.5 2.0 Residual chlorine (mg [Cl2] mL-1) Time (seconds) 425 Fig. 6. Initial and final MS2 concentrations when influent B spiked with MS2. 426 The superior efficiency of in situ electrochlorination over chemical chlorination and disinfection in 427 electrochlorinated water was statistically confirmed for MS2 inactivation. Furthermore, after undergoing a 428 single pass through the electrochemical flow cell and 1 minute of contact time in the storage reservoir, the 429 MS2 reduction of 5.8 log met the minimum requirements for viral inactivation of 4 log (EPA, 2020). In situ 430 electrochlorination showed 5.8 log of reduction with a Ct value of 1.75 min mg L-1, while suggested Ct 431 value by EPA to achieve 4 log reduction is 3 min mg L-1 at pH range of 6-9 and a temperature of 20 ˚C. 432 These findings highlight the effectiveness of in situ electrochloirnation in viral reduction, even with lower 433 Ct value requirement compared to traditional active chlorine treatment guidelines. The ability to achieve 434 substantial viral reduction with lower Ct values implies potential advantages in terms of treatment 435 efficiency and resources utilization. 436 LogN0 LogNt 0 1 2 3 4 5 6 7 LogNt, (PFU mL-1) LogN0LogNt 3.3 Techno-economic analysis 437 In order to assess economic feasibility of electrochlorination systems for decentralized water 438 disinfection applications, a preliminary techno-economic analysis (TEA) was conducted. The goal of this 439 analysis was to compare two electrochemical chlorination scenarios versus a chemical chlorination baseline 440 treatment. Two electrochemical scenarios involved two different approaches: one utilizing power from the 441 electrical grid as energy source, and the other relying on consumption of locally generated energy from 442 solar panels. The evaluation encompassed all the associated expenses of three scenarios considering the 443 equipment and operational parameters details employed in this study as summarized in Table 3. The 444 economic evaluation is simple, but appropriate for concept feasibility or screening study at early stages of 445 technology development (Class 1 Estimate, AACE International). 446 In order to compare cost efficiencies of the systems over their lifetimes, the levelized cost of water 447 method was used. Levelized costs are calculated as the discounted sum of costs over the assumed lifetime 448 of the device divided by the discounted sum of clean water produced over the device lifetime. The levelized 449 cost of water is measured in cost per unit volume treated, e.g. $ L-1. An equation defining the calculation 450 is shown below: 451 𝐿𝐶𝑂𝑊 ($ 𝐿−1) = ∑𝐶𝑡/(1 + 𝑟)𝑡 ∑𝑄𝑡/(1 + 𝑟)𝑡, (6) where Ct is the total cost including capital and operating costs in year ($), Qt is the treated water production 452 in year t by system of interests (L), and r is the discount rate (% yr-1). 453 Operating expenses refer to the quantity of consumables, such as electricity and chemicals, were 454 calculated based on yearly water treatment. The electric energy consumption by cell was estimated to be 455 0.13 kWh per day and be 0.17 kWh per gram of generated chlorine under used operating parameters (S2 456 and S1). The cost of electricity consumed by electrochemical system when using a centralized electricity 457 source is estimated to be 0.136 $ kWh-1. If an alternative energy source is employed, such as solar energy, 458 the levelized cost of energy (LCOE) is assumed to be 0.05 $ kWh-1 (EIA, 2023; SETO, 2021). For chemical 459 water disinfection baseline scenario, the average prices of gallon of bleach (7.5% NaOCl) and 43 fl oz of 460 muriatic acid were taken as average costs in the United States resulting in $10.00 and $3.19, respectively 461 (S4). 462 Calculations of the material costs required to build the electrochemical systems indicate that the 463 electrodes and cell could require less than $15 of materials (S4). The estimated power consumption was 464 only 5W, and 24V/5W power supplies are readily available as wall adapters for $3-$10 retail. We predict 465 that the costs of creating and delivering the electrochemical system as a retail product will be dominated by 466 sales, distribution, installation, and design costs rather than the material costs which can be calculated rather 467 precisely. Electrochemical chlorination devices are readily available for saltwater pools from retailers for 468 $500-$1500, albeit at higher Cland much higher flow rates employed. Likewise, dosing pumps for 469 chemical disinfection are typically used in highly controlled scientific or manufacturing environments with 470 very high precision and/or flow rates. Small pumps are available in retail for aquariums for roughly $100471 $300. Quotes from dosing pump vendors indicate that small, highly accurate pumps are available for 472 $1000-$2500. Two would be required for the chemical treatment concept including a pump for NaOCl and 473 one for muriatic acid addition to adjust water pH. Doing a detailed cost engineering effort on these concepts 474 is beyond the scope of this effort. Therefore, we have reasoned using the cost data available that both 475 devices for electrochemical treatment and chemical addition could be purchased for $500. In addition, both 476 devices would require one trip of qualified installer which is assumed to result in expenses of $250. Initial 477 equipment and installation costs for both systems are assumed to result in $750 applying to C0. 478 Yearly operating expenses encompass costs related to monitoring and service labor, needed to 479 ensure proper systems functionality. The life of utilized devices including pumps and electromechanical 480 cells are generally considered to have 10–15-year lifespan (S4). Without detailed lifetime performance 481 data, this assumption could be thought of as a target for the electrochemical concept. The discount rate used 482 is taken as a middle value between public water utilities and households since public treatment often 483 employs very low weighted average cost of capital and household tends to have higher discount rates when 484 making energy efficiency purchase decisions. 485 486 Table 3 487 Summary of capital and operational expenses. 488 Electrochemical (Grid) Electrochemical (Solar) Chemical Water Production (L yr-1) 157,680 157,680 157,680 Current Density (mA cm-2) 20 20 N/A Cell Potential (V) 27 27 N/A Chlorine Dose (mg L-1) 1.75 1.75 1.75 Effluent pH 7 7 7 Cost of Electricity ($ kWh-1) 0.136 0.05 N/A Cost of Bleach (7.5% NaOCl) ($ L-1) N/A N/A 2.50 Cost of Muriatic Acid (37%) ($ L-1) N/A N/A 2.6 Equipment Purchase ($) 500 500 500 Equipment Installation ($) 250 250 250 Yearly Operating Expenses ($) 6.43 2.37 0.45 Device Life/ Analysis Period (yr) 10 10 10 Discount Rate (% yr-1) 10 10 10 Levelized Cost ($ m-3) 0.815 0.777 0.789 489 Based on the levelized cost analysis ($ m-3), the utilization in situ electrochlorination system for 490 decentralized water disinfection with alternative energy source resulted in the lowest expenses of $0.78 per 491 liter of the treated water. 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