Due to its high poly(vinyl) alcohol (PVA) concentration, Embroidery industry wastewater is very resistant to biodegradation and classical methods such as coagulation, and adsorption. On the other hand, high chemical consumption is required in chemical oxidation processes. In this study, wastewater containing PVA was treated by an electrooxidation process with Boron Doped Diamond (BDD) electrodes for the first time in literature. The effects of pH (2.0–8.0), temperature (25–45 ̊C), current (3.0–6.0 A), electrolyte concentration (NaCl of 3.0–8.0 g/L), and oxidation time (0-600 min) were determined. The total organic carbon (TOC) concentration in the wastewater was decreased from 6584 mg/L to 159.10 mg/L (97.6%) at pH 5.01, 6.0 A, 3.0 g/L of electrolyte concentration, 25°C, and 10 hours. In addition, the removal mechanism of the process was elucidated by using HPSEC, FT-IR, UV-VIS, and Fluorescence Spectroscopy methods. Characterization studies showed that the high amount of PVA in the embroidery process wastewater is degraded and the halogens have attached to the part of the polymer where the structure has deteriorated through oxidation. Raw wastewater contains a wide range of pollutants (103-1010 Da). The results showed that the EO process is not a pollutant-selective technique in terms of molecular weight and it shows its superiority over other processes.
Research Article
Treatment of Embroidery Wastewater Containing Poly (Vinyl Alcohol) by Electrooxidation Process
https://doi.org/10.21203/rs.3.rs-2252846/v1
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Poly (Vinyl Alcohol) Wastewater
Electrooxidation
Boren Doped Diamond Electrode
Embroidery is the art of creating decorative designs on fabric, leather, or paper. It is also known as chemical or Schiffli lace. In this lace-making method, a pattern is formed on a sacrificial fabric. Then that sacrificial fabric (usually poly(vinyl) alcohol; PVA) is chemically dissolved. In the world, a total of 930 kilotons of PVA is produced every year in more than 20 countries and regions [1]. When PVA is selected as the sacrificial fabric, the chemical solvent is water which causes forming of a high amount of wastewater. Due to the embroidery textile industry containing a series of processes (finishing, bleaching, mercerization, etc.), this wastewater contains a high concentration of organic and colored substances except for PVA. Unless this wastewater is subjected to a treatment process that provides high removal efficiency, it creates a huge foam layer on the surface in the aquatic environments where it is discharged. These foams prevent entering sunlight into the aquatic environment and therefore photosynthesis. In the next stage, the oxygen concentration in the water decreases, and anaerobic conditions occur. In addition, the presence of PVA causes heavy metals to accumulate in the aquatic environment [2].
In recent years, many treatment processes have been tried in the treatment of PVA[3, 4]. One of them is the biological treatment systems that come to the forefront with their low cost in the treatment of other wastewater. However, the ratio of BOD to COD of PVA is quite low (BOD5/COD < 1). This shows that a very high amount of oxidation process is required for the degradation of PVA, but this degradation does not occur with conventional biological systems. In a study using mixed cultures acclimated in PVA solution, only about 40% of the PVA was found to be mineralized after 48 days of incubation. In these processes, where 800–1000 m3 of wastewater is produced daily, it is quite long for the treatment time of 48 hours. In the treatment of PVA with biological methods, the foam formed on the equipment makes it difficult to obtain operating stability and acceptable treatment efficiencies [5–7]. High treatment efficiencies are obtained with the salting method, however, this process requires very high chemical consumption [8]. PVA is neutrally charged, highly hydrophilic, and contains a large amount of -OH groups. This polymer does not respond to the coagulation process due to a neutral charge. [9]
The advanced oxidation processes (AOPs) have a high potential for treating PVA wastewater and improving its biodegradability [10], concluding the wet oxidation process [11], the Fenton process [1], a persulfate-Fe system [12], Peroxymonosulfate [13], and ionizing radiation [14]. However, due to some disadvantages and gaps in the literature, more studies are needed on the treatment of PVA-containing wastewater with AOPs. For instance, the Fenton method alone cannot effectively mineralize high concentrations of PVA; it is particularly promising concerning pretreatment [9, 15]. The long PVA molecule chains were broken up, and the abundant PVA’s hydroxyl groups were partially oxidized to -COOH or -C = O groups during the Fenton process. In this way, the biodegradability of the wastewater increases, or it becomes responsive to coagulation through its charged groups.
A method that can treat this wastewater with less chemical and power consumption is an important area of interest in the scientific community [9]. At this point, the electrooxidation (EO) process emerges as a very important alternative as it performs the full mineralization of pollutants in a low oxidation time and without the need for high consumption of chemicals. The EO process achieves the treatment of the pollutants by two different methods which are classified as direct and indirect oxidation. In indirect oxidation, oxidizing reagents such as chlorine, hypochlorous acid, hypochlorite, hydrogen peroxide, and ozone are formed on the electrode surface and these reagents oxidize pollutants. On the other hand, electron transfer from the anode surface to pollutants and/or the formation of hydroxyl radicals is possible during direct oxidation [16, 17]. The factor that determines the success of electrooxidation is the chosen electrode. A wide variety of anode electrodes have been used in both direct and indirect oxidation, but Ti/IrO2-RuO2, SnO2, PbO2, and boron-doped diamond (BDD) electrodes have come to the fore with their high efficiencies [18].
Since the BDD electrode is an inactive anode, it is very effective in the treatment of wastewater. The formation of hydroxyl radicals, one of the highest oxidizing reagents, on the BDD surface has been reported as the main removal mechanism [19, 20]. BDD has an inert surface, and this inert surface causes low physisorption of hydroxyl radicals. Complete mineralization of pollutants in wastewater can be possible due to these physisorbed hydroxyl radicals (Reaction 1 and 2, R: organic pollutants in wastewater) [21–27].
In this study, pollutant removal in wastewater containing PVA was tested using BDD electrodes. For this purpose, the most suitable pH, temperature, current, electrolyte concentration, and operational time in the electrooxidation process were determined.
2.1. Wastewater and The Experimental Set-up
The wastewater was obtained at the entrance of an embroidery plant treatment system. In this plant, temperature, the pH, conductivity, turbidity, and TOC were measured as 60 ̊C, 4.76, 804 µS/cm, 46.9 NTU, and 6584 mg/L, respectively. Since the conductivity value of the raw wastewater is very low, the conductivity was increased by adding NaCl to the wastewater in the initial trials of the study. In addition, since the temperature of the raw wastewater from the production process is high, the effect of temperature in the EO process was investigated.
In the EO process, experiments were carried out in a double-walled reactor with a volume of 600 mL. 15 cm x 6 cm x 0.2 cm BDD by DiaCCon GmbH was used as the anode, and a stainless-steel plate was used as the cathode. For a sample volume of 500 mL, the active surface area of a single plate is 60 cm2. The electrodes are placed vertically and parallel to each other and the distance between the plates was fixed at 2 cm. The electrodes were connected to a digital DC power supply (Agilent 6675A model; 30 V, 6A) operated at galvanostatic mode and the current was held constant at desired values for each run.
During EO, electrical energy consumption (ENC, kWh) was calculated with Eq. 3.:
$$ENC=U\times i\times {t}_{EC}$$
3
where U is cell voltage (V), I is current (A), t is operating time (s) [28].TOC levels were determined by combustion of the samples at 680°C using a non-dispersive IR source (Shimadzu, TOC-L model)[29, 30]. The analyses were triplicated and the average data were reported. The accuracy of these measured values for TOC was calculated at around 1.0%. Hach 2100Q turbidimeter for measuring the turbidity, Varian Agilent Cary Clipce instrument for obtaining fluorescence spectra, and PG instruments T80 + UV/VIS spectrophotometer for UV/VIS spectrum were used.
The molecular weight distribution (MWD) of the pollutants in the wastewater was determined using a high-performance liquid chromatography instrument (Hewlett-Packard HPLC 1260 series) with a refractive index detector (RID). In this system, two ultrahydrogel columns [PL aquagel-OH 30 (100 − 60,000 MW) and PL aquagel-OH MIXED-H (6,000–10,000,000 MW)] and polysaccharide (Pollulan) standards with different molecular weights (186, 100, 58, 48, 23.7 and 12.2 kDa) were used. Ultrapure water was chosen as the mobile phase and used at a flow rate of 1 mL/min. 100 µL of sample is injected into the system [31–33].
3.1. The effect of operational parameters on removal efficiencies
Initial pH effect
There is a lot of powerful data that pH affects the efficiency of treatment technologies [33–36]. However, the efficacy of pH in treatment technologies varies according to the system or wastewater type. For this reason, the effect of pH on treatment efficiency was examined at the beginning of the study. Therefore, the initial pH of the wastewater was adjusted to between 3.0 and 8.0, then the removal efficiency was monitored for 120 min. at 3 A.
Similar to our previous study [35, 37, 38], as can be seen from the results of this study, very small differences were observed in the TOC removal after 120 minutes at different initial pHs (Fig. 3.1.). These results show that the BDD electrode is not pH sensitive in this case and it shows its superiority over other processes that require pH adjustment [35, 39]. It was observed that the TOC amount could be reduced to a maximum of 5025 mg/L at pH 5.0 which is very close to the original pH value of wastewater. Thus, the optimum pH value was selected as pH 5.0. On the other hand, during these experiments, the energy consumption (EC) was monitored and after 120 min the lowest EC was obtained at pH 5.0 as 0.1699 kWh (Fig. 3.1.b). When the pH value was measured after experiments, it was observed that during the EO process with BDD, the final pH always tended to shift to the neutral zone, regardless of the starting pH (Fig. 3.1.c).
Initial electrolyte effect
It has been observed in previous studies that the type and concentration of electrolyte used in the EO process play an active role in color, COD, and TOC removal, and current efficiency [35, 40]. It was decided to use NaCl as the electrolyte since it showed higher removal efficiency than other electrolytes [41]. In the EO process, the removal of pollutants usually occurs with the formation of hydroxyl radicals. Simultaneously, the presence of catalytic components such as chlorine in wastewater leads to the production of active chlorine as shown below by electro-generation. This causes the active chlorine to be a role in the disintegration of pollutants and is called indirect EO [41].
$${Cl}^{-}\to {Cl}_{2}+2{e}^{-}$$
4
$${Cl}^{-}+{H}_{2}O\to {ClO}^{-}+{H}^{+}+{e}^{-}$$
5
$$Organics+ {ClO}^{-}\to Intermediates \to {CO}_{2}+{Cl}^{-}+{H}_{2}O$$
6
When the concentration of NaCl electrolyte varied between 3.0–8.0 g/L, the lowest TOC value was obtained as 4779 mg/L at 3.0 g/L NaCl concentration. It was found that increasing the NaCl concentration did not make increase the removal efficiency. With the addition of electrolytes, a decrease in efficiency was observed due to the increased ion concentration. This has also been observed in similar studies. Neto and Andrade reported a decrease in oxidation efficiency when the NaCl concentration exceeded 3 g/L. They considered that the increase in Cl2 evolution is enhanced towards forming more oxidizable species \((\text{H}\text{C}\text{l}\text{O}, {\text{C}\text{l}\text{O}}^{-})\). Also, they reported another possibility as the formation of chlorate (\({\text{C}\text{l}\text{O}}_{3}^{-}\)), which is favored at high current density and high chloride concentration [42]. On the other hand, the most important advantage of electrolyte production is that it reduces energy consumption as it increases the conductivity of wastewater. In this study, when the concentration varied to 3.0, 4.0, 5.0, 6.0 and 8.0 g/L, the energy consumption varied as 0.17102, 0.11257, 0.12064, 0.09647 and 0.07717 kWh. Although the addition of more electrolytes reduces electricity consumption, it also means an increase in chemical consumption and cost. In addition, a positive effect of electrolyte addition on treatment efficiency was not observed. Therefore, the study was continued with a concentration of 3.0 g/L NaCl. In addition, the experimental results in Fig. 3.2. showed that the time increase contributed positively to the removal efficiency at all NaCl concentrations.
Initial Temperature Effect
Until this stage of the study, the experiments were carried out at 25°C. To understand the effect of temperature on pollutant removal, the initial temperature was set to 25°C, 35°C, and 45°C and kept constant there until the end of the experiment. The experiments were realized at 500 A/m2 (3.0 A), 3g/L of NaCl, and an initial pH of 5.0. As a result of these studies (Fig. 3.3.), it is observed that there is no increase in TOC removal at temperatures above 25°C, and TOC was reduced to 4777 mg/L at 25°C, resulting in 19.48% TOC removal. The effect of temperature change on EC was also investigated as seen in Fig. 3.3.b. Since the temperature increases, the movement of the molecules in the wastewater increases too and thus the conductivity has increased. As a result, energy consumption has decreased. In this part of the study, it is seen that the increased operating time increases the removal efficiency.
Current Effect
One of the most important parameters in EO is current. In the studies conducted in the past years, it has been observed that the current has a direct effect on the removal of COD and TOC [39–41, 43]. To examine the effect of current, it was decided to conduct experiments at 3.0A (500 A/m2), 4.5A (750 A/m2), and 6.0A (1000 A/m2) by adjusting the temperature to 25 oC. The amount of TOC was reduced up to 4782 mg/L in 3.0 A, up to 4610 mg/L in 4.5 A, and up to 3867 mg/L in 6.0 A. It is seen that time has a serious effect on the removal efficiency in all currents applied. The energy consumptions were calculated as 0.17102, 0.19308, and 0.22475 kWh for 3.0, 4.5, and 6.0 A, respectively. It has been determined that the increase in energy consumption with the increase in current is not linear.
The effect of time
A positive effect of time on removal efficiency has been observed in all parts of the study carried out so far. Therefore, it was decided to experiment with 6.0 A (1000 A/m2), 3.0 g/L NaCl and an initial pH of 5.0 for 600 min. hours.
It was observed that the amount of TOC decreased to 159.10 mg/L at the end of 600 min. and the removal efficiencies were realized as 97.2%. During 600 min, the energy consumption was calculated as 1.193 kWh. According to Faraday's law, the number of electrons transferred in EO systems will increase as the current and time increase. For this reason, it is expected that the treatment efficiency will increase as the current and time increase. However, the limiting factor is the number of pollutants or mediators diffused on the electrode surface. At this point, increasing the electrode surface or time minimizes this limiting factor.
3.2. Characterization
The measurements in Fluorescence Spectrophotometer and UV/VIS are made in 10 times diluted wastewater. As a result of both measurements, a significant decrease in the axis of y during the treatment process is observed. (Fig. 3.6).
The fluorescence spectrum of pure PVA showed fluorescence peaks at 370 nm and 500 nm at room temperature in Fig. 3.6a. According to the PVA fluorescence data in the literature, it is possible to give a weak sudden fluorescence peak in the range of 400 to 510 nm [44, 45]. As seen in each figure, it was determined that the peaks thought to belong to the existing PVA after 600 minutes of EO disappeared and it supported the other characterization results. UV-Vis results are supported by other characterization data that PVA in Embroidery wastewater degrades after EO process. The absorbance of pure PVA in UV-Vis is between 200–400 nm. The maximum absorbance is in the UV region and a significant absorbance could not be measured in Vis region [46–48].
Structural analysis of the Embroidery process raw and treated wastewater was performed with the FT-IR spectrum. It was carried out from solid samples obtained by evaporation of 5 mL aqueous solutions taken from these samples in Petri dishes. FT-IR spectra were given in Fig. 3.7. The company that produces this wastewater stated that water-soluble interlining is used in embroidery production. The raw material of the interlining is PVA. In this case, it has been determined that a large proportion of the wastewater generated was PVA. It is seen in many publications in the literature that the FT-IR spectrum of the Embroidery process wastewater (0 min.) is the same as the pure PVA spectrum [49–51]. Absorption peaks of wastewater were found at-OH stretching at -3290 cm− 1 and low-intensity asymmetric stretching of -CH2 peaks at 2935 cm− 1 [50–52]. These two peaks were not seen in the FT-IR spectrum of treated wastewater. This indicates that -H is separated from the -OH group attached to the -CH group in the structure. The bending peak of the 1082 cm− 1 -OH group is in the form of a large peak, and it is known that this peak becomes a severe peak at 1091 cm− 1 after EO and this peak is known as the -CO stretching peak [49, 53]. The -CH2 peak at 918 cm− 1 and the C-C tension peak at 830 cm− 1 disappeared after the process and gave the peak of 1,2 disubstituted -C-H bending structure at 880 cm− 1. When we consider both spectra, it is predicted that the high amount of PVA in the embroidery process wastewater is degraded and the possible halogens (peaks at 655 − 615 cm− 1) are attached to the structure in where the polymer structure has deteriorated.
The shown FT-IR spectrums were obtained after evaporation of the water and other volatile components at 80 0C. As shown in Fig, after 600 min of the EO, the peak at 1100 (1/cm) in the spectrum became a sharp peak, in addition, two new peaks were observed in the 610 and 1400.
Wastewater consists of pollutants with a wide range of molecular weights, depending on the process they are produced, and the raw materials used. Monitoring the molecular weight distributions of these pollutants provides important information about the treatment mechanism. For this purpose, molecular weight distributions of pollutants in wastewater are performed using the HPSEC technique. In the HPSEC technique, the wastewater is filtered through a 0.45 µm filter paper and loaded into the HPLC device. Therefore, this technique monitors dissolved pollutants in wastewater. In Fig. 3.8, the molecular weight distribution of wastewater containing PVA, which is monitored at regular intervals throughout the EO process, is shown.
As seen in Fig. 3.8, raw wastewater contains a wide range of pollutants (103-1010 Da) of these, components from 109 to 1010 are of low density and immediately revert to smaller molecular weight components within the first 30 minutes of the purification process. In parallel, there is a decrease in the number of pollutants with molecular weights between 103 and 106. These two results show that the EO process is not a pollutant-selective technique in terms of molecular weight. If there was a process that primarily converts large-molecule components to small-molecule components, the TOC change should not be linear. However, as can be seen in the Figure (the part shown with the red line), there is an accumulation of pollutants with molecular weights of 104 to 105 as the treatment time increases. This means that some intermediate species resistant to the EO process are formed from high molecular components. However, TOC results finally show us that the EO process successfully decomposes these intermediate species.
In this study, the treatment of real wastewater from the Embroidery industry using BDD (Boron Doped Diamond) was tried, and NaCl was used to increase the conductivity of the wastewater. In the process, the most appropriate pH, temperature, current, electrolyte concentration, and process time parameters were determined. The amount of TOC in the wastewater was increased from 6584.0 mg/L to 159.1 mg/L at the initial pH of the wastewater (5.01), 6.0 A current, 3.0 g/L NaCl and 25°C for 10 hours.
In addition, the removal mechanism of the process was elucidated by using HPSEC, FT-IR, UV-VIS, and Fluorescence Spectroscopy methods. Characterization studies have shown that the main pollutant in wastewater is PVA and the high amount of PVA in the embroidery process wastewater is degraded and the halogens have attached to the part of the polymer where the structure has deteriorated. The results show that the EO process is not a pollutant-selective technique in terms of molecular weight.
Acknowledge
We thank the Scientific and Technical Research Council of Turkey (TUBITAK, Project Number: 115Y305) for funding the research project.
Author contribution
Erhan Gengec: Conceptualization, Methodology, Data Curation, Writing- Review&Editing.: Seda Karayunlu Bozbas: Conceptualization, Investigation, Supervision, Writing- Review&Editing.: Hasan Huseyin: Conceptualization, Investigation, Data Curation.:
Data availability The datasets generated during and/or analyzed during the current study are available from the corresponding author upon reasonable request.
Conflict of interest The authors declare no competing interests.
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No competing interests reported.