Treatment of the spin cycle effluent using electrocoagulation and granular activated carbon Open Access

Since their introduction, washing machines have become a staple in households worldwide with millions of units in use (Pakula & Stamminger 2010). Washing machines utilize nearly 10 L of water to launder 1 kg of clothes which makes them a water-intensive appliance (Akarsu & Deniz 2021). Washing machines use around 19 billion m³ of freshwater every year and hence are major contributors to the household freshwater demand (Barthel & Götz 2013). Due to the rise in global population and standard of living, there is a rapid surge in demand for freshwater which is creating pressure on freshwater sources. The current state of the water supply is made worse by uneven resource distribution, extreme droughts, groundwater depletion, declining water quality, and climate change (Chen et al. 2013). One option to reduce household water demand is by treating and recycling the domestic washing machine effluent to form a closed-loop system.

The washing machine process consists of three cycles: the main wash cycle, the rinse cycle, and the spin cycle in succession (Mathews 2015). Water is taken at the start of each cycle and used for laundering the clothes, at the end of the cycle, the water is drained out of the machine as an effluent. The main wash cycle is the primary cycle and uses detergent mixed with water to wash clothes using agitation achieved by slow rotation of the washing machine drum. This cycle removes the maximum amount of dirt from the clothes and produces higher-strength wastewater. The rinse cycle removes the residual dirt and detergent and produces effluent of less strength than the main wash cycle. The spin cycle utilizes high rotation speeds to remove excess water from the clothes and this water also contains residual detergent and dirt that were not removed during the first and second wash cycles. The spin cycle utilizes rotation speeds in the range of 400–1,400 rpm and is typically operated at 1,000 rpm for domestic laundry. The spin cycle, hence, has a lower strength for its effluent water when compared with the main wash cycle and the rinse cycle. Due to its lower strength, it requires less intensive treatment before it can be reused. If this water is treated, it can be reused for household purposes or even reused inside the washing machine for the next main wash cycle.

Although the literature did not contain any studies on spin cycle effluent treatment or its characterization, several studies have attempted the treatment of laundry wastewater and washing machine effluents. Physical and chemical treatment methods, such as membrane filtration, chemical coagulation, electrocoagulation (ECG), and adsorption along with anaerobic and aerobic biological treatments, were employed previously to treat laundry wastewater (Kumar et al. 2022). Laundry water has high chemical oxygen demand (COD) and turbidity content due to the presence of detergent, dirt, and grease (Al Hinai 2020). Phosphates are introduced in laundry wastewater through phosphate-containing detergents. If such water is discharged into water bodies without treatment, it leads to eutrophication (Kundu et al. 2015). Recent research on washing machine effluent treatment has highlighted the effectiveness of various methods. Dos Santos et al. (2018) demonstrated complete color and organic removal using the solar photoelectro-Fenton process, while Duran et al. (2018) found that electro-oxidation with boron-doped diamond (BDD) anodes achieved higher removal efficiency and faster COD decay compared with Ti/Pt anodes. Rodrigues et al. (2022) achieved significant turbidity removal using aluminum sulfate and Moringa oleifera seed extract. Additionally, Ghanbari & Martinez-Huitle (2019) reported nearly complete organic removal with the photoelectro-Fenton process combined with peroxymonosulfate. Furthermore, Oliveira et al. (2023) investigated a BDD electrode as the anode and a Ni–Fe-based SS (stainless steel) mesh as the cathode and solar-powered polymer electrolyte membrane (PEM-type) cell to treat the washing machine effluent while simultaneously producing green hydrogen and carboxylic acids.

However, these methods have notable disadvantages. The photoelectro-Fenton process requires acidification and the addition of iron, making it complex and costly, and it depends on sunlight availability, which can be inconsistent (Dos Santos et al. 2018). Electro-oxidation, while effective, comes with several disadvantages such as high energy consumption, potential generation of toxic byproducts, short electrode lifespan due to fouling, and the need for supporting electrolytes, which increase operational costs (Navarro-Franco et al. 2022). Additionally, it requires expensive electrodes such as Ti/Pt and BDD, resulting in high initial costs. For instance, washing machine effluent treatment, as demonstrated by Duran et al. (2018), achieved more than 90% total organic carbon (TOC) reduction but required costly BDD electrodes and the addition of 7 g/L of Na₂SO₄, which introduces undesirable sulfate ions and adds to the overall cost.

Coagulation of the washing machine effluent also presents challenges, including the production of significant amounts of sludge and the potential introduction of undesirable ions into the water. For example, Rodrigues et al. (2022) used aluminum sulfate and M. oleifera for coagulating the washing machine effluent. However, the use of aluminum sulfate introduces unwanted sulfate anions into the water (Wang & Zhang 2019), while M. oleifera requires extraction with calcium nitrate, which adds to the cost and introduces undesirable nitrate ions in water (Shrimali & Singh 2001). If the dosage is not optimized, Moringa extract, being organic, can also cause higher COD and BOD levels in treated water due to improper coagulation. The use of compounds like peroxymonosulfate for washing machine effluent treatment (Ghanbari & Martinez-Huitle 2019) is another approach, but it is expensive and requires careful handling. These challenges highlight the need for a quick, safe, and cost-effective method that does not introduce undesirable anions or produce excessive sludge. One potential method that meets these criteria is electrocoagulation.

Electrocoagulation is a well-known wastewater treatment technology in which current is applied across metal electrodes resulting in the electrodissolution of sacrificial anodes which gives rise to coagulant species in water. Electrocoagulation has many advantages such as lower sludge production than chemical coagulation, easy automation, and control of the process via galvanostatic or potentiostatic mode. Electrocoagulation treatment with aluminum electrodes has been found to be highly effective in the treatment of COD, turbidity, and phosphorous content of laundry water. Janpoor et al. (2011) showed that electrocoagulation with aluminum electrodes is an effective technology for the treatment of laundry effluent with a 95.9% reduction in turbidity, a 93.2% reduction in COD, and a 96.7% reduction in phosphorous content. Wang et al. (2009) showed a reduction of 62% in COD for simulated laundry water using simultaneous electrocoagulation and ultrasound treatment. However, the treatment has not been specifically utilized for treating the spin cycle effluent. Research showed that a four-plate bipolar setup was superior to a monopolar setup due to its higher treatment efficiency (Janpoor et al. 2011). The chemical reactions that occur during electrocoagulation using aluminum electrodes given by Janpoor et al. (2011) are as follows.

Under appropriate pH conditions, aluminum hydroxide species hydrolyze and polymerize to form monomeric species such as ⁠, ⁠, ⁠, and and polymeric species such as ⁠, ⁠, ⁠, ⁠, and ⁠. These species effectively remove pollutants through adsorption, charge neutralization, and precipitate enmeshment.

Polymeric species have reactive groups that bind to colloidal particle surfaces, with the remaining long-chain molecule extending into wastewater. The polymer can bridge the gap between colloidal particles. If no other colloidal particles are available, it can wrap around the original particle and stabilize it. Aggressive mixing or extended agitation can cause restabilization by breaking interparticle bridging and allowing freed polymeric sections to enclose the original particles. Electrocoagulation creates metal hydroxide flocs with high surface areas, allowing for faster adsorption of soluble organic compounds and the capture of colloidal particles. A mixing step with a stirring mechanism can be used to increase the size of the flocs and aid flocculation until they are heavy enough to settle, leaving behind a clean supernatant.

As summarized by Mohammadi et al. (2019), several mechanisms are involved in the removal of total nitrogen (TN), which consists of organic nitrogen, ammonium, nitrite, and nitrate during electrocoagulation. Chloride ions naturally present in wastewater react under an applied electrical current, producing chlorine gas, which is then hydrolyzed into hypochlorous acid and hypochlorite ions – highly oxidative compounds capable of converting ammonium into nitrogen gas. In addition, nitrates can be reduced to nitrogen gas and ammonia near the cathode, with the ammonia potentially undergoing further oxidation into nitrogen gas. Mohammadi et al. (2019) also described the formation of hydroxo-amorphous polymeric complexes due to anode dissolution, which act as adsorbents for nitrite and nitrate, resulting in floc formation and precipitation. Some lighter flocs may attach to hydrogen bubbles and float to the surface, further aiding in nitrate and nitrite removal.

Washing machines typically operate at temperatures between 25 and 60 °C. Cold washes are often used for gentle washing cycles, while hot washes are employed when more aggressive cleaning is required typically in the case of heavily soiled clothing. The temperature variability in washing machines can thus significantly impact the electrocoagulation process and its efficiency. Dimoglo et al. (2019) found that as the temperature increased at a constant current density, aluminum yield in the electrocoagulation of laundry water improved up to a certain threshold (60–80 °C). Beyond this point, the yield decreased. This decline was attributed to the increased passivation of aluminum anodes, caused by the compaction and swelling of colloidal aluminum hydroxide in the micropores of the material. Although there is limited research specifically examining how temperature affects electrocoagulation efficiency in laundry water, other studies on different types of wastewater have shown both improvements and declines in treatment efficiency as the temperature rises (Yilmaz et al. 2008; Vasudevan et al. 2009; Katal & Pahlavanzadeh 2011; Attour et al. 2014).

However, while electrocoagulation followed by flocculation is excellent for removing suspended and colloidal impurities, it is relatively less effective at removing dissolved pollutants. As a result, granular activated carbon (GAC) was tested following flocculation to further reduce dissolved pollutants in the water. GAC has a large surface area, porous structure, and high adsorption capacity, making it an excellent adsorbent. Activated carbon adsorption is a well-established method for removing organic and inorganic pollutants from water and wastewater. This method is effective at removing surfactants, total dissolved solids, dyes, pesticides, personal care products, nitrates, etc., from water (Jjagwe et al. 2021). Activated carbon was traditionally made out of coal or charcoal but now many agricultural wastes, such as coconut shells, rice husk, and almond shells, are being used. The coconut shell-based activated carbon is an attractive option as it is a more sustainable alternative than traditional coal-based activated carbon and its raw material, i.e., waste coconut shells are available in abundance in countries like India (Vilén et al. 2022). Due to these reasons, coconut shell-based GAC was utilized in this study. Adsorption of pollutants on GAC takes place either due to chemisorption or through the physisorption mechanism (Velasco & Ania 2011). The use of GAC post-electrocoagulation has been tested before for wood-based industry wastewater, dairy wastewater, etc., but no such study has been conducted for spin cycle effluent treatment (Hansson et al. 2014; Eulmi et al. 2019). In this study, an optimum time for electrocoagulation was determined based on COD reduction. Once the samples were treated for the optimum time, they were then further treated using GAC to lower the pollutant concentration within permissible limits. The treatment efficiency after each treatment step was calculated. Solid and liquid detergents were employed in this study to generate the spin cycle effluent.

This research focuses specifically on spin cycle effluent treatment, an area that has not yet been explored. The combination of electrocoagulation and GAC treatment was employed in this study to reduce COD levels to below 50 ppm. The spin cycle effluent was chosen specifically for the advantage of having a lower pollutant load compared with the combined laundry effluent, allowing for a less intensive treatment to achieve water reusability. The study utilized solid and liquid detergents and a thorough analysis was done concerning characterization, treatment time optimization, and cost of electrocoagulation. The findings from this research will help the development of treatment systems that can enable treated spin cycle water to be reused in the washing machine for the main wash cycle or for other non-potable applications, such as toilet flushing, gardening, or car washing. This aligns with the broader goal of reducing freshwater demand in households and promoting more sustainable water practices.

For our study, the EXECUTIVE PLUS VXR 8.5KG washing machine sold by IFB Ltd was procured and was operated at full load in cotton normal mode at a temperature of 40 °C and a spin speed of 1,000 rpm. The dosage of solid or liquid detergent for the wash is 60 g or 1 scoop for solid detergent and 60 mL for liquid detergent. The mode consists of three cycles: the first cycle is the main wash cycle which lasts for 120 min followed by a rinse cycle of 22 min, and finally, the spin cycle of 16 min. Soiled clothing used for the washing was collected from university students and was part of their daily wear. The water from the spin cycle was collected from the drain pipe of the washing machine in a clean plastic bucket and mixed thoroughly and the sample necessary for the study was drawn from this in sterilized sampling bottles. Six popular detergents, three solid, and three liquid detergents, were purchased from the local market. The detergents used were labeled as a type of detergent (solid/liquid) followed by letters A, B, and C for the sake of convenience in this study. The actual names of the detergent are as follows: Surf excel Matic front load powder detergent (solid det A), Ariel Matic front load powder detergent (solid det B), Tide Ultra Antigerm for front and top load powder detergent (solid det C), Ariel Matic liquid detergent front load (liquid det A), IFB fluff Matic liquid detergent front load (liquid det B), and Surf excel Matic liquid detergent front load (liquid det C).

Water samples were analyzed for pH using an Oakton Ion 700 benchtop pH meter. The conductivity was analyzed using a digital conductivity meter model no. CC-01 which was procured from CONTECH Instruments Limited. Biochemical oxygen demand (BOD) for 5 days at 20 °C was estimated using WTW oxitop bottles sold by Xylem Analytics Germany GmbH. The turbidity of the samples was measured using a digital nephelometer model no. 341 procured from Electronics India Limited. The COD of the water samples was measured using the closed reflux colorimetric method (Rice et al. 2012). Whereas orthophosphate concentration was measured using the Vanadomolybdophospheric acid colorimetric method (Rice et al. 2012). The TN of water samples was measured using the total nitrogen measuring (TNM-L) unit from Shimadzu. The weighing of aluminum electrodes was carried out using the weighing balance model AJ-1200E from Shinko Denshi Co. Ltd. COD and orthophosphate analyses were done in duplicates, whereas the rest of the parameters were done in singlicates. A hot air oven manufactured by Meta-Lab Scientific Industries was utilized to dry the electrode plates before weighing.

The washing machine was operated 12 times to generate spin cycle effluent samples, as described in Section 2.1. Each of the six detergents, as mentioned in Section 2.1, was utilized for two separate operations. Two liters of spin cycle samples obtained after each operation were tested for pH, conductivity, turbidity, COD, BOD, orthophosphate, and TN.

For the electrocoagulation reactor, four rectangular electrodes of size 13.5 cm × 11 cm, a handle dimension of 10.5 cm × 2 cm, and a thickness of 4 mm made of aluminum (99.9% purity) were purchased from Ti Anode Fabricators Pvt Limited (Figure 1). Only the rectangular part of the electrode, with an area of 148.5 cm², was submerged in the liquid sample and thus participated in the reactions. The electrode handle, on the other hand, was kept outside the liquid. All four electrodes were connected in a bipolar setup with an electrode spacing of 1 cm and electrode surfaces parallel to each other, as shown in Figure 1. The electrode spacing was maintained using 3D-printed 1 cm spacers between the electrode handles. Four electrode handles with three spacers in between them were compressed from both sides with a 3D-printed clamp. The material used for 3D printing was polylactic acid (PLA). A 2-L glass beaker (Borosil) was filled with 2 L of the spin cycle effluent sample and the rectangular portion of all the electrodes was submerged in the sample in such a way that it did not touch the bottom of the beaker. Electrodes on the extreme ends were connected to a laboratory DC power supply (GWINSTEK GPS-4303) in potentiostatic mode with a supply voltage fixed at 30 V.

For the GAC treatment, a cubical container of inner dimension volume of 20 cm × 20 cm × 20 cm was fabricated from Plexiglass of 1 cm thickness. The container was filled with 4 kg of GAC. The activated carbon had iodine no. 1200, size 2–4 mm, and was made from coconut shells. It was procured from Numatik Engineers Pvt Limited. The container can be opened from the top so that a water sample can be poured into it. The container is also equipped with a tap at the bottom to drain the water. A cotton plug was used to stop carbon particles from escaping into the tap to avoid clogging of the tap.

The washing machine was operated six times with six different detergents and the spin cycle effluent was collected. Three solid detergents and three liquid detergents of popular brands purchased from the consumer market were used for this study, as mentioned previously. A 12-L spin cycle effluent for each sample was taken in six 2-L beakers. Each 2-L beaker was subjected to electrocoagulation for a fixed time. The electrocoagulation was done for 5, 10, 15, 20, 25, and 30 min. The same electrode assembly was used for all the beakers sequentially.

The supernatant was achieved by treating 2 L of water for 15 min of electrocoagulation followed by mixing and settling for the 12 samples mentioned previously were poured into the GAC setup from the top. After a contact time of 6 min, the treated water was drained from the bottom and analyzed for pH, conductivity, turbidity, COD, orthophosphate, and TN.

Data from treatment efficiency and conductivity measurements for solid and liquid detergent samples were analyzed using Microsoft Excel. For treatment efficiency, both groups had a sample size of n = 6, while for conductivity, both groups had n = 12. Data were assumed to follow the normal distribution, and outliers were removed by eliminating points beyond 1.5 times the interquartile range. No normalization was applied as the raw data were deemed appropriate for analysis. Separate two-sample t-tests assuming unequal variances (Welch's t-test) were conducted for treatment efficiency and conductivity. The choice of this test was based on an F-test confirming unequal variances, ensuring robustness when comparing the independent groups.

The null hypothesis (H₀) for treatment efficiency was that there was no significant difference in the mean treatment efficiency between solid and liquid detergent samples. Similarly, for conductivity, the null hypothesis was that there was no significant difference in the mean conductivity between the two groups. A p-value of less than 0.05 was considered statistically significant, and rejection of the null hypothesis indicated a significant difference between the groups.

The mean values for COD, BOD, pH, conductivity, turbidity, TN, and orthophosphate for the spin cycle effluent were found to be 241.8 ± 88.4 ppm, 36.9 ± 14.3 ppm, 7.2 ± 0.9, 223.4 ± 46.7 μS/cm, 85.9 ± 20.1 NTU, 7.2 ± 3.1 ppm, and 4.6 ± 1.7 ppm, respectively. As shown in Table 1, when compared with the washing machine effluent parameters reported by Ghanbari & Martinez-Huitle (2019), the mean COD, BOD, and conductivity values for the spin cycle effluent are lower. This is because the main wash and rinse cycles remove most of the particulate and dissolved carbon-containing pollutants, leaving only residual carbonaceous matter in the spin cycle, which results in a substantial reduction in these values. Additionally, the number of ions is drastically reduced after both the washing and rinse steps, leading to much lower conductivity in the spin cycle effluent compared with the washing machine effluent. However, the mean pH of the spin cycle effluent was slightly higher than that reported by Ghanbari & Martinez-Huitle (2019).

The domestic laundry effluent range presented in the review by Kumar et al. (2022) shows a similar trend in terms of COD and BOD. However, the conductivity of the spin cycle effluent falls at the lower end of their reported range. The lower COD and BOD levels indicate that the wastewater is of lower strength, suggesting that it requires less intensive treatment before reuse. The pH of the spin cycle effluent, however, was slightly lower compared with the range given by Kumar et al. (2022).

The graph for COD vs time, as shown in Figure 2(a) and 2(b), for all the detergents shows that, on average, maximum reduction was obtained at 15 min beyond which COD either increased or remained the same. Anodic dissolution during electrocoagulation releases aluminum ions in water which forms aluminum hydroxide species that are positively charged and act as coagulating agents. These species neutralize the negative charge of suspended particulate matter, thus causing charge neutralization. After charge neutralization, the suspended particulate matter flocculates and settles down. COD reduction is observed due to a reduction in suspended matter containing carbon. Reduction is also caused due to the adsorption of pollutants on the flocs which precipitate (Tegladza et al. 2021). The amount of coagulating agent generated in situ directly depends on duration. As the duration for which the current is passed through the electrodes increases, the amount of anodic dissolution increases, thus causing an increase of coagulating agents in the water. However, after a certain point, all the positive charges will be neutralized. This corresponds to an optimum dosage point and the time corresponding to this would be the optimum treatment time. After this point, the excess dosage will not cause any more coagulation, it might even restabilize the suspended particles, thus reducing flocculation and COD removal (Ghernaout et al. 2011). Hydrogen gas bubbles released during the reaction also cause the lighter floc to float on the surface, a phenomenon known as electroflotation. However, in this study, the floc eventually settled post-stirring and settling, as the stirring step increased the size and density of the flocs. Along with precipitation, direct anodic oxidation and indirect oxidation via hypochlorous and hypochlorite ions also play a role in the reduction of COD (Janpoor et al. 2011).

Although no studies are reported for spin cycle effluent treatment, Janpoor et al. (2011) showed the most resemblance with the operational parameters and electrode configurations utilized in this study. Although they worked on laundry wastewater, they have also demonstrated the use of a bipolar electrocoagulation setup consisting of aluminum electrodes operated at 30 V with a similar surface area as the electrodes used in this study. But they treated a smaller amount of water (1.5 L) and used a greater inter-electrode distance of 1.5 cm. Their study reported a COD reduction of 79.7% post-electrocoagulation at 15 min which was higher than the average reduction of 67.8 ± 8% achieved in this study; however, they reported higher COD removal as reaction time was increased beyond 15 min.

Several works in the literature report using various operating conditions and electrode configurations, different from the current study, to achieve electrocoagulation for laundry water and graywater using aluminum as the cathode and anode (see Table 2). The current study, along with the maximum treatment efficiency reported by Janpoor et al. (2011), is included in the table for comparison. The table shows that the treatment efficiency achieved for the spin cycle effluent in this study was higher than those reported by Wang et al. (2009) for laundry water and Nasr et al. (2016) for graywater, but lower than that achieved in other studies.

Chen et al. (2000), however, proposed other possible mechanisms as well, including the release of CO₂, which becomes oversaturated in acidic solutions and is released due to H₂ bubble formation, and the exchange of anions like Cl⁻ and with OH⁻ ions in Al (OH)₃, both leading to pH increase. However, in our study, the initial pH was near neutral, so the CO₂ stripping effect would be muted. Chloride and sulfate ions are expected in the spin cycle effluent due to their presence in detergents (Bajpai & Tyagi 2007). Hence, the anion exchange effect on pH cannot be ruled out.

A decrease in pH was observed after GAC treatment. This effect is likely due to the presence of acidic functional groups within the GAC. These functional groups interact with and neutralize the basic components present in the effluent, leading to a reduction in pH. A similar effect was reported by Vargas & Lopes (2020), who observed a reduction in pH after passing the raw laundry effluent through activated carbon derived from flamboyant pods, demonstrating the influence of GAC on the pH of the treated effluent. The pH of treated water was found to be within the range of 6.5–7.5 and was thus safe for discharge into the environment (Ministry of Environment and Forest 1986). Figure 4(b) depicts the conductivity of effluent water produced by solid and liquid detergents. The conductivity values for effluent from solid detergents were significantly higher (p = 0.0002) compared with those from liquid detergents. This disparity can be attributed to the elevated concentration of ions in solid detergents. The water before GAC treatment had relatively low conductivity, yet post-treatment, a major increase in conductivity was observed. This indicates that GAC must have released ions into the water, which contributed to the higher conductivity. This phenomenon was unexpected, as previous studies, such as those by Vargas & Lopes (2020), who attempted the treatment of the laundry effluent using activated carbon derived from flamboyant pods, did not observe an appreciable change in conductivity post-treatment. In their study, both raw and treated laundry effluents exhibited no appreciable shifts in conductivity, suggesting that the specific characteristics of activated carbon used are influencing the outcome. To determine the exact cause and mechanism for this increase, a separate investigation is necessary.

The variance in overall reduction was found to be much lower than that found in individual treatment steps. GAC effectively removes turbidity from water through a combination of adsorption and filtration mechanisms. The porous structure of GAC adsorbs fine particulates, while its granular form physically filters out suspended solids, thereby reducing turbidity levels. Hatt et al. (2013) investigated turbidity removal using GAC for the secondary treated wastewater using a different bed design and operating parameters with similar influent turbidity levels, but this demonstrated very high turbidity removal more than 80% compared with the below 50% reduction. In this case, the indicated bed design and operating parameters along with the characteristics of influent water play a role in removal efficiency.

It was observed that the current had negligible variation within ±0.02 A during the 15 min of reaction time, and the remaining was nearly constant. The current values at each time point (0, 5, 10, and 15 min) were averaged to obtain the average current for each of the 12 trials, and then further averaged by the detergent type. The average current for solid detergent was 0.58 ± 0.08 A, and for liquid detergent, it was 0.27 ± 0.02 A. Consequently, the average current density was 3.905 mA/cm² for solid detergent and 1.818 mA/cm² for liquid detergent.

These current densities are low compared with the optimum current density of 5.5 mA/cm² utilized by Dimoglo et al. (2019) for 5 min and 10 mA/cm² by Tabash et al. (2024) for 50 min in electrocoagulating laundry water. The distinct current density requirements for solid and liquid detergents underscore the importance of tailoring operational parameters to the specific characteristics of wastewater. This understanding can guide future system designs and optimizations for various types of detergent wastewater.

The average power consumed for the 15 min of electrocoagulation for the 2-L scale study in the case of solid detergents was 15.9 ± 2.4 W. For liquid detergents, however, the average power consumption was 8.2 ± 0.7 W. The average current mentioned in the previous section was multiplied by 30 V to get the average power consumption. This is due to a relatively lower current value during potentiostatic electrocoagulation owing to the lower conductivity of spin cycle water generated using liquid detergents. For treating 1 m³ of water, our study found that using liquid detergent required 1.025 kWh/m³, while solid detergent needed 1.9875 kWh/m³. When compared with previous studies, the energy consumption for liquid detergent was lower than the 1.25 kWh/m³ established by Dimoglo et al. (2019) as optimal for laundry water treatment under controlled conditions (5.26 mA/cm² current density, pH 5.5, and 5-min treatment time). In contrast, the solid detergent exceeded this benchmark, suggesting a higher energy demand. Both treatments, however, required considerably less energy than the 20.54 kWh/m³ reported by Caetano et al. (2022) for laundromat graywater treated under harsher conditions (400 A/m² current density and 17.14 V over 30 min), highlighting the relative efficiency of our methods. The lower energy consumption of liquid detergent suggests that it may be particularly suitable for systems with limited energy resources or for facilities prioritizing energy conservation and environmental sustainability.

As aluminum anode dissolves, it releases aluminum ions which reduce the weight of the electrode. The weight reduction due to 15 min of electrocoagulation gives us an idea about the aluminum dosage required to treat 2 L of spin cycle water. The average dosage for treating 2 L of spin cycle water is 342.8 ± 14 mg in the case of solid detergents and 118.5 ± 6 mg in the case of liquid detergents. Since electrocoagulation was carried out at constant voltage, the amount of current was dependent on the electrical conductivity of the spin cycle effluent. Due to the higher conductivity of the effluent generated using solid detergent, higher current and consequentially higher anodic dissolution were observed.

The electricity cost is calculated based on the average power consumption during the 15-min electrocoagulation process for both solid and liquid detergents at a 1 m³ scale. The formula used to determine the electricity cost is:

• Electricity cost (€) = Power (kW) × Time (h) × Electricity rate

For solid detergents:

• Electricity cost = 7.95 kW × 0.25 h × 0.1294 €/kWh = 0.2572 €

For liquid detergents:

• Electricity cost = 4.1 kW × 0.25 h × 0.1294 €/kWh = 0.1328 €

Thus, the electricity cost for treating 1 m³ of wastewater is:

• Solid detergents: 0.2572 €

• Liquid detergents: 0.1328 €

The cost of aluminum plates dissolved during the electrocoagulation process is calculated using the formula given by Silva et al. (2022):

• DP = (TM × AL)/ρ

For solid detergents:

• Aluminum cost = (171.4 g × 32,888 €/m³)/2,700,000 g/m³ = 2.09 €

For liquid detergents:

• Aluminum cost = (59.25 g × 32,888 €/m³)/2,700,000 g/m³ = 0.72 €

Thus, the cost of dissolved aluminum plates for treating 1 m³ of wastewater is:

• Solid detergents: 2.09 €

• Liquid detergents: 0.72 €

The total cost for treating 1 m³ of wastewater is the sum of the electricity cost and the aluminum plate cost.

For solid detergents:

• Total cost = 0.2572 € (electricity) + 2.09 € (aluminum) = 2.35 €

For liquid detergents:

• Total cost = 0.1328 € (electricity) + 0.72 € (aluminum) = 0.85 €

Thus, the total cost for treating 1 m³ of wastewater is:

• Solid detergents: 2.35 €

• Liquid detergents: 0.85 €

As of 6 November 2024, the exchange rate is approximately 1 € to 1.09 USD.

Using this rate, the prices are converted as follows:

• Solid detergents: €2.35 × 1.09 = $2.56

• Liquid detergents: €0.85 × 1.09 = $0.93

One important aspect to consider in this cost analysis is that no additional cost is incurred for sludge disposal. The amount of sludge generated at the household level is low and can easily be released into the sewer system, where it will be diluted with other wastewater. Therefore, sludge disposal does not require any special handling or incur any significant costs. There are also no extra reagents added to increase conductivity; hence, there is also no reagent cost. In our study, the unit costs considered were $13.20/kg for aluminum and $0.14/kWh for electricity, leading to a total treatment cost of approximately $2.56/m³ for solid detergents and $0.93/m³ for liquid detergents. For comparison, Yaranal et al. (2023) reported a slightly lower operational cost of $0.53/m³ for general laundry wastewater treatment via electrocoagulation. This lower value can be attributed to highly economical unit costs for aluminum and electricity in their calculations ($2.82/kg for aluminum and $0.0924/kWh). In contrast, Caetano et al. (2022) observed a significantly higher treatment cost of $4.10/m³ for laundromat greywater treatment (unit costs of $6.23/kg for aluminum and $0.180/kWh for electricity).

These findings suggest that treating household-level spin cycle wastewater via electrocoagulation with the methodology used in this study, especially with liquid detergents, remains cost-effective compared with laundry effluents.

In this study, the spin cycle effluent was characterized, and it was found to have lower pollutant concentrations than reported for laundry water and washing machine effluents, making it an appealing target for treatment and reuse. The treatment efficiency of the process, involving electrocoagulation with aluminum electrodes, mixing, settling, and GAC treatment steps carried out in succession, was tested at a 2-L scale. The treatment method achieved excellent reductions, successfully lowering COD by 92.7 ± 3.6%, orthophosphate by 91.4 ± 2.9%, TN by 87.2 ± 9.7%, and turbidity by 93.3 ± 4.7%. The final water quality was found to meet local discharge standards. The optimal electrocoagulation time for COD reduction, using a bipolar setup with 30 V applied voltage and an aluminum electrode area of 148.5 cm² was determined to be 15 min at a 2-L scale. There was no significant difference in overall treatment efficiency between the evaluated parameters for solid- and liquid detergent-generated spin cycle effluents. Also, there was no significant difference in treatment efficiencies between solid- and liquid-generated spin cycle effluents with respect to electrocoagulation for the measured parameters, with the exception of TN, where solid detergent-generated samples exhibited slightly better treatment efficiency. The solid detergent-generated samples had significantly higher conductivity than the liquid detergent-generated samples. The power consumption for electrocoagulation varied depending on the detergent type, with samples generated using solid detergents requiring higher power due to their higher conductivity. The solid detergent-generated spin cycle effluent utilized higher aluminum dosage than its liquid counterparts.

Overall, the treatment method was shown to be effective for the treatment of both solid- and liquid detergent-generated spin cycle effluents. The treatment method was rapid, lasting only 21 min. It is also cost-efficient because it uses low-cost materials such as aluminum electrodes and GAC, and the procedure also consumes very little electrical power. The approach is safe because no hazardous chemicals were used during the operation. Further scope of this study encompasses determining the longevity of GAC, given its eventual saturation.

The authors thank the Birla Institute of Technology and Science, K K Birla Goa Campus for providing laboratory facilities.

This research did not receive any specific grant from funding agencies in the public, commercial, or not-for-profit sector.

All relevant data are included in the paper or its Supplementary Information.

The authors declare that they have a patent awarded for the spin cycle effluent treatment process. The authors also confirm that this did not influence the design, execution, or interpretation of the research presented.