Ozonation for Low-Load Greywater Treatment: A Review and Experimental Considerations for Small-Scale Systems

Abstract

The effectiveness of ozone (O₃) in eliminating various types of microorganisms, as well as in oxidizing a wide range of contaminants present in wastewater, and drinking water, is extensively documented in the literature, along with the required concentrations, contact times (Ct values), reaction mechanisms for different pollutants, and overall efficiency. This article presents a comprehensive review on the use of aqueous O₃ for treatment and disinfection, specifically for low-contaminant domestic greywater (LGW), providing information for its integration into the design of small-scale treatment systems. Additionally, to complement the theoretical findings, experimental tests were conducted using a portable O₃ generator in an operational facility treating greywater (GW) from handwashing sinks. The results confirmed that O₃ concentration increases over time but decreases as the volume of water to be treated increases. Water analysis results showed significant reductions in BOD₅, turbidity, and total suspended solids after treatment. Furthermore, the results demonstrated that the presence of microorganisms in LGW is minimal, as in the case of fecal coliforms, ensuring a 1 Log disinfection level in this type of system. O₃, as the sole treatment and disinfection system, with an oxidation potential nearly twice that of chlorine, proved to be highly effective in small-scale treatment systems, promoting sustainable practices, water resource conservation, environmental protection, and public health.

1. Introduction

Only 2.5% of the Earth’s total water is freshwater, a resource vital for sustaining life [1]. By 2050, this critical resource is expected to face significant stress due to population growth, urbanization, and climate variability [2,3,4,5,6].

This escalating demand for water and energy has driven the development of sewage systems that capture domestic, industrial, and hospital waste [7]. In this context, improving water management efficiency in sustainable human settlements has become imperative, making the reuse of GW an attractive option [8], especially in areas facing water stress, and especially in rural areas [4,6,9].

Regenerated GW can be used for irrigation, toilet flushing, industrial applications, firefighting, and other purposes, depending on the country and its regulations [1,8,10]. Key factors affecting its quality include population lifestyle, socioeconomic level, family structure, customs, and the type of water supply available [11,12,13].

On the other hand, inadequate management of surface runoff causes direct contamination of underground reservoirs, and river and lakes, potentially reaching drinking water supplies [14]. This scenario, aggravated by climate change, which in recent decades has led to changes in the frequency and intensity of extreme events, highlights the urgency of identifying new water sources [15].

The presence of biological contamination in water resources, such as bacteria, viruses, and protozoa, poses a serious health threat, as it can cause a wide range of waterborne diseases [2,3,16]. Effluents contaminated with industrial, agricultural, and domestic waste severely pollute this resource. The United Nations Development Program (UNDP) reported that 80% of wastewater is discharged into water bodies without adequate treatment, raising concerns for the World Health Organization (WHO), which has detected the direct release of chemical and pharmaceutical contaminants into rivers, for example [7].

In Chile, compounds such as diclofenac, ibuprofen, naproxen, carbamazepine, fluoxetine, caffeine, and sulfamethoxazole have been detected in treatment plants [15]. As such, sanitation and water reuse are core elements of Sustainable Development Goal 6 (SDG), which aims to expand activities and programs related to reuse technologies by 2030 [17].

In high-income countries, approximately 70% of municipal wastewater, a mixture of blackwater and greywater [18], receives treatment, while this proportion decreases to 38% in upper-middle-income countries, 28% in lower-middle-income countries, and 8% in low-income countries [17]. In Chile, sewerage and treatment coverage reached 97.5% and 100%, respectively, for the population connected to the sewer system in urban and concessioned areas, positioning the country as a leader in Latin America [19].

Globally, pioneering countries in GW reuse include the United States (California, Arizona, Texas, and Florida), Australia, Japan, the United Kingdom, France, Germany, Israel, Saudi Arabia, Jordan, Cyprus, and Spain, as noted by Diaz et al. [11]. Moreover, GW represents around 80% of household wastewater, with per capita consumption ranging from 60 to 245 L per day in developed countries, aligning with figures reported in other studies [20]. The reuse of GW for toilet flushing alone could yield water savings of up to 30% for households and up to 60% for office buildings.

The negative environmental impacts associated with the water treatment lifecycle are mainly due to electricity consumption in pumping and distribution systems [5]. Consequently, O₃ has gained popularity as a treatment agent in recent years due to reductions in O₃ production costs [21,22], positioning it as an innovative, fully natural, cost-effective, and environmentally sustainable technology [23].

O₃ is easily generated in situ from oxygen or air [22] using commercially available generators [16] through the application of UV radiation or electricity [24]. In our case, a corona discharge generator fed with ambient air is used for O₃ production, which is then injected into greywater through a bubble diffuser plate for the ozonation process. O₃ is produced when oxygen molecules (O₂) are dissociated by an energy source, generating oxygen atoms that subsequently collide with another oxygen molecule to form an unstable gas with a characteristic blue color and sharp odor [25,26,27,28,29].

Its half-life in distilled water is approximately 20 min (±10 min, depending on environmental conditions) before it reverts to an oxygen molecule [23,30,31,32,33]. By definition, the half-life is the time required to reduce the concentration of a species to half of its initial value [34]. Studies indicate that O₃ solubility increases as water becomes purer or less contaminated, as well as at lower temperatures [30,35,36]. With an oxidation potential of 2.07 V, O₃ is one of the most powerful oxidants, surpassing chlorine with 1.36 V [10,22,28,37,38,39]. In comparison to the reaction rate of molecular O₃, the hydroxyl radical (·OH), with an oxidation potential of 2.80 V [39], is a particularly important oxidant since its assault is 10⁶–10⁹ times faster, and pH affects the ratio of O₃ and ·OH radicals in wastewater [25]. In general, an increase in the concentration of O₃ in the liquid phase leads to an increase in the oxidation rate and a direct selective reaction with the substrate or organic substances. On the other hand, it has been observed that ·OH, generated by the self-decomposition of O₃, can non-selectively attack refractory organic compounds [25,28,34], as shown in Figure 1.

Thus, O₃ emerges as a viable solution for LGW reuse [40]. A detailed economic analysis is not included, as it is well-documented that the return on investment is usually unfavorable due to the high initial and operational costs of treatment systems [41]. The research aims to provide a unique, compact, low-volume, and cost-effective system capable of treating (oxidizing or reacting with organic matter) and disinfecting greywater using O₃. This system can be replicated within the same building or property, enabling the reuse of slightly contaminated water in compliance with regulatory quality standards.

2. Methodology in the Literature Review

To compile the appropriate material, an exhaustive state-of-the-art review was conducted, including academic literature, research studies, theses, journal articles, and relevant books. Searches were performed on Google Scholar and indexed journal databases using specific keywords such as “aqueous ozone”, “ozone water treatment”, “greywater”, “disinfection”, and “reuse”, among other terms. This review establishes a solid foundation to guide the development and analysis of this research.

This article is structured into three clear sections: first, the investigation and collection of information on ozonation systems for water treatment and disinfection, as well as systems for GW treatment; second, an analysis of the mass transfer parameters of O₃; and finally, experimental trials to compare the results of the review and thus verify its implementation in the treatment of LGW. While the reviewed literature provides substantial support for this approach, this study seeks to further validate its feasibility through in situ trials; in this way, it is intended to provide additional background information to support decision making during the design stage.

2.1. Description of O₃ and Its Applications in Water Treatment

Ozone is poorly soluble in water and highly volatile [22,38,42]; it remains in solution for only a few minutes after application (just 10% of its production), eliminates bacteria by rupturing their cell membranes, and has a strong capacity to eliminate viruses [21]. Its solubility is almost 10 times greater than that of pure oxygen [9,43] and is recommended even for highly contaminated water as a treatment in advanced oxidation processes (AOPs) or tertiary treatment [28,38,44,45,46], enabling, in many cases, the reuse of treated wastewater [3,21,35,47,48].

The reaction of ozone and organic pollutants can be divided into two groups: acidic medium favors direct oxidation, and alkaline medium favors indirect oxidation [28]. O₃ degrades organic contaminants both directly (through molecular ozonolysis) and indirectly via free radicals produced during its decomposition [2,14,21,23,46,49,50]. O₃ is commonly applied for the removal of microcontaminants, reduction of chemical oxygen demand (COD), and decolorization [7,9,48]. Several studies have validated the efficiency of the O₃ process as a pretreatment option for wastewater [35,46] and have also employed it for drinking water treatment [51,52].

Other applications include the treatment of industrial wastewater, soil disinfection, medical and dental applications, swimming pools, odor treatment, pesticide degradation, and sludge conditioning, among others [14,30,53,54]. Additionally, one of the most important applications of O₃ in drinking water treatment is the oxidation of arsenic, iron, and manganese [25,49], as well as the degradation of potentially carcinogenic compounds, such as polycyclic aromatic hydrocarbons [53]. O₃ has a disinfectant effect and does not form chloroorganics, which are undesirable by-products in water and potentially toxic to humans [2,34,47,55,56,57,58], providing environmental advantages over chlorine [22]. Aqueous O₃ was introduced by the United States Environmental Protection Agency (USEPA) in 1989 for the design of water ozonation reactors [59].

It is important to note that chlorine treatment does not guarantee the elimination of certain types of pathogenic microorganisms, such as Giardia and Cryptosporidium cysts [34,59]; however, O₃ has been described as an effective inactivator [49]. Moreover, O₃ is a highly attractive disinfection alternative for several reasons: it is a strong germicide against bacteria, viruses, and protozoa [30,43,55]; it reduces the color of poor-quality water and removes compounds that cause taste and odor issues [3,35]; and simultaneously, it oxidizes organic matter, thereby improving wastewater quality [3,47].

As a gas, it is more stable, with a half-life of approximately 12 h in air, enough time to penetrate all areas within a room, including cracks, fixtures, fabrics, and inaccessible surfaces on furniture, making it much more efficient than aerosols, manually applied liquids, and even UV radiation [32,33,57]. However, it can damage some surfaces or materials such as rubber, plastics, fabrics, paint, and metals [43].

O₃ alone does not fully oxidize certain refractory or recalcitrant organic compounds and has a low reaction rate [22]. The combination of O₃ with hydrogen peroxide and UV radiation, known as an AOPs, enhances its efficiency in water treatment [8,14,22,50,54].

Table 1. Comparison of different disinfection methods, modified from [35].

Characteristic	Chlorine	Sodium Hypochlorite	Calcium Hypochlorite	Ozone	UV
Toxicity to microorganisms	High	High	High	High	High
Solubility	Slight	High	High	High	N.A.
Stability	Stable	Slightly stable	Relatively stable	Unstable	Must be generated upon use
Toxicity to higher life forms	High	Toxic	Toxic	Toxic	Toxic
Interaction with foreign matter	Oxidizes organic matter	Active oxidant	Active oxidant	Oxidizes organic matter	Moderate
Corrosion	Highly corrosive	Corrosive	Corrosive	Highly corrosive	N.A.
Deodorizing capacity	High	Moderate	Moderate	High	None

presents the characteristics and compares the most commonly used disinfection methods.

O₃ can be toxic to the lungs and eyes depending on concentration, temperature, humidity, and exposure time [29,30,33]. The Occupational Safety and Health Administration (OSHA) in the United States has established safe air exposure limits of 0.1 ppm for 8 h or 0.3 ppm for 15 min [33,35,43]. In the same country, O₃ has received GRAS (generally recognized as safe) approval for direct contact with food [30,37,48,54]. It can be used as a food disinfectant when applied at levels and through methods compatible with good manufacturing practices, and it is classified by the Food and Drug Administration (FDA) as a secondary direct food additive (processing aid) for food [27,30,37] and as medical equipment for sterilization [25,60]. The olfactory perception threshold of O₃ is 0.04 mg/m³, equivalent to 0.02 ppm, which has no effect on human health [43]. In water, there are no health or skin-related issues [25,61].

The main disadvantages of O₃ include:

• Its ability to corrode certain materials with prolonged exposure.
• Its potential toxicity to humans, which can be mitigated by temporarily isolating people during treatment and sealing properly to prevent gas escape into the environment [31,33].
• The presence of bromide in raw water can lead to the formation of brominated organic by-products during ozonation, which can be carcinogenic at very high levels [49,59]. However, if the pH is below 8, hypobromite ions, which are precursors to bromate ions, can be avoided [39]. For their part, the WHO and the EPA have established a maximum limit of 10 μg/L for bromate in drinking water, and given that LGW comes directly from washbasins, it is highly likely that this condition is met. Moreover, if the concentration of bromide ions is below 20 μg/L, there is no significant risk. However, if it exceeds 50 μg/L, bromate formation must be evaluated [62].
• It is a rather costly process at an industrial scale due to the low solubility of O₃ in water, which increases the need for O₃ supply to treat the contaminant and requires a long contact time [63]. Due to the low transfer efficiency of O₃, the residual gas must be removed when concentration levels exceed the limits established by OSHA. This removal can be achieved through thermal, catalytic, or activated carbon adsorption methods. The thermal method is the easiest to operate, while the latter is not recommended for large-scale plants due to safety concerns and high maintenance requirements [35].

2.2. Factors Affecting the Ozonation Process in Treatment

Temperature, pH, and flow rate with O₃ are three factors that directly and significantly impact the decomposition rate and half-life of O₃ [30,32,34,37,55]. The operating temperature affects the half-life, solubility, self-decomposition of ozone, and the kinetics between O₃ and pollutants. The solubility of ozone decreases with the increase in temperature [28]. O₃ reacts with multiple organic and inorganic compounds in wastewater in a process known as “ozone demand”, which is crucial in the design of ozonation systems, as reacted O₃ is not available for disinfection [24].

In O₃ disinfection, pH may be an important factor, even though the quality of sterilization at a specific concentration of O₃ solution is not considered to significantly alter across pH [25,35,64]. At alkaline pH, reactions follow a radical pathway and generate more ·OH, whereas at acidic pH, they proceed via a selective direct reaction mechanism. Therefore, it is suggested that a higher solution pH enhances ozone decomposition into ·OH, which improves the degradation of refractory organic compounds and leads to greater total organic carbon (TOC) removal [28]. The results reported in the literature show that in natural and synthetic water (deionized water with or without buffer) an increase in the pH-value increases the reaction rate, and for the combined oxidation processes, the effect of pH is even more complex [34]. The optimal pH range is quite broad, between 5.5 and 9.5, and temperatures between 5 °C and 35 °C do not significantly affect its efficiency [58]. Studies have shown that elevated pH values (close to 10) maximize the reduction of contaminants [9], as in the removal of certain pesticides, where 65% degradation was achieved at pH 9 and an HRT of 60 min [14].

It has been shown that O₃ pretreatment converts recalcitrant compounds into biodegradable organic compounds, enabling the transformation of high-molecular-weight compounds into smaller, more polar, oxygenated, and non-toxic ones [23,35]. Examples of these types of wastewaters include those originating from wineries, textile plants, or the paper industry. O₃ is effective in treating such waters because it reacts indirectly through the formation of hydroxyl radicals, which are highly unstable and immediately react with other molecules to recover the missing electron. The reaction between O₃ and hydroxide ions leads to the formation of a superoxide anion and a hydroperoxyl radical. The anion reacts with O₃ to form the ozonide anion, which rapidly decomposes into an ·OH radical through the formation of hydrogen trioxide. The hydroxyl radical then reacts with O₃ to form a tetraoxohydroxyl radical (HO₄·), which decomposes into a hydroperoxyl radical upon reacting with O₂. The hydroxyl radical also reacts with the organic molecule R, which can act as a promoter, forming either an organic radical (R·) or peroxyorganic radicals (ROO·) if O₂ is present [46].

However, some studies indicate that pathogen inactivation increases at pH 6 [39], and disinfection is more efficient at a slightly acidic pH or within a range of 6 to 9 for various microorganisms [37]. An aqueous O₃ concentration of 2 to 4 ppm (1 ppm = 1 mg/L in water) was sufficient to reduce manure-based pathogens (MBP) below detectable limits in 2 min on the aforementioned smooth surfaces. The same publication mentions minimum concentrations of 1 ppm in water and 20 to 30 ppm for gaseous O₃ for bacteria [32].

The intake flow rate directly affects the liquid phase state. Increasing the intake flow rate accelerates the update of the gas–liquid interface and decreases the mass transfer resistance of ozone; however, a higher intake flow rate does not indicate better results, as the excessive intake flow rate will cause huge turbulence, resulting in a decrease in the dissolving efficiency of ozone in the liquid phase [28].

2.3. Considerations for the Design of GW Treatment Systems

As urbanization and industrialization increase, so does the demand for water. Domestic wastewater contains significant amounts of nutrients that can be reused for irrigation, reducing the demand for freshwater in agriculture [10,21,65]. With the depletion of natural water reserves, reclaimed water, also referred to as “wastewater treatment” (WWT) is considered a sustainable source and alternative due to its continuous production, which is relatively independent of climatic conditions [3,4].

Domestic wastewater (DWW) includes millions of intestinal bacteria and other organisms that pose health risks [3,66]. Treatment systems for this type of water range from biological processes to simple physical treatments combined with disinfection, where biological processes achieve effective removal of organic matter [67,68]. However, these systems are often complex to operate and require trained personnel, making them impractical for household-scale or small-building applications [68].

The lower degree of fecal contamination and high availability make GW an attractive and easy-to-treat water resource [41,67,68,69] compared to other wastewater categories, such as “yellow” (containing urine), “brown” (containing feces), “black” (a mixture of urine, feces, and bacterial activity), and “grey” (originating from kitchen, laundry, shower, and bathroom sinks) [3,20,65,66,70]. GW is generally classified into low, medium, and high-load greywater (HGW) [11,12], with HGW being the most contaminated especially when it comes from laundry and kitchen sources, which are rich in detergents, oils, and recalcitrant organic pollutants [12]. In contrast, LGW excludes kitchen or laundry effluents, has low levels of organic matter, and exhibits lower biochemical oxygen demand (BOD) and COD [11,12,41,65,68], which, according to the literature, should not exceed 290 mg/L of BOD₅, 330 mg/L for TSS, and 200 NTU for turbidity [11,13]. Based on this, it is considered necessary to conduct tests with HGW to validate the literature findings and determine its regeneration potential.

The main components of GW include organic materials such as proteins, carbohydrates, and fats, in proportions of approximately 50%, 40%, and 10%, respectively [9], although this varies depending on the GW source. The separation of blackwater and greywater, with in situ treatment of GW, is a promising strategy for reuse in toilet flushing and garden irrigation [11,18]. The generation of GW is directly related to human activity, which allows the water supply to be matched to demand [20]. In a typical dwelling, GW accounts for between 50% and 80% of the wastewater generated daily [9,18,41,66], characterized with low organic resistance and high volume [20]. Díaz et al. [11] reported that between 77% and 88% of domestic wastewater corresponds to GW, with the main contributors being the toilet and shower, accounting for 21% and 35%, respectively, figures that are consistent with other studies [66,68]. According to the European Commission, promoting GW reuse and rainwater harvesting could reduce potable water usage by 5% by 2050 [18]. Studies have reported significant savings in potable water consumption through GW reuse for toilet flushing: a university building in Brazil achieved a 25.73% reduction, one in Peru achieved a 12.67% reduction [6], and in schools in northern Chile, savings reached 26% [41].

In Chile, the reuse of GW must comply with Law No. 21.075 [70], updated in November 2023, which defines its approved uses for urban, recreational, ornamental, industrial, and environmental purposes, with emphasis on garden irrigation and toilet flushing in urban applications [11]. In the same year, the new regulation for this law [71] establishes the sanitary conditions for reuse systems aligned with international guidelines [65], thus enabling the possibility of implementing more flexible treatment alternatives.

The efficiency of O₃ also depends on pH, as mentioned earlier: at acidic pH, molecular O₃ directly attacks organic compounds, while at alkaline pH, the process occurs through the formation of hydroxyl radicals [46,72]. It is essential to consider this for proper treatment and to achieve optimal results. O₃ reacts rapidly with compounds such as nitrites, sulfides, and ammonium [24,48], although turbidity can reduce its effectiveness in the inactivation of certain microorganisms [49]. One study reports ammonium reductions from 28 mg/L to 8% at pH 7.0 in 30 min, reaching 42% at pH 8.4 and up to 70% at pH 9.0 [24].

For water with low COD (20–30 mg/L), O₃ concentrations between 0.5 and 1.0 mg/L are required, which can be considered low-load water. Other authors report reductions of up to 88% in COD, 68% in BOD₅, 75% in TSS, and 89% inactivation of fecal coliforms in water treated for irrigation purposes [48]. A study conducted in Jordan achieved removal efficiencies of 15% for TOC and 7.5% for BOD₅ [8]. At high O₃ concentrations, another study in Baghdad, Iraq, reported reductions of 35% to 60% for COD, 21% to 71% for BOD₅, 33% to 54% for TOC, and 16% to 72% for oils and fats at pH 8 with a contact time of 30 min [9]. Ozonation is also known to be effective for the removal of heavy metals; enhances the biodegradation of surfactants (detergents), pharmaceuticals, and personal care products; and eliminates color, COD, and TOC [21,22]. Longer contact or HRT help degrade up to 90% of COD, BOD₅, and TOC within 150 min [9].

Ozonation generates a wide range of carbonyl disinfection by-products, including carboxylic acids, aldehydes, ketones, and aldo-keto acids. However, these compounds are readily oxidized and converted into biodegradable intermediates, particularly through the action of hydroxyl radicals. In LGW, the formation of such by-products is expected to be minimal due to the low availability of organic matter and the lower O₃ dosages applied. The most common water quality parameters are presented in

Table 2. Quality parameters for residential use in toilets, modified from [11].

Toilet Flushing	[73] Hong Kong	[13] Spain	[13] R.D. 1620/2007 Spain	UK [74]	Ministry of Health Canada [67]	[75]	[76] China	[76]	Effluent Quality [70]
Suspended Solids (mg/L)	≤5	-	≤20	-	≤20	-	-	-	≤10
Turbidity (NTU)	≤5	≤2	≤10	<10	<5	≤2	<5	<5	≤5
BOD5 (mg/L)	≤10	-			≤20	≤10	≤10	≤10	≤10
Escherichia coli (UFC/100 mL)	No	No	No	10	<200	No	3	≤10
Fecal Coliforms (NMP/100 mL)	-	-	-	-	-	-	-	-	≤10
Free Residual Chlorine (mg/L)	≥0.2	0.5–2.0	-	<2	0.5–1.5	≤1	>1	≤1	0.5–2.0

and

Table 3. Quality parameters for residential use in irrigation, modified from [11].

Garden Irrigation	[77] Spain	[13] Spain	[13] R.D. 1620/2007 Spain	NSF350 [78]	UK [74]	[75]	[76] China	[76]	Effluent Quality [70]
Suspended Solids (mg/L)	<25	-	≤20	≤30	-	-	-	≤30	≤30
Turbidity (NTU)	<5	<10	≤10	<5	<10	≤2	<20	<5	≤10
BOD5 (mg/L)	-	-		≤25	-	≤30	≤20	≤30	≤30
Escherichia coli (UFC/100 mL)	<200	<200	≤200	≤200	10	≤200	3	≤10
Fecal Coliforms (NMP/100mL)	-	-	-	-	-	-	-	-	≤200

, based on findings from national and international studies and regulations [73,74,75,76,77,78].

2.4. Relevance of Bubble Size in the Reactor

An important parameter in the effectiveness of O₃ in water treatment is bubble size, which also affects its solubility. Smaller bubbles, such as nanobubbles, rise more slowly compared to larger bubbles, increasing the residence time of O₃ in water and improving the specific contact surface, efficiently destroying pathogens [10,22,24,30,46]. Smaller bubbles are beneficial to the mass transfer of ozone, thereby accelerating the removal of pollutants [28], and with a higher surface-to-volume ratio than conventional bubbles, they improve ozone dissolution [51]; for instance, 13 trillion nanobubbles can fit in 1 coarse bubble [23]. Transfer efficiencies close to 90% have been reported when using fine bubbles [72,79]. Studies have compared the lifespan of bubbles produced by a conventional diffuser versus a nanodiffuser, showing that after 1 h of stabilization, nanobubbles retained O₃ in water almost four times longer [80]. Researchers indicate that the concentration delivered by nanobubbles is much higher than that of a regular diffuser, also slowing down its reduction or degradation in water [81], as they do not follow buoyancy and instead move in water through Brownian motion [23].

Microbubbles, which vary in size between 1 and 50 microns, achieve volumetric mass transfer values (kLa) higher than those of conventional bubbles [44]. They also generate hydroxyl radicals in higher concentrations and increase zeta potential in ozonation processes [72]. Tests with micro-nano bubbles of air and O₃ report reductions of 90.2% in BOD and 92.5% in COD in hospital wastewater [46]. Other studies report COD reduction efficiencies close to 50% [50]. O₃ nanobubbles facilitate the removal of contaminants through direct reactions with O₃ molecules or free radicals, as well as enhancing the flotation of oils and fats [80]. O₃ micro-nano bubble technology (MNBT) has been used for algae treatment over the past five years, achieving 100% efficiency in their removal [1,23].

In India, ozone nanobubbles (NBO) smaller than 200 nm were used for cleaning artificial water ponds, achieving reductions in ammonia, total suspended solids (TSS), turbidity, BOD, and COD, with removal values ranging from 55% to 82% for COD, 63% to 91% for BOD, and 83% to 99% for TSS [23]. In Figure 2, the difference can be observed.

This is due to the fact that small-sized O₃ bubbles effectively generate hydroxyl radicals, which are highly efficient in decomposing organic molecules in both acidic and alkaline water environments [46]. These solutions are environmentally sustainable and cost-effective, as they do not introduce chemicals into the water and eliminate many taste and odor issues in the treated water [23].

2.5. Microorganism Inactivation with O₃

There is a recognized relationship between the morphological characteristics of microorganisms and their resistance to O₃, following this descending order of resistance: fungi > spore-forming bacteria > non-spore-forming bacteria > viruses [34,35].

Unlike drinking water treatment processes, where efficacy depends more on O₃ concentration than on contact time, in wastewater, the initial transferred dose is crucial [21]. It has been demonstrated that after 30 s of exposure, 99% of viruses show damage to their envelope proteins, with enveloped viruses being the most susceptible to O₃ action [43]. The microbicidal action of O₃ is a complex process, as it attacks numerous cellular constituents [30]. Studies indicate the removal of organic micropollutants exceeding 85%, while bacterial counts (total heterotrophs at 37 and 22 °C, E. coli, and enterococci) decreased to values below the limits allowed for drinking water, even after 3 days of storage, for concentrations of 3 mg/L and a hydraulic retention time (HRT) of 10 min [64].

Detecting all microorganisms present in a water sample is complex; however, fecal indicators such as Escherichia coli offer a faster, simpler, and more cost-effective alternative for assessing contamination, as their presence indicates recent contamination [2]. The detention time of wastewater is another factor influencing the performance of water disinfection and treatment. It has been demonstrated that contact times, measured in seconds, are sufficient for the inactivation of E. coli and only minutes for coliforms, achieving at least a 1 Log reduction for the latter [24].

Viruses are small particles composed of macromolecules that, unlike bacteria, require a host cell to replicate. Unable to repair oxidative damage, they are more vulnerable to the action of O₃ [43]. Viruses (from the Latin venenum, meaning poison) are entities smaller than bacteria, composed of genetic material surrounded by a protective envelope, with a size ranging from 0.01 to 0.1 mm. They lack reproductive systems and independent life but can replicate within living cells. If it is an RNA virus, it first converts its RNA into DNA, using the machinery of the host cell [43,63]. Among them, coronaviruses (CoVs) are large spherical RNA viruses with spike-like projections that form a crown. They can cause diseases in both animals and humans [33,82,83]. This crown or lipid envelope surrounding the virus, which gives the family its name, makes the virus sensitive to external factors and agents, such as heat, lipid solvents, oxidants such as O₃, and UV radiation [82].

Seven types of human coronaviruses have been identified, four of which cause the common cold, while the other three (SARS-CoV-1, MERS-CoV, and SARS-CoV-2) are more pathogenic, with the latter responsible for the COVID-19 pandemic [33,61,83]. Compared to other pathogens, SARS-CoV-2 has a remarkably high transmission capacity [83,84], similar to the Ebola virus. The coronavirus structure contains cysteine-rich regions and a tryptophan crown, both of which are vulnerable to oxidation by O₃ [29,83,85]. Both SARS viruses exhibit relatively long viability on stainless steel and polypropylene compared to copper or cardboard [86]. Other studies indicate that O₃ attacks glycoproteins and glycolipids in bacterial cell membranes, causing their rupture and inactivation [3,21,29,32,53,55,82]. Carbohydrates and fatty acids react slowly with O₃, whereas amines, amino acids, nucleic acids, and proteins react more rapidly [38,57]. However, the requirement for humidity to achieve optimal efficacy suggests that ions and hydroxyl radicals derived from O₃ decomposition may be involved [87], as observed in aqueous environments [31,88].

The antimicrobial activity of aqueous O₃ against bacterial pathogens, including E. coli, Pseudomonas aeruginosa, and Vibrio spp., has been studied to ensure food safety [35,89]. It has been reported that at concentrations of 1 mg/L, O₃ inactivates the hepatitis A virus (HAV) in 60 s, consistent with similar findings [48,60]. Studies of direct ozonation at the source have shown disinfection levels comparable to those achieved through conventional activated sludge processes combined with additional O₃ treatment for antimicrobial-resistant bacteria [7,28,31,39].

In wastewater with high organic content, E. coli inactivation has been reported with a dose of 10 mg/L and a 5-min contact time [49]. This is confirmed by a study where virus reduction rates after 5 s at O₃ concentrations of 1, 4, and 7 mg/L were 81.4%, 93.2%, and 96.6%, respectively [61].

However, a laboratory-scale study added 8 Log/100 mL of pathogens, including E. coli, Salmonella enterica, Pseudomonas aeruginosa, and the MS2 bacteriophage, to evaluate the effectiveness of O₃ disinfection at a concentration of 5 mg/L. The results showed limited inactivation due to the high organic content of the graywater and the size of the ozone generator. Additionally, the same study indicates that poliovirus is more resistant to O₃ than MS2 [68]. It is clear that the effectiveness in eliminating microorganisms depends on the achieved concentration, equipment size, and contaminant load.

2.6. Criterion in Inactivation with O₃ Ct

The first inactivation criteria emerged in the 1970s, when the WHO adopted the poliovirus inactivation criterion as the basis for O₃ disinfection, establishing that a residual concentration (C) of 0.4 mg/L maintained for a contact time (t) of 4 min is sufficient to achieve more than 99.99% inactivation of this proposed baseline virus. This design criterion, known as “viricidal”, was the starting point for multiple studies [39], some of which indicate a 99.999% or 5 log inactivation of E. coli after 10 min of contact [2].

The USEPA, in a 2003 proposal regarding “new” microorganisms to be considered in drinking water quality, such as enteroviruses, Legionella, and Giardia cysts [47], fully introduced the concept Ct (product of dissolved O₃ concentration C in mg/L and hydraulic residence time t in water), which has since as a regulatory guideline [34,37,39,90] and become a common concept in disinfection processes. In chemical disinfection, the Ct concept is frequently applied, based on the Chick (1908) and Watson (1908) law [2,24,49,87,91], although its application should be adapted to the type of water being treated [34]. Lastly, research has shown that parasites require a much higher Ct value than viruses to achieve a two- or three-log reduction [34]. Studies have reported that enveloped viruses, such as feline calicivirus (FCV) and human parainfluenza virus type 2 (hPIV2), achieved reductions of more than 3 log following exposure to 1 ppm O₃ concentrations for 5 s and 1 min, respectively, resulting in undetectable levels [92]. In another study, FCV in water with organic load showed complete inactivation after 36 min of exposure to 1 ppm O₃ [89]. Other authors report a 99.9% Ct for enteric adenoviruses, ranging from 0.04 to 0.10 mg·min/L [91].

Table 4. Ct Values (mg·min/L) for 2-log inactivation [93].

		Ct (mg/min/L)
pH	6–7	8–9	6–7	6–7
Microorganism	Free Chlorine	Chloramine	Chlorine Dioxide	Ozone
E.coli	0.034–0.05	95–180	0.4–0.75	0.02
Poliovirus	1.1–2.5	768–3740	0.2–6.7	0.1–0.2
Rotavirus	0.01–0.05	3800–6500	0.2–2.1	0.006–0.06
Giardia lamblia (cysts)	47–150	2200	26	0.5–0.6
Giardia muris (cysts)	30–630	1400	7.2–18.5	1.8–2.0
Cryptosporidium parvum	7200	7200 *	78 *	5–10 *
Cryptosporidium parvum (1 °C)			200	10
Cryptosporidium parvum (22 °C)			120 **	7 **

Notes: Temperature: 25 °C; * 1 log; ** 3.5 log.

illustrates the significant advantage of O₃ over other disinfectants.

Additionally, inactivations greater than 1.5 log have been reported for 0.15 ppm O₃ over 2 min for poliovirus 1 and 2, echovirus 1 and 5, coxsackievirus A9 and B5 [55], while for murine norovirus (MNV), 99% inactivation was achieved with O₃ at 0.72 mg/L [55,91]. Reductions of 1.43 log have been reported for poliovirus with a Ct of 0.2 mg·min/L [49].

Table 5. Ct Values (mg·min/L) for virus inactivation by ozone, pH 6–9 [93].

Log Inactivation	Temperature, °C
	≤1	5	10	15	20
2.0	0.90	0.60	0.50	0.30	0.25
3.0	1.40	0.90	0.80	0.50	0.40
4.0	1.80	1.20	1.00	0.60	0.50

shows the Ct values recommended by the USEPA for viral inactivation using O₃ [93], highlighting the strong influence of water temperature on the required Ct value.

The USEPA has evaluated virus inactivation in pilot plants, achieving more than 3 log inactivation of poliovirus 2 and coxsackievirus with a contact time of 5 min and O₃ concentrations of 0.8 mg/L and 1.7 mg/L, respectively [63].

Other studies in hospital wastewater have demonstrated the elimination of more than 99.9% of antibiotic-resistant bacteria after 10 min of treatment with a Ct value of 1.0 mg·min/L [7].

Table 6. Ozone dosage for various waterborne viruses, modified from [35].

Application	Theoretical Ozone Dose (ppm)
Coxsackie Virus	0.51
Poliovirus	0.012–0.015
Porcine Pigina Virus	0.024

and

Table 7. Disinfection of secondary effluent under different ozonation conditions [47].

		Mesophilic Aerobes UFC/mL	Total Coliforms NMP/100 mL	Fecal Coliforms NMP/100 mL
Before Ozonation		4.6 × 102	7.6 × 103	5.2 × 103
After Ozonation (t = 5 min)
Dose	Ct
(mg/L)	(mg·min/L)
7	5	2.5 × 101	negative	negative
14	10	2.2 × 101	negative	negative
21	14	1.5 × 101	negative	negative

show the efficiency of pathogen inactivation under experimental disinfection conditions, indicating that a Ct value of 5 mg·min/L is sufficient to inactivate the analyzed population in secondary wastewater treatment [47].

2.7. Mass Transfer Parameters

The mass transfer of O₃ in water depends on hydrodynamic and physicochemical factors [22], such as the concentration of O₃ in the gas phase, gas flow rate, temperature, and pH of the liquid phase [38], as well as gas pressure and water pressure [79]. When a gas phase rich in compound A comes into contact with a liquid phase poor in A, compound A spontaneously diffuses across the gas-liquid interface. The mass flow transferred can be modeled using Fick’s law of molecular diffusion, assuming that the resistance to mass transfer is located in a thin layer near the interface in the liquid phase [42]. A commonly used theory to quantify this transfer is the Lewis-Whitman double-film theory [28,42,52,87,94]. The solubility of O₃ in water is fundamental to its disinfection capacity, as it depends on the amount of O₃ transferred. The amount of O₃ that will dissolve in a given volume of water at a constant temperature is proportional to the partial pressure of O₃ above the water [24,26]. The equilibrium at the gas-liquid interface can be expressed by the following equation, known as Henry’s law of solubility or its inverse constant [24].

$${\mathrm{H}}^{\mathrm{c}\mathrm{p}}\stackrel{\mathrm{def}}{=}{\mathrm{c}}_{\mathrm{a}}/\mathrm{p}$$

(1)

In this expression, c_a is the concentration of a species in the aqueous phase in mol/m³, and $\mathrm{p}$ is the partial pressure of that species in the gas phase under equilibrium conditions, measured in Pascals (Pa). In the International System, H^cp has units of mol·m⁻³·Pa⁻¹ [24,95]. Another important definition, known as the inverse of H^cp, is Henry’s Law constant or Henry’s volatility [34,96], which has a value of 9090 Pa·m³/mol at 25 °C with a van’t Hoff equilibrium constant [94] equal to 2300 K, or 1.1·10⁻⁴ mol·m⁻³·Pa⁻¹ for its inverse H^cp [87,96].

Henry’s Law constant is represented by the following equation [24,28,96]:

$${\mathrm{K}}_{\mathrm{H}}^{\mathrm{c}\mathrm{p}}\stackrel{\mathrm{def}}{=}\mathrm{p}/{\mathrm{c}}_{\mathrm{a}}=1/{\mathrm{H}}^{\mathrm{c}\mathrm{p}}$$

(2)

In the SI system, the unit for ${\mathrm{K}}_{\mathrm{H}}^{\mathrm{c}\mathrm{p}}$ is Pa·m³/mol. Alternatively, Henry’s solubility can also be expressed as the dimensionless ratio between the concentration in the aqueous phase c_a and its concentration in the gas phase c_g, commonly referred to as the dimensionless constant or “Air-Water Partition Coefficient” H^cc [96,97].

$${\mathrm{H}}^{\mathrm{c}\mathrm{c}}\stackrel{\mathrm{def}}{=}{\mathrm{c}}_{\mathrm{a}}/{\mathrm{c}}_{\mathrm{g}}$$

(3)

It can also be expressed in a dimensionless form as:

$${\mathrm{K}}_{\mathrm{H}}^{\mathrm{c}\mathrm{c}}\stackrel{\mathrm{def}}{=}{\mathrm{c}}_{\mathrm{g}}/{\mathrm{c}}_{\mathrm{a}}=1/{\mathrm{H}}^{\mathrm{c}\mathrm{c}}$$

(4)

The effect of temperature on O₃ solubility is the most significant, as H^cp values are plotted against temperature on a semilogarithmic graph. Henry’s constant increases exponentially with temperature [94,96] and is expressed as:

$${\mathrm{H}}^{\mathrm{c}\mathrm{p}}\left(\mathrm{T}\right)={\mathrm{H}}^{\mathrm{c}\mathrm{p}}\times \mathrm{e}\mathrm{x}\mathrm{p}\left({\displaystyle \frac{-{∆}_{\mathrm{s}\mathrm{o}\mathrm{l}}{\mathrm{H}}^{\mathrm{c}\mathrm{p}}}{\mathrm{R}}}\left({\displaystyle \frac{1}{\mathrm{T}}}-{\displaystyle \frac{1}{{\mathrm{T}}^{\mathrm{c}\mathrm{p}}}}\right)\right)$$

(5)

The temperature dependence of equilibrium constants can generally be described by the van’t Hoff equation [94,95] mentioned above.

$${\displaystyle \frac{\mathrm{d}\mathrm{L}\mathrm{n}{\mathrm{H}}^{\mathrm{c}\mathrm{p}}}{\mathrm{d}(1/\mathrm{T})}}=\left({\displaystyle \frac{-{∆}_{\mathrm{s}\mathrm{o}\mathrm{l}}{\mathrm{H}}^{\mathrm{c}\mathrm{p}}}{\mathrm{R}}}\right)$$

(6)

Expressions regarding the concentrations at the interface are difficult to determine. The equations for the gas phase and liquid phase, respectively, are:

$$\overline{{\mathrm{N}}_{{\mathrm{O}}_3}}={\mathrm{K}}_{\mathrm{g}}\left({\mathrm{P}}_{{\mathrm{O}}_3}-{\mathrm{P}}_{{\mathrm{O}}_3}^{\mathrm{*}}\right)\mathrm{S}$$

(7)

$$\overline{{\mathrm{N}}_{{\mathrm{O}}_3}}={\mathrm{K}}_{\mathrm{L}}\left({\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}-{\mathrm{C}}_{\mathrm{L}}\right)\mathrm{S}$$

(8)

where the molar flux of transferred O₃ N_O3 is in mol/s, and S represents the gas-liquid exchange surface area, K_g and K_L are the overall transfer coefficients in the gas phase (mol·m⁻²·Pa⁻¹·s⁻¹) and in the liquid phase (m·s⁻¹), respectively, ${\mathrm{P}}_{{\mathrm{O}}_3}^{\mathrm{*}}$ is the partial pressure of O₃ in the gas phase at equilibrium, and ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}$ is the concentration in the liquid phase at equilibrium [28,52,98,99]. ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}$ is also known as the theoretical saturation concentration of O₃ in the liquid phase [94] and represents the maximum concentration that can be supplied to the fluid. For compounds with high Henry’s Law constants or low solubility, such as O₃, nitrogen, and oxygen, the transfer resistance is located on the liquid film side, and the gas-side resistance can be neglected [38,98]. This theoretical expression for the transfer coefficient, also called the liquid film rate constant, is represented by k_L in (m·s⁻¹) [34]. Thus, the equation becomes as (8). The determination of the exchange surface S is complex; therefore, the interfacial area per unit volume, a = S/V, is generally introduced into the transfer equation. This yields the final expression for the average transfer flux, applicable to mass transfer when a chemical reaction is not considered. It should be noted that these equations are based on an idealized approximation; however, significant deviations are not expected, as the levels of suspended solids and organic load are low.

$$\overline{{\mathrm{N}}_{{\mathrm{O}}_3}}={\mathrm{k}}_{\mathrm{L}}\mathrm{a}\mathrm{V}\left({\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}-{\mathrm{C}}_{\mathrm{L}}\right),$$

(9)

where:

• k_La: Volumetric mass transfer coefficient (s⁻¹).
• V ≈ VL: Reactor volume and volume of liquid in the reactor (m³).

The k_La coefficient follows a phase that is typically adjusted to first-order kinetics, with values ranging from 10⁻⁴ s⁻¹ for groundwater to 0.01 s⁻¹ for treated wastewater [87]. A similar observation is made by [34], with moderate k_La values in the range of 0.005–0.01 s⁻¹ in simple bubble columns. The interfacial area of the exchange volume, “a” is also considered a hydrodynamic parameter, expressed in units of m²·m⁻³.

Its value reflects the exchange surface area per cubic meter of liquid [93,98]. The area ratio is given as:

$$\mathrm{a}=6/d_{\mathrm{b}}$$

(10)

The denominator d_b corresponds to the bubble diameter. Under certain thermodynamic conditions, it is possible to experimentally determine the values of k_La and the interfacial area “a” and consequently obtain the value of k_L. Many experimental results are expressed as k_La due to the lack of information on “a”. The parameter k_La encompasses effects such as bubble size and high gas residence times in the reactor [28,87].

2.8. Reactor Configuration

The most common laboratory approach is to use a batch configuration [28,34], where O₃ free (clean) water is gasified with an ozone/air or ozone/oxygen mixture. This bubbling is maintained under partial pressure and isothermal conditions [99]. The change in liquid O₃ concentration over time is measured with an O₃ probe or a photometer [34]. For the case study, the configuration is more of a semi-batch, considering that effluent is received during the day, but there is no inflow at night, which is the intended period for bubbling. The rate of decrease is commonly represented by pseudo first-order kinetics [63,88,94]. A balance on a fluid volume element where mass transfer occurs from the gas phase and dissolved O₃ decomposition leads to the following equation [94,100]:

$${\displaystyle \frac{{\mathrm{d}\mathrm{C}}_{\mathrm{L}}}{\mathrm{d}\mathrm{t}}}={\mathrm{k}}_{\mathrm{L}}\mathrm{a}\left({\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}-{\mathrm{C}}_{\mathrm{L}}\right)-{\mathrm{k}}_{\mathrm{d}}{\mathrm{C}}_{\mathrm{L}}^{\mathrm{m}}$$

(11)

where k_d is the self-decomposition coefficient of O₃ (s⁻¹); and m is equal to 1, assuming first-order O₃ decomposition in the liquid phase.

To determine k_d, the stationary regime in the evolution of O₃ transfer from gas to liquid is considered [97,99], yielding the first-order O₃ decomposition kinetics with the following equation [94,99]:

$${\displaystyle \frac{{\mathrm{d}\mathrm{C}}_{\mathrm{L}}}{\mathrm{d}\mathrm{t}}}={-\mathrm{k}}_{\mathrm{d}}{\mathrm{C}}_{\mathrm{L}}^{\mathrm{m}}$$

(12)

By integrating for a first-order reaction (m = 1), a common value in ozonation within bubble columns [44], a value of m = 2 could make sense under high O₃ concentration conditions, which is unlikely in this treatment scenario. The rate coefficients −(k_La + k_d) represent the slope in the linear relationship between ln(${\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}$ − ${\mathrm{C}}_{\mathrm{L}}$) and time t [34,72,94]. If the decomposition term is not considered in the O₃ mass balance, assuming a first-order reaction, the O₃ concentration at steady state can be calculated according to the following equation:

$$\mathrm{l}\mathrm{n}\left({\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}-{\mathrm{C}}_{\mathrm{L}}\right)=-{\mathrm{k}}_{\mathrm{L}}\mathrm{a}\times \mathrm{t}+\mathrm{C}$$

(13)

Generally, the value of k_d is less than 0.001 s⁻¹ for pH levels below 8 [94,100], and if pH increases, it is due to the increased decomposition rate [97]. The k_La values are at least 40 times higher than k_d values; for drinking water, it is considered a slow reaction, with k_d negligible and insignificant within the studied pH range [34,94,100]. The following equation is assumed [38,44,72]:

$${\displaystyle \frac{{\mathrm{d}\mathrm{C}}_{\mathrm{L}}}{\mathrm{d}\mathrm{t}}}={\mathrm{k}}_{\mathrm{L}}\mathrm{a}\left({\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}-{\mathrm{C}}_{\mathrm{L}}\right)$$

(14)

The steady-state or equilibrium concentration in the liquid phase is known as ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{\infty }}$. This is related to the theoretical saturation concentration ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}$, which can be experimentally determined if k_La and k_d are known [34]. The relationship can be expressed as follows:

$${\displaystyle \frac{{\mathrm{C}}_{\mathrm{L}}^{\mathrm{\infty }}}{\mathrm{d}\mathrm{t}}}={\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}{\displaystyle \frac{{\mathrm{k}}_{\mathrm{L}}\mathrm{a}}{{\mathrm{k}}_{\mathrm{L}}\mathrm{a}+{\mathrm{k}}_{\mathrm{d}}}}$$

(15)

It is important to note that, for O₃, the steady-state concentration in the liquid, ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{\infty }}$, even in clean water, is generally not equal to ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}$ due to O₃ decomposition. Since k_d is negligible depending on pH, we can say that the equilibrium concentration is ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{\infty }}$ = ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}$. To consider this, a rapid mass transfer must be ensured so that the rate of O₃ decomposition is negligible compared to the rate at which O₃ is absorbed [93]. Over a sufficiently long time, ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{\infty }}$ assumes the specific value ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}$ [34,94,99]. With system monitoring during O₃ mass transfer into the liquid, and once steady state is reached, ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{\infty }}$ = ${\mathrm{C}}_{\mathrm{L}}^{\mathrm{*}}$= ${\mathrm{c}}_{\mathrm{a}}^{\mathrm{*}}$. This is the goal in each ozonation trial, as shown in Figure 3.

3. Materials and Methods

After the literature review and background collection, tests are conducted to validate an appropriate treatment system as previously described and to compare results with the research findings. This approach helps establish a foundation for future designs.

3.1. System Configuration

The system is designed for the treatment and disinfection of LGW (limited to handwashing sink effluents), as it represents the most feasible scenario for achieving pollutant reduction in compliance with current regulatory standards and is easy to maintain and operate. This approach avoids the high costs associated with robust treatment systems when it is not truly necessary [68]. The system diagram is shown in Figure 4.

Initially, GW from a treatment system installed at a university in Santiago was analyzed. This system captures water from handwashing sinks in men’s restrooms located near the dining area. The water was stored in a 100 L tank to equalize the influent flow and then pumped into a 200 L tank, where ozonation tests were conducted.

Continuing with the evaluations, and to validate the decision to use this system exclusively for treating LGW in small-scale systems, as well as to provide tangible evidence supporting the exclusion of O₃ use in HGW, a separate ozonation test was performed in a 70 L tank with water from kitchen sinks after intensive use during utensil washing. This smaller volume was selected because the daily discharge from kitchen facilities is lower than that from sinks, representing approximately 12% of the total daily wastewater volume in a household [11]. For this reason, a smaller accumulation volume and a shorter HRT were applied to the HGW.

3.2. Description of Equipment

The equipment used in this study for the treatment and disinfection of LGW included the following:

Multiparameter Meter: Used to measure the concentration of O₃ in water, the Super Pen 5, model AM 005 by Yalitech Instruments (Santiago City, Chile). This device allows for the measurement of pH, ORP, and temperature, with all parameters properly calibrated.

Turbidimeter: For turbidity measurements, a Lutron Electronic Enterprise turbidimeter, model TU-2016 (Taipei City, Taiwan), was used. This device features an automatic range of 0 to 50 NTU and 50 to 1000 NTU, allowing for accurate measurements at each time interval.

Ozone Generator: For in situ O₃ generation via corona discharge, an Ambiente Ozono AO3-0010 PLUS unit (Chillán City, Chile) rated at 230 W and supplied with ambient air, was used. The equipment integrates an O₂ concentrator and a silica gel layer with hygroscopic properties to remove moisture from the O₂ stream before entering the generator. The production is estimated at 10 g/h under optimal humidity and temperature conditions, with an approximate airflow of 0.24 Nm³/h. The maximum O₃ concentration in the injected air is estimated at approximately 42 gO₃/m³ or 42 mg/L in the gas phase.

To verify the results measured with portable instruments and to obtain additional relevant values for assessing water quality, samples are analyzed in an authorized laboratory, which conducts tests according to the following standards:

• Free residual chlorine according to St. Methods Ed. 23rd 2017-4500 Cl G.
• Turbidity—Nephelometric according to St. Methods Ed. 23rd 2017-2130 B
• Fecal coliforms according to Chilean standard Nch 2313/22:1995 [101]
• BOD₅ according to Chilean standard Nch 2313/5:2005 [102]
• Total suspended solids according to Chilean standard Nch 2313/3:1995 [103]
• pH according to Chilean standard Nch 2313/1:1995 [104]

4. Discussion and Results

It had already been mentioned that tests would be conducted for both LGW and HGW, and these tests were carried out in situ with the same generator and fine bubble diffuser plate. O₃ proves to be a powerful oxidant, but it requires higher levels of O₃ production to oxidize the available organic matter in HGW and achieve suitable quality for its reuse, as described in the literature [11,12,41]. These results would increase both equipment costs and energy consumption, and the values obtained for HGW are show in

Table 8. Ozonation results for high-load greywater (HGW).

Parameters	Raw Greywater	Treated Greywater	Reduction
pH	7.61	7.69	-
BOD5 mgO2/L	880	372	58%
Total Suspended Solids mg/L	193	52	73%
Turbidity UNT	164	124	24%
Fecal Coliforms NMP/100 mL	<2	<2	-

.

In the analysis of HGW, a significant reduction in BOD₅ and TSS is observed, which is consistent with previous studies reporting the effectiveness of ozone as an oxidizing agent for both biodegradable and non-biodegradable organic compounds. The removal of BOD₅ is mainly attributed to the direct oxidation of ozone on double bonds and functional groups present in organic matter, as well as the generation of hydroxyl radicals, which non-selectively attack refractory compounds. The reduction in TSS can be explained by the action of ozone breaking organic bonds and destabilizing colloids, thereby facilitating their sedimentation. However, although the removal percentages are significant, the final levels do not meet reuse standards due to the high initial load of the water. Results are shown in Table 8, and its profile can be observed in Figure 5.

In Figure 5, it can be observed that a steady state is reached at 50 min, without achieving an increase in concentration. The high BOD₅ values recorded align with those reported in previous studies on kitchen wastewater or HGW [105], which indicate values ranging from 100 to 1850 mg/L, for a pH between 5.6 and 8, and TSS levels between 134 and 1300 mg/L. To achieve better results, a mandatory pretreatment, such as an oil and grease separator chamber, must be included, which increases implementation and maintenance costs. For this reason, HGW is typically excluded, as it would require higher O₃ production from the generator or additional pretreatment elements; on its own, it does not comply with discharge regulations. For comparison purposes and to gather data for design, volumes of 2 and 20 L were saturated, in addition to the 200 L tank, solely to observe the behavior of O₃ according to the available volume for ozonation. This is illustrated in Figure 6.

It is evidently observed that the volume of water influences the achieved O₃ concentration, despite having the same contamination characteristics. This can simply be explained by the fact that a larger volume results in greater interaction between the present contaminants and O₃, referred to earlier as “ozone demand”; the greater the volume, the higher the demand. To obtain a proportional result, the equipment and production should also increase. Increasing the O₃ dosage enhances the driving force for O₃ transfer to the effluent, thereby increasing the generation of hydroxyl radicals, which improves degradation efficiency [22]. The rest of the analysis focuses exclusively on the 200 L volume, equivalent to the LGW volume in a household of four to five people, which can be replicated in similar homes or buildings that lack physical space. These systems are tailored to the contamination level and do not increase implementation costs, aligning with the pillars of sustainability in construction, one of which is economic sustainability [106].

Regarding the monitoring of ozone concentration in the system, factors such as temperature and contact time affect O₃ transfer into the liquid; however, after a certain time and with the liquid kept covered by a gas layer at constant concentration, equilibrium can be considered achieved, where the dissolved O₃ is equal to the amount of O₃ released [95]. Regarding response time, t90 is defined as the time required to reach 90% of the established concentration; for example, oxygen electrodes and O₃ probes generally reach t90 within 10 s [34], which explains the non-linearity in certain readings recorded by the measuring device.

With constant bubbling, equilibrium was achieved in 60 min for the 200 L tank. Once the equilibrium condition is reached, Henry’s Law Solubility can be used to calculate the amount of dissolved O₃ in the water, where H^cp = 0.00011 mol·m⁻³·Pa⁻¹ at a temperature T^cp of 298.15 K or 25 °C. Since the test temperature corresponds to 286.15 K, the new solubility is calculated by combining Equations (5) and (6):

$${\mathrm{H}}^{\mathrm{c}\mathrm{p}}\left(286.15\right)={\mathrm{H}}^{\mathrm{c}\mathrm{p}}\times \mathrm{e}\mathrm{x}\mathrm{p}\left(\left({\displaystyle \frac{\mathrm{d}\mathrm{L}\mathrm{n}{\mathrm{H}}^{\mathrm{c}\mathrm{p}}}{\mathrm{d}(1/\mathrm{T})}}\right)\left({\displaystyle \frac{1}{\mathrm{T}}}-{\displaystyle \frac{1}{{\mathrm{T}}^{\mathrm{c}\mathrm{p}}}}\right)\right)$$

(16)

Resulting in H^cp (286.15) = 0.00015 mol·m⁻³·Pa⁻¹, as expected, an increase in solubility is observed as temperature decreases. The saturation concentration of dissolved O₃ can be used along with Henry’s Law Solubility at the study temperature to estimate the O₃ gas concentration. If we now calculate the dimensionless Henry’s Law constant with the air-water partition coefficient H^cc = 0.00015 mol·m⁻³·Pa⁻¹ × 8.31 Pa·m³·mol⁻¹ K⁻¹ × 286.15 K, the result is H^cc = 0.356 (dimensionless) and its inverse K_H^cc = 2.81, which matches the solubility values and Henry’s Law constant in [24,34]. Given H^cc and ${\mathrm{c}}_{\mathrm{a}}^{\mathrm{*}}$, ${\mathrm{c}}_{\mathrm{g}}$ is obtained, where the latter becomes the average gas concentration at steady state.

The tests were designed to assess the system’s performance under different scenarios and establish the feasibility and limitations of O₃ based treatment.

Table 9. Calibration for converting ozone concentration from ORP to ppm [107].

mV-orp	ppm mg/L	mV-orp	ppm mg/L	mV-orp	ppm mg/L	mV-orp	ppm mg/L
100	0.00	875	1.00	1125	3.50	1375	6.00
200	0.04	900	1.25	1150	3.75	1400	6.25
300	0.08	925	1.50	1175	4.00	1425	6.50
400	0.13	950	1.75	1200	4.25	1450	6.75
500	0.16	975	2.00	1225	4.50	1475	7.00
600	0.20	1000	2.25	1250	4.75	1500	7.25
700	0.22	1025	2.50	1275	5.00	1525	7.50
750	0.25	1050	2.75	1300	5.25	1550	7.75
800	0.39	1075	3.00	1325	5.50	1575	8.00
860	0.50	1100	3.25	1350	5.75	1600	8.25

was used to perform unit conversions, such as from mV to ppm.

The ca∗ value measured with the probe corresponds to 469 mV according to Table 9, equivalent to 0.15 g/m³ or mg/L in the steady-state concentration. Dividing this by the molecular weight of O₃ of 48 g/mol, we obtain the molar concentration of ${\mathrm{c}}_{\mathrm{a}}^{\mathrm{*}}$ = 0.00314 mol/m³. Then, by isolating the gas concentration (cg) from the air–water partition coefficient H^cc, ${\mathrm{c}}_{\mathrm{g}}$ = 0.0088 mol/m³ or 0.42 g/m³ is obtained. The resulting equilibrium gas (O₃) transfer into water corresponds to 36%, a value very close to the 32% experimental mass transfer values reported in the literature [99] and close to the previously mentioned 10% production. The O₃ concentration levels achieved are shown in Figure 7, while the kinetic parameters k_La determined in the transient stage present the results in Figure 8.

A k_La slope value of 0.0009 s⁻¹ is obtained, which is close to the value reported by [44] for microbubbles, confirming behavior consistent with Henry’s Law.

The behavior of the slope in Figure 8 shows a variation at 2000 s or approximately 30 min of bubbling. It should be noted that O₃ initially reacts directly with the organic matter or contaminants present in the water. During this process, radicals are generated that continue to react indirectly. These radicals have a much higher redox potential than O₃, which leads to an increase in the concentration reading measured by the equipment. On the other hand, the indirect reaction forms radicals with lower reactivity that do not contribute to the formation of superoxide radicals, which in turn promotes the decomposition of O₃ [34,94]. Additionally, the half-life of O₃, which is approximately 25 to 30 min, should be considered, as it produces a small plateau in the concentration shown in Figure 7, before continuing to increase and eventually stabilizing.

As a general reference, for evidence and to corroborate the information reported by various authors, readings were taken during the ozonation process and graphed to observe their evolution and behavior throughout the tests, which can be reviewed in Figure 9 and Figure 10. The results clearly show no variation in pH, as evidenced in Figure 9, for both high- and low-contaminant load waters, remaining stable during the ozonation stage. This aligns with the literature on the quality of water to be treated, indicating that these pH levels are optimal for the behavior of O₃ in water, ensuring good results.

Regarding turbidity, a significant reduction is observed, as well as in total suspended solids. Despite the low concentration achieved, oxidation had a positive effect due to the retention time. This effect is further enhanced by the sedimentation time following ozonation, which promotes the settling of remaining particles. Readings for these parameters were consistently taken during in situ tests, and in the case of turbidity, multiple measurements were conducted for different trials, as reflected in Figure 10.

While some studies recommend treatment times of up to 4 h for high-load wastewater in advanced oxidation processes [108], this study prioritizes moderate use to minimize generator wear and energy consumption, a criterion further supported by the fact that low-load water is considered for the treatment.

This approach aligns with the design of sustainable buildings, which incorporates economic sustainability, green infrastructure, and water purification [106].

The generator consumes approximately 0.23 kWh per day, as it is designed to operate 1 h per day with a power rating of 230 W, translating to monthly operating costs of around 7 kWh, or approximately USD 2 per month, based on the kWh cost in Chile. This low consumption makes O₃ a cost-effective and accessible technology for small-scale GW treatment systems, especially in applications with low organic load or low BOD₅ values. Other studies estimate energy consumption below 0.01 kWh/m³ for larger installations [108]; in our case, consumption is approximately 1.1 kWh/m³, with costs and rates decreasing as the scale increases.

These results are compared with those obtained by the instruments. Additionally, we evaluate the disinfection and contaminant oxidation efficiencies, as described in

Table 10. Comparison of treated water with regulatory requirements.

Parameters	Urban Use [71]	Irrigation for Recreational [71]		Afluent Greywater Quality	Efluent Greywater Quality	Reduction
	Irrigation or Toilets	Surface	Subsurface
pH	-	-	-	7.24	7.24	-
BOD5 mgO2/L	10	30	50	10	<2	80%
Total Suspended Solids (mg/L)	10	30	50	33	<10	70%
Turbidity UNT	5	10	-	4.8	1.7	65%
Fecal Coliforms NMP/100 mL	10	200	1000	2	<2	1 Log

, which provides the limit values for treated water for urban use, recreational irrigation, and ornamental irrigation as required by the current regulations.

The results shown in the treated water quality in Table 10 indicate reductions of over 65% in turbidity and TSS, and an 80% reduction in BOD₅ for low-contaminant load water, fully meeting the objective of this study and the required standards for reuse in irrigation and toilet flushing, as shown in Table 2 and Table 3. Regarding fecal coliforms, these are already low in LGW from handwashing sinks, so they are assumed to reach almost a logarithmic reduction level after ozonation, considering that the Ct ratio achieved in this case with only 5 min of ozonation is 0.75 mg·min/L, higher than the 0.4 and 0.48 mg·min/L values evidenced in [109].

5. Conclusions

Aqueous O₃ is confirmed to be effective in contaminant reduction and more effective than chlorine for microorganism elimination, further reinforced by the low levels of these in LGW. This positions O₃ as a safe and environmentally friendly option for GW treatment, ensuring effective contaminant reduction in small-scale, low-contaminant-load systems, in comparison with other disinfectants.

HGW does not meet the country’s current water quality standards for reuse, as identified in Table 2 and Table 3, confirming its exclusion from treatment systems and validating the reuse of LGW in small-scale systems. However, this does not imply that, with the consideration of appropriate technology and systems, it cannot be reused in the same way.

For a generator with a production capacity of 10 g/h, this study considers a minimum ozonation time of 60 min to be adequate based on the results obtained in the tests and the findings in the literature, with an HRT of 40 min in a volume not exceeding 200 L. This ensures a Ct value above 6 mg·min/L, sufficient for excellent disinfection in LGW. It is proposed that bubbling occur overnight when there is no significant flow to alter the HRT, thus simulating a batch system that enables even more efficient and prolonged treatment of LGW.

Temperature, which is considered as a potential factor affecting O₃ efficacy due to its influence on solubility and decomposition, did not show significant effects on O₃ concentrations at 13 °C and 20 °C. However, according to the literature, changes become significant below 5 °C. This could be considered a practical range for performing bubbling or could be associated with climatic conditions similar to those in central Chile.

Due to its compact size, efficiency, and ease of implementation, this technology is suitable for homes and small facilities in schools, universities, or offices, enabling the in situ reuse of water from handwashing sinks. Despite being a small and individual system, it is replicable for communities while maintaining the same volumes.

Given the characteristics of this water and the effective action of O₃, it ensures safe and non-harmful levels for people and the environment, meeting discharge standards and directly contributing to Sustainable Development Goal No. 6: Clean Water and Sanitation.

The literature review shows little evidence of greywater treatment with O₃ and is rather scarce, making it relevant to continue researching and expanding the evidence for this type of system. This would help establish a more precise range of contaminant parameters and potential volumes to be treated.

To address these challenges, future research should focus on:

• Treatment of water with a higher contaminant load, such as water from showers, as it represents a considerable volume and lower contamination than HGW.
• Feasibility of scaling up the presented system to assess the possibility of treating larger volumes efficiently while complying with regulations.
• Testing the combination of physical systems such as sand and carbon filters together with higher-production O₃ generators to evaluate the effectiveness in medium- and high-load systems without a significant increase in costs.