The non textile dyes exhibited remarkable decolorization in presence of both the Redox mediators at 1.0 mM concentration. There was no significant change on dye decolorization of non textile dyes at concentrations above 1.0 mM. It has already been reported that Redox mediators have the potential to mediate oxidation reaction between a substrate and an enzyme (d'Acunzo et al., 2006). Different Redox mediators have different mediation efficiency which is governed by Redox potential of the mediator and the oxidation mechanism of the substrate (Baiocco et al., 2003). Oxidation of substrate occurs by free radical formation by the mediator. The free radicals can be formed either by one-electron oxidation of substrate or by abstraction of a proton from the substrate (Xu et al., 2001; Fabbrini et al., 2002). In this study, Redox-mediating property of two different compounds as peroxidase mediators was evaluated and extensive study on textile, non-textile and dye

Laccase have been used with Redox mediators to oxidize non-phenolic compounds (Bourbonnais & Paice, 1990). The mechanism of action of laccase mediator system has been extensively studied and it is used in the textile industry in the finishing process for indigo stained materials. Several workers have demonstrated that the use of Redox mediator system enhanced the rate of dye decolorization by several folds but these mediators were required in very high concentrations (Soares et al., 2001a,b; Claus et al., 2002). In the present study we have shown the decolorization of both textile and non-textile dyes as well as dye mixtures mediated by *T. dioica* peroxidase under the influence of low concentration of Redox mediators. The peroxidase reacted well to decolorize at concentration of 0.8 mM and 1.0 mM of H2O2 for disperse and reactive dyes respectively and remained substantially unaffected till 1.2 mM **[Figure-8]**. Although the concentration of H2O2 greater than 0.75 mM acted as an inhibitor of peroxidase activity by irreversibly oxidizing the enzyme ferri-heme group essential for peroxidase activity (Satar and Husain, 2009), in this study we observed a relatively higher working concentration of H2O2 . The enzyme in the presence of Redox mediators was much effective in performing the decolorization of the dyes, implying dye decolorization was a result of Redox mediated H2O2-dependent enzymatic reaction. Our results are consistent and very near to values reported earlier for maximum functional concentration of H2O2 (Camarero et al., 2005).The enzyme could function better in an acidic medium of pH range 3-6, whereas its decolorizing/degrading activity was adversely affected in alkaline medium **[Figure-9]**. This finding supports the earlier view of disperse dye decolorization in acidic medium by *T. dioica* and reports further that acid medium favours catalysis of reactive dyes too by this peroxidase. It has earlier been reported that the degradation of industrially important dyes by enzymes such as horse radish peroxidase, polyphenol oxidase, BGP and laccase was also maximum in

presence of relatively higher concentration of vanillin.

mixtures has been performed with HOBT.

the buffers of acidic pH (Galindo et al., 2000).

The reaction temperature is an important parameter which effects the decolorization of dyes. The maximum decolorization for both the reactive and disperse dyes were in the temperature range of 40°C to 50°C in the presence of fixed concentration of HOBT **[Figure-10]**. The extent of decolorization was remarkable for Reactive Blue 15, Reactive Orange 15, Reactive Red 4 at 50°C whereas disperse dyes could be decolorized in the range of 61.2% to 79% at 40°C. The decolorization varies with the nature of the dyes but Redox mediated decolorization with *T. dioica* peroxidase was a better solution for effective decolorization of recalcitrant compounds. The rate of decolorization varied with time and maximum decolorization in the presence of HOBT was observed within one hour of incubation for Reactive Blue15 and Disperse Red 19 **[Figure-11]**. However, for other dyes the extent of decolorization progressed slowly and reached a plateau after 4 h of incubation. This data is consistent with reports that decolorization rate varies, depending upon the type of dye to be treated (Camarero et al., 2005).

The non-textile dyes studied for decolorization with *T. dioica* peroxidase exhibited enhanced decolorization in the presence of 1.0 mM HOBT, whereas in the absence of Redox mediator decolorization was much slower. Coomassie Brilliant Blue R250, Methylene Blue, Eriochrome Black T, Martius Yellow, Methyl Orange, Naphthol Blue Black and Rhodamine were extensively decolorized by *T. dioica* peroxidase under the influence HOBT **[Table-2]**.The decolorization of Evans blue was not significantly affected by the presence of Redox mediator, although in the presence of HOBT decolorization achieved was higher. The performance of this system was maximal during 160 min of incubation. Celestine blue was decolorized to 79.5% in the presence of HOBT at higher concentration of enzyme as well as longer incubation time. The data in **[Table-2]** is suggestive of *T. dioica* peroxidase in conjunction with low concentration of HOBT to be wonderful decolorization/degradation system for non-textile dyes as well. Further, dye mixtures simulating industrial effluents exhibited more than 82% decolorization with 1-hydroxybenzotriazole **[Figure-12]**. The rate of decolorization of dye mixture was slower in comparison to that of individual dyes both in the presence and absence of HOBT. The application of inexpensive peroxidase from easily available source can overcome the limitations in current wastewater treatment strategies. The use of peroxidases can be extended to large-scale treatment of wide spectrum of structural dyes by using immobilized PGP along with relatively cheaper Redox mediators. Thus, the study demonstrated that the peroxidase enzyme isolated from *T. dioica* can be coupled with Redox mediator into a system that can serve as an effective biocatalyst for the treatment of effluents containing recalcitrant dyes from textile, dyeing and printing industries.

#### **3. Future perspective**

Dye wastewater discharged from textile and dyestuff industries needs to be treated due to their impact on water bodies and to address growing public concern over their toxicity and carcinogenicity. Many different and complicated molecular structures of dyes make dye wastewater treatment difficult by conventional biological and physico-chemical processes. Therefore innovative treatment technologies need to be investigated. The studies performed using peroxidases from different sources indicates that novel enzyme systems can be created to decolorize wide spectrum of textile, non-textile dyes and dye mixtures under varying set of conditions. The efficacy of decolorization drastically improves with Redox mediators and dyes/dye mixtures recalcitrant to peroxidase exhibited remarkable decolorization. The application of inexpensive peroxidases from easily available sources can overcome the limitations in current wastewater treatment strategies. The use of peroxidases can be extended to large-scale treatment of wide spectrum of structural dyes by using immobilized peroxidases along with relatively cheaper Redox mediators.

#### **4. References**

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**4. References**

## **Effect of Photochemical Advanced Oxidation Processes on the Bioamenability and Acute Toxicity of an Anionic Textile Surfactant and a Textile Dye Precursor**

Idil Arslan-Alaton and Tugba Olmez-Hanci *Istanbul Technical University Turkey*

#### **1. Introduction**

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Surfactants are frequently being used as cleaning, dissolving and wetting agents in household activities and several industries (Utsunomiya et al., 1997; Van de Plassche et al., 1999). For instance in various textile preparation operations (scouring, mercerising, bleaching), surfactant formulations are employed in order to allow thorough wetting of the textile material, emulsification of lipophilic impurities and dispersion of insoluble matter and degradation products (Arslan-Alaton et al., 2007; EU, 2003). Anionic (including alkyl sulphonates, alkyl aryl sulphonates, alkyl sulphates, dialkylsulphosuccinates, and others) and especially nonionic surfactants are the chemicals being more often used for this particular purpose. These multi-purpose surfactants create the main organic pollution load in effluents originating from the above mentioned activities. Surfactants enter the environment mainly through the discharge of sewage effluents into natural water and the application sewage sludge on land for soil fertilizing purposes (Petrovic & Barceló, 2004). Although most of these surfactants are designated as "biodegradable" according to different long-term biodegradability tests (Euratex, 2000) former studies dealing with the biodegradability of surfactants have indicated that their biodegradation in conventional activated sludge treatment systems is in most cases rather incomplete, resulting in the accumulation of partial biodegradation products (Ikehata & El-Din, 2004; Swisher, 1987). A lot of commercial surfactants used today by different industries tend to sorb and hence accumulate on sludge and soil sediments in receiving water bodies due to their amphiphilic characteristics (Swisher, 1987; Staples et al., 2001). Hence, when a surfactant is discharged into the environment without proper treatment at source (e.g. at the treatment facilities of the textile factory), it will enter the sewage treatment works or directly natural waters without any significant structural alteration and may cause serious ecotoxicological consequences due to its bioaccumulation tendency (Ikehata & El-Din, 2004). Consequently, more effective and at the same time economically feasible treatment processes have to be applied to alleviate the problem of partially adsorbed and/or metabolized surfactants in the environment. Moreover, the management of biologically-difficult-to-degrade effluent discharges bearing surfactants remains an important challenge that has to be urgently solved to reduce the concentration of surfactants in effluent discharge at source.

Aryl sulphonates are commercially important textile dye precursers; their global production for dye synthesis is limited to a few countries (e.g. China, India, Pakistan and Egypt) where they create serious environmental problems since they are biologically inert in natural as well as engineered aerobic biotreatment systems. Moreover, due to their very polar and hydrophilic nature, sulphonated textile dye precursers are very mobile in the aquatic ecosystem and hence very difficult to treat via conventional methods (i.e. coagulation, precipitation, adsorption). In addition, it has been postulated that the anaerobic metabolites of sulphonates bearing aromatic amines in their molecular structure are potentially mutagenic and even carcinogenic. From the environmental point of view, the fate of aryl sulphonates and their degradation products in biological treatment units as well as receiving water bodies is still not very clear since until now only limited attention has been paid towards their occurrence, fate and degradability in engineered systems as well as in the natural environment (Jandera et al., 2001). Due to the fact that conventional biological, physical and chemical treatment methods are not very effective in the removal of naphthalene sulphonates, alternative options have to be considered and evaluated.

In the last three decades, so-called Advanced Oxidation Processes (AOPs) have been developed and explored for the treatment of toxic and/or refractory pollutants and wastewaters. Their efficiency and oxidation power relies on the formation of very active species including hydroxyl radicals (HO●) that violently and nonselectively react with most organic pollutants at diffusion limited rates. Bimolecular rate coefficients obtained for the free radical chain reaction initiated by HO● are typically in the range of 107-1011 M-1 s-1 for most pollutants depending on their structural properties. Among the AOPs, photochemical processs including the UV-C photolysis of H2O2 and the Photo-Fenton reaction (UV-C photolysis of H2O2 catalyzed by Fe2+ ions under acidic pH conditions) are known as the most efficient, feasible and kinetically favorable types. The major drawbacks of AOPs can be listed as (i) their relatively high electrical energy requirements dramatically affecting their operating costs and hence real-scale application as well as (ii) the possibility of undesired oxidation products (more toxic and less bioamenable than the mother pollutant) formation and build-up due to kinetically difficult to control chain reactions in the treatment system. Baxendale and Wilson (1957) first studied the UV-C photolysis of H2O2 in water that is the most direct method of HO generation through the homolytic cleavage of H2O2. The following set of reactions explains the main steps of the mechanism of the H2O2 photolytic decomposition in water;

$$\text{H}\_2\text{O}\_2 + \text{hv} \to 2\text{HO}^\* \tag{1}$$

$$\text{HO}^{\bullet} + \text{H}\_{2}\text{O}\_{2} \rightarrow \text{HO}\_{2}^{\bullet} + \text{H}\_{2}\text{O} \tag{2}$$

$$\text{HO}^{\bullet} + \text{HO}\_{2}^{\bullet} \rightarrow \text{HO}\_{2}^{\bullet} + \text{OH}^{\bullet} \tag{3}$$

$$2\text{HO}^\* \rightarrow \text{H}\_2\text{O} + \frac{1}{2}\text{O}\_2\tag{4}$$

The quantum yield of reaction (4) is 1.0 mole HO per mole of photons (1.0 einstein) absorbed by H2O2 (Prousek, 1996). The molar absorption coefficient of H2O2 at 254 nm is only 18.6 M-1 cm-1 and hence the efficiency of the H2O2/UV-C process decreases drastically with increasing pollutant concentration and UV254 absorbance of the target chemical (CCOT 1995). For heavily polluted effluents, high UV-C doses and/or H2O2 concentrations are required.

Aryl sulphonates are commercially important textile dye precursers; their global production for dye synthesis is limited to a few countries (e.g. China, India, Pakistan and Egypt) where they create serious environmental problems since they are biologically inert in natural as well as engineered aerobic biotreatment systems. Moreover, due to their very polar and hydrophilic nature, sulphonated textile dye precursers are very mobile in the aquatic ecosystem and hence very difficult to treat via conventional methods (i.e. coagulation, precipitation, adsorption). In addition, it has been postulated that the anaerobic metabolites of sulphonates bearing aromatic amines in their molecular structure are potentially mutagenic and even carcinogenic. From the environmental point of view, the fate of aryl sulphonates and their degradation products in biological treatment units as well as receiving water bodies is still not very clear since until now only limited attention has been paid towards their occurrence, fate and degradability in engineered systems as well as in the natural environment (Jandera et al., 2001). Due to the fact that conventional biological, physical and chemical treatment methods are not very effective in the removal of

naphthalene sulphonates, alternative options have to be considered and evaluated.

most direct method of HO

HO

HO

2HO <sup>→</sup> H2O + <sup>1</sup>

The quantum yield of reaction (4) is 1.0 mole HO

decomposition in water;

required.

In the last three decades, so-called Advanced Oxidation Processes (AOPs) have been developed and explored for the treatment of toxic and/or refractory pollutants and wastewaters. Their efficiency and oxidation power relies on the formation of very active species including hydroxyl radicals (HO●) that violently and nonselectively react with most organic pollutants at diffusion limited rates. Bimolecular rate coefficients obtained for the free radical chain reaction initiated by HO● are typically in the range of 107-1011 M-1 s-1 for most pollutants depending on their structural properties. Among the AOPs, photochemical processs including the UV-C photolysis of H2O2 and the Photo-Fenton reaction (UV-C photolysis of H2O2 catalyzed by Fe2+ ions under acidic pH conditions) are known as the most efficient, feasible and kinetically favorable types. The major drawbacks of AOPs can be listed as (i) their relatively high electrical energy requirements dramatically affecting their operating costs and hence real-scale application as well as (ii) the possibility of undesired oxidation products (more toxic and less bioamenable than the mother pollutant) formation and build-up due to kinetically difficult to control chain reactions in the treatment system. Baxendale and Wilson (1957) first studied the UV-C photolysis of H2O2 in water that is the

following set of reactions explains the main steps of the mechanism of the H2O2 photolytic

H2O2 + hv <sup>→</sup> <sup>2</sup>HO (1)

absorbed by H2O2 (Prousek, 1996). The molar absorption coefficient of H2O2 at 254 nm is only 18.6 M-1 cm-1 and hence the efficiency of the H2O2/UV-C process decreases drastically with increasing pollutant concentration and UV254 absorbance of the target chemical (CCOT 1995). For heavily polluted effluents, high UV-C doses and/or H2O2 concentrations are

+ H2O2 → HO2

+ HO2

generation through the homolytic cleavage of H2O2. The

+ H2O (2)

<sup>2</sup> O2 (4)

per mole of photons (1.0 einstein)

(3)

The Fenton process is being increasingly used in the treatment of organic pollutants. The conventional dark Fenton process involves the use of one or more oxidizing agents (usually H2O2 and/or oxygen) and a catalyst (a metal salt or oxide, usually iron) at acidic pHs. Fenton process is efficient only in the pH range 2-4 and is usually most efficient at around pH 2.8 (Pignatello, 1992). In acidic water medium, the most accepted scheme of this radical mechanism is described in the following equations (Sychev & Isak, 1995):

$$\text{Fe}^{2+} + \text{H}\_2\text{O}\_2 \rightarrow \text{Fe}^{3+} + \text{HO}^\cdot + \text{HO}^\bullet \tag{5}$$

$$\text{Fe}^{3+} + \text{H}\_2\text{O}\_2 \rightarrow \text{Fe}^{2+} + \text{HO}\_2^{\bullet} + \text{H}^+ \tag{6}$$

$$\text{Fe}^{2+} + \text{HO}^{\bullet} \rightarrow \text{Fe}^{3+} + \text{HO}^{\bullet} \tag{7}$$

$$\text{HO}^{\bullet}\text{+}\text{H}\_{2}\text{O}\_{2}\rightarrow\text{HO}\_{2}^{\bullet}\text{+}\text{H}\_{2}\text{O}\tag{8}$$

$$\text{HO}\_2^{\bullet} \leftrightarrow \text{O}\_2^{\bullet^\*} \text{+ H}^+ \quad \text{K} = 1.58 \times 10^{-5} \text{ M} \tag{9}$$

$$\text{Fe}^{3+} + \text{HO}\_2^{\bullet}/\text{O}\_2^{\bullet \text{-}} \rightarrow \text{Fe}^{2+} + \text{O}\_2/\text{+H}^+\tag{10}$$

$$\text{Fe}^{2+} + \text{HO}\_2^{\bullet}/\text{O}\_2^{\bullet^{\bullet}} \rightarrow \text{Fe}^{3+} + \text{H}\_2\text{O}\_2\tag{11}$$

Photo-Fenton oxidation is the photo-catalytically enhanced version of the Fenton process; in the Photo-Fenton process; UV light irradiation (180-400 nm) causes an increase in the HO• formation rate and efficiency via photoreduction of Fe3+ to Fe2+, thus continuing the redox cycle as long as H2O2 is available (Pignatello, 1992; Wadley & Waite, 2004). The reason for the positive effect of irradiation on the degradation rate include the photoreduction of Fe3+ to Fe2+ ions, which produce new HO with H2O2 (Eq. 6) according to the following mechanism;

$$\text{Fe}^{3+} + \text{H}\_2\text{O} \xrightarrow{\text{hv}} \text{Fe}^{2+} + \text{HO}^\* + \text{H}^+ \tag{12}$$

$$\text{H}\_2\text{O}\_2 \xrightarrow{\text{hu}} 2\text{HO}^\bullet \quad (\lambda \lessapprox 400 \text{ nm}) \tag{13}$$

The main compounds absorbing UV light in the Fenton system are ferric ion complexes, e.g. [Fe3+(OH)-]2+ and [Fe3+(RCO2)-]2+, which produce additional Fe2+ by the following photoinduced, ligand-to-metal charge-transfer reactions (Sagawe et al., 2001);

$$\text{-}\left[\text{Fe}^{3+}\text{(OH)}\right]^{2+} \rightarrow \text{Fe}^{2+} \text{+HO}^{\bullet} \text{ (\less\text{ca. 450 nm})}\tag{14}$$

$$\left[\text{Fe}^{3+}\text{(RCO)}\right]^{2+} \rightarrow \text{Fe}^{2+} + \text{CO}\_2 + \text{R}^{\bullet} \text{ ($$

Additionally, reaction (14) yields HO , while reaction (15) results in a decrease of the total organic carbon (TOC) content of the system due to the decarboxylation of organic acid intermediates. It is very important to note that both reactions produce ferrous ions required for the Fenton reaction (Eq. 6). The overall degradation rate of organic compounds is considerably increased in the Photo-Fenton process, even at lower concentration of iron salts present in the system (Chen & Pignatello, 1997). The main disadvantage of the iron based AOPs is the necessity to work at low pH, because at higher pH ferric ions would begin to precipitate as hydroxide. Furthermore, depending on the iron concentration used, it has to be removed after the treatment in agreement with the regulation established for wastewater discharge (Rodríguez, 2003).

Information about the toxicity of organic pollutants and advanced oxidation intermediates appears to be a key factor in order to evaluate the applicability of the AOPs. It should be considered that advanced oxidation may result in the formation of more inert and/or toxic degradation products, since free radical chain reaction-based treatment methods are difficult to control thus requiring the toxicity assessment of the formed advanced oxidation intermediates (Farré et al., 2007; Oller et al., 2007; González et al., 2007). Furthermore, in most cases the final goal of AOPs is to improve the biocompatibility of the effluent in order to apply a biological treatment (Scott & Ollis, 1995; Marco et al., 1997). Consequently, the use of short-term biological screening tests that aim to a rapid estimation of the acute toxicity of a pollutant and its oxidation products during AOP applications has become essential. Until now, several studies investigating relative changes in the toxicity of model pollutants have been reported using photobacteria, cladocerans, microalgae and other test organisms (Wang et al., 2002; Dalzell et al., 2002; Stasinakis et al., 2008). Among the available acute toxicity test protocols, the assessment of respirometric inhibition (decrease in oxygen uptake rate of microorganisms relative to a control sample; e.g. synthetic or real sewage) is an attractive and advantageous option since heterotrophic biomass being present in biological treatment systems are used as the toxicity test organism that enables the simultaneous assessment of acute toxicity and biodegradability by a single test protocol (Paixão et al., 2002; Gendig et al., 2003; Henriques et al., 2005; Stasinakis et al., 2008; Olmez-Hanci et al., 2009).

Bearing the above-mentioned facts in mind, the present work aimed at investigating the photochemical advanced oxidation of a commercially important anionic textile surfactant (dodecyl sulpho succinate; DOS) and a frequently used vinyl sulphone dye intermediate, namely Para base (aniline-4-beta-ethyl sulphonyl sulphate-2-sulphonic acid; PB) employing H2O2/UV-C and Photo-Fenton treatment processes. Treatment performances of the H2O2/UV-C and Photo-Fenton processes were evaluated in terms of DOS, PB, COD and TOC removals. Treatment efficiencies, pseudo-first-order reaction rate coefficients and associated electrical energy requirements were comparatively evaluated for DOS and PB degradation with the H2O2/UV-C and Photo-Fenton processes to decide for the technically and economically more attractive treatment solution for each chemical. In the second part of the study, activated sludge inhibition experiments were conducted with untreated and photochemically pretreated DOS and PB samples in order to examine the inhibitory effect of the model pollutant and advanced oxidation intermediates on the respirometric activity of activated sludge microorganisms.

#### **2. Materials and methods**

#### **2.1 Chemicals and reagents**

DOS and PB model compounds were supplied by a local textile chemical company and used as received without further purification. Some physicochemical and environmental characteristics of the studied model compounds were summarized in Table 1.

AOPs is the necessity to work at low pH, because at higher pH ferric ions would begin to precipitate as hydroxide. Furthermore, depending on the iron concentration used, it has to be removed after the treatment in agreement with the regulation established for wastewater

Information about the toxicity of organic pollutants and advanced oxidation intermediates appears to be a key factor in order to evaluate the applicability of the AOPs. It should be considered that advanced oxidation may result in the formation of more inert and/or toxic degradation products, since free radical chain reaction-based treatment methods are difficult to control thus requiring the toxicity assessment of the formed advanced oxidation intermediates (Farré et al., 2007; Oller et al., 2007; González et al., 2007). Furthermore, in most cases the final goal of AOPs is to improve the biocompatibility of the effluent in order to apply a biological treatment (Scott & Ollis, 1995; Marco et al., 1997). Consequently, the use of short-term biological screening tests that aim to a rapid estimation of the acute toxicity of a pollutant and its oxidation products during AOP applications has become essential. Until now, several studies investigating relative changes in the toxicity of model pollutants have been reported using photobacteria, cladocerans, microalgae and other test organisms (Wang et al., 2002; Dalzell et al., 2002; Stasinakis et al., 2008). Among the available acute toxicity test protocols, the assessment of respirometric inhibition (decrease in oxygen uptake rate of microorganisms relative to a control sample; e.g. synthetic or real sewage) is an attractive and advantageous option since heterotrophic biomass being present in biological treatment systems are used as the toxicity test organism that enables the simultaneous assessment of acute toxicity and biodegradability by a single test protocol (Paixão et al., 2002; Gendig et al., 2003; Henriques et al., 2005; Stasinakis et al., 2008; Olmez-Hanci et al.,

Bearing the above-mentioned facts in mind, the present work aimed at investigating the photochemical advanced oxidation of a commercially important anionic textile surfactant (dodecyl sulpho succinate; DOS) and a frequently used vinyl sulphone dye intermediate, namely Para base (aniline-4-beta-ethyl sulphonyl sulphate-2-sulphonic acid; PB) employing H2O2/UV-C and Photo-Fenton treatment processes. Treatment performances of the H2O2/UV-C and Photo-Fenton processes were evaluated in terms of DOS, PB, COD and TOC removals. Treatment efficiencies, pseudo-first-order reaction rate coefficients and associated electrical energy requirements were comparatively evaluated for DOS and PB degradation with the H2O2/UV-C and Photo-Fenton processes to decide for the technically and economically more attractive treatment solution for each chemical. In the second part of the study, activated sludge inhibition experiments were conducted with untreated and photochemically pretreated DOS and PB samples in order to examine the inhibitory effect of the model pollutant and advanced oxidation intermediates on the respirometric activity of

DOS and PB model compounds were supplied by a local textile chemical company and used as received without further purification. Some physicochemical and environmental

characteristics of the studied model compounds were summarized in Table 1.

discharge (Rodríguez, 2003).

2009).

activated sludge microorganisms.

**2. Materials and methods 2.1 Chemicals and reagents**

Table 1. Physicochemical properties of the anionic surfactant (DOS) and the vinyl sulphone dye precursor (PB). The COD and TOC equivalencies shown in Table 1 were experimentally obtained by preparing individual calibration curves for varying DOS and PB concentrations in water and determination of the corresponding COD and TOC values.

35% w/w H2O2 (Fluka, USA) was used as received without any dilution. Residual H2O2 was destroyed with enzyme catalase derived from *Micrococcus lysodeikticus* (100181 U mL-1, Fluka, USA). The ferrous iron catalyst source was prepared for daily use by dissolving FeSO4.7H2O (Fluka, USA) in distilled water to obtain a 10% (w/v) stock solution. Several concentrations of HNO3 (Merck, Germany) and NaOH (Merck, Germany) solutions were used for pH adjustment. HPLC-grade acetonitrile (Merck, Germany) was used as the mobile phase in the HPLC measurements. All other reagents were analytical grade and used without purification.

#### **2.2 UV-C photoreactor and light source**

All photochemical treatment experiments were performed using a 3250 mL-capacity batch stainless steel batch reactor (length=84.5 cm; width=8 cm) equipped with a 40W lowpressure, mercury vapor sterilization lamp that was located at the center of the reactor in a quartz glass envelope. The incident light flux of the UV lamp and effective light path length were determined via H2O2 actinometry (Nicole et al., 1990) as 1.6 x 10-5 einstein L-1 s-1 and 5.1 cm, respectively, at 254 nm. During the photochemical experiments, the reaction solution was continuously circulated through the UV-C photoreactor using a peristaltic pump at a rate of 400 mL min-1. For the H2O2/UV-C experiments, H2O2 was added to the pH-adjusted aqueous DOS and PB solutions and the reaction mixture was fed to the photoreactor via a peristaltic pump. After the sample t=0 was taken, the reaction was initiated by turning on the UV-C lamp. In case of Photo-Fenton oxidation experiments, after adjusting the initial pH of the reaction solutions to 3.0 0.1, e.g. the optimum pH for Photo-Fenton oxidation (Oliveros et al., 1997), Fe2+ catalyst was added. In order to eliminate the effect of the dark Fenton reaction, the other reactant (H2O2) was added to the reaction mixture at the very end of the suction period. As the whole solution was fed into photoreactor, UV-C lamp was turned on and photo-oxidation was started. For photochemical oxidation experiments, samples were taken at regular time intervals for analyses up to 180 min.

#### **2.3 Experimental approach**

For all photochemical oxidation experiments aqueous solutions of DOS and PB were individually prepared and adjusted to an initial COD of 450 mg L-1 since a typical textile preparation and textile dye synthesis wastewater exerts a COD in the range of 400-500 mg L-1 (Zollinger, 2001). It is known that an optimum H2O2 concentration exists for the H2O2/UV-C treatment system depending upon the reaction pH, initial organic carbon content, type and molecular structure of the pollutant under investigation (Muruganandham & Swaminathan, 2004; Arslan-Alaton & Erdinc, 2006). In our preliminary experiments the H2O2 concentrations (30 mM) were initially selected close to the stoichiometric oxygen equivalent of the aqueous solutions of DOS and PB (=COD 2.12 g H2O2/g O2 28 mM H2O2). For H2O2/UV-C treatment of DOS at 30 mM H2O2 concentration resulted in significant degradation both in DOS and organic carbon content. However, the theoretically established H2O2 concentration of 30 mM did not result in significant organic carbon removals for PB (data not shown). Considering our preliminary results, it was decided to increase the initial H2O2 concentration to 60 mM for H2O2/UV-C treatment of PB. The initial reaction pH was selected as 10.5 for H2O2/UV-C treatment of aqueous DOS solutions, since the effluent pH of a typical textile preparation is in the range 10–11 (EU, 2003), and in previous studies it was observed that the H2O2/UV-C oxidation process was practically pH-independent, and no significant changes in the reaction rate constants were found over a wide pH range of 3-11 (Arslan-Alaton & Erdinc, 2006; Arslan-Alaton et al., 2007). H2O2/UV-C treatment of aqueous PB solutions was investigated at a reaction pH of 3.0.

For Photo-Fenton experiments, H2O2 (30 mM) and Fe2+ (0.5 mM) concentrations were selected upon consideration of optimum Fe2+: H2O2 molar ratios obtained in former related research work and our own experiences in similar case studies (Arslan-Alaton et al., 2009a; 2009b; 2010). Photo-Fenton experiments were conducted at a fixed pH value of 3.0 0.1 that is generally accepted as the optimum pH for iron-based AOPs (Pignatello, 1992).

As being expected for photochemical AOPs (Oppenländer, 2003), the degradation of the original pollutants (in our case DOS and PB) is usually faster than that of their oxidation products, which are typically expressed in terms of the collective parameters COD and TOC. It was also taken into account that treatment conditions may vary for the removal of original pollutant and its degradation products that are usually collectively presented with the environmental sum parameters COD and TOC. Thus in the present study, the process efficiencies were assessed both in terms of DOS and PB, COD and TOC abatement rates. Residual (unreacted) H2O2 and pH were also recorded during the entire treatment period. The pH and ionic strength of the synthetic solutions were not controlled throughout the experiments. All experiments were conducted at a constant temperature of 20C.

As aforementioned, information about the toxicity of degradation intermediates as well as original pollutants appears to be a key issue for the application of AOPs in water and wastewater treatment as well as integrated chemical and biochemical treatment options (Scott & Ollis, 1995; Farré et al., 2007). By considering this important issue in the last stage of the experimental work, relative changes in the inhibitory effect of DOS and PB photochemical degradation intermediates on heterotrophic biomass (sewage sludge) during the application of photochemical oxidation were also examined.

#### **2.4 Activated sludge inhibition experiments**

The procedure of the activated sludge inhibition test was based on the test protocol given in the ISO 8192 Method (2007). This method describes the way to assess the inhibitory effects of a test substance on oxygen consumption of activated sludge (resembling sewage) by measuring the respiration rate under defined conditions in the presence of a known biodegradable substrate and different concentrations of the test substance (Ubay Cokgor et al., 2007). The activated sludge used as the heterotrophic biomass was obtained from a municipal tertiary wastewater treatment located in Istanbul and daily fed with synthetic sewage in accordance with the same test protocol (ISO 8192, 2007).The heterotrophic biomass concentration in the activated sludge was determined according to Standard Methods (APHA/AWWA/WEF, 2005) and expressed as mg L-1 of mixed liquid volatile suspended solids (MLVSS). All experiments were run at constant heterotrophic active biomass concentration of 1500 mg L-1 to obtain appropriate oxygen uptake rates (OUR) of around 100 mg L-1 h-1.The COD of the synthetic sewage was adjusted to 480 mg L-1 and used as the "readily biodegradable substrate" in the activated sludge inhibition experiments. Any unreacted H2O2 remaining in the reaction samples was destroyed with catalase enzyme and the pH of the solutions was adjusted to 7.0±0.2 before conducting the toxicity tests in order to eliminate potential inhibitory effects of hydrogen peroxide and pH, respectively. The blank control was prepared using the same amount of activated sludge and synthetic sewage solution as in the test dilutions, but without adding untreated or treated aqueous DOS and PB samples. The decrease in dissolved oxygen concentration in the blank control and reaction samples was periodically monitored using a WTW InolabOxi Level 2 oxygen meter at different incubation times (15-180 min). OUR values were calculated on the basis of the linear part of the decreasing dissolved oxygen concentration curves versus test time. Percent inhibition of oxygen consumption (IOUR), for every test sample was calculated using the following equation;

OUR B T <sup>B</sup> I (%) R R 100 R (16)

where RB stands for the OUR in the sample blank and RT is the OUR in the sample effluent mixture, respectively The IOUR values were, thereafter, plotted against the logarithm of the DOS and PB concentrations. The DOS and PB concentrations resulting in percent decrease in OUR's (in mg L-1 DOS or PB) after 15 min incubation period was eventually calculated by interpolation of the "log DOS" or "log PB" versus percent "IOUR" plots obtained for different DOS and PB concentrations.

#### **2.5 Analytical procedures**

78 Advances in Treating Textile Effluent

treatment system depending upon the reaction pH, initial organic carbon content, type and molecular structure of the pollutant under investigation (Muruganandham & Swaminathan, 2004; Arslan-Alaton & Erdinc, 2006). In our preliminary experiments the H2O2 concentrations (30 mM) were initially selected close to the stoichiometric oxygen equivalent of the aqueous solutions of DOS and PB (=COD 2.12 g H2O2/g O2 28 mM H2O2). For H2O2/UV-C treatment of DOS at 30 mM H2O2 concentration resulted in significant degradation both in DOS and organic carbon content. However, the theoretically established H2O2 concentration of 30 mM did not result in significant organic carbon removals for PB (data not shown). Considering our preliminary results, it was decided to increase the initial H2O2 concentration to 60 mM for H2O2/UV-C treatment of PB. The initial reaction pH was selected as 10.5 for H2O2/UV-C treatment of aqueous DOS solutions, since the effluent pH of a typical textile preparation is in the range 10–11 (EU, 2003), and in previous studies it was observed that the H2O2/UV-C oxidation process was practically pH-independent, and no significant changes in the reaction rate constants were found over a wide pH range of 3-11 (Arslan-Alaton & Erdinc, 2006; Arslan-Alaton et al., 2007). H2O2/UV-C treatment of aqueous PB solutions was

For Photo-Fenton experiments, H2O2 (30 mM) and Fe2+ (0.5 mM) concentrations were selected upon consideration of optimum Fe2+: H2O2 molar ratios obtained in former related research work and our own experiences in similar case studies (Arslan-Alaton et al., 2009a; 2009b; 2010). Photo-Fenton experiments were conducted at a fixed pH value of 3.0 0.1 that

As being expected for photochemical AOPs (Oppenländer, 2003), the degradation of the original pollutants (in our case DOS and PB) is usually faster than that of their oxidation products, which are typically expressed in terms of the collective parameters COD and TOC. It was also taken into account that treatment conditions may vary for the removal of original pollutant and its degradation products that are usually collectively presented with the environmental sum parameters COD and TOC. Thus in the present study, the process efficiencies were assessed both in terms of DOS and PB, COD and TOC abatement rates. Residual (unreacted) H2O2 and pH were also recorded during the entire treatment period. The pH and ionic strength of the synthetic solutions were not controlled throughout the

As aforementioned, information about the toxicity of degradation intermediates as well as original pollutants appears to be a key issue for the application of AOPs in water and wastewater treatment as well as integrated chemical and biochemical treatment options (Scott & Ollis, 1995; Farré et al., 2007). By considering this important issue in the last stage of the experimental work, relative changes in the inhibitory effect of DOS and PB photochemical degradation intermediates on heterotrophic biomass (sewage sludge) during

The procedure of the activated sludge inhibition test was based on the test protocol given in the ISO 8192 Method (2007). This method describes the way to assess the inhibitory effects of a test substance on oxygen consumption of activated sludge (resembling sewage) by measuring the respiration rate under defined conditions in the presence of a known biodegradable substrate and different concentrations of the test substance (Ubay Cokgor et al., 2007). The activated sludge used as the heterotrophic biomass was obtained from a municipal tertiary wastewater treatment located in Istanbul and daily fed with synthetic

is generally accepted as the optimum pH for iron-based AOPs (Pignatello, 1992).

experiments. All experiments were conducted at a constant temperature of 20C.

the application of photochemical oxidation were also examined.

**2.4 Activated sludge inhibition experiments**

investigated at a reaction pH of 3.0.

Reaction samples were analyzed for DOS-PB, COD, TOC, residual H2O2 and pH before, during and after photochemical treatment. CODs were determined by the closed reflux titrimetric method according to ISO 6060 (1989). In order to prevent the positive interference of H2O2 to COD analyses, the pH of each sample solution was adjusted to 6.5-7.5 and thereafter catalase enzyme made from *Micrococcus lysodeikticus* (1AU destroys 1 mol H2O2 at pH=7, 100181 U mL-1, Fluka grade) was added to destroy any residual H2O2.

TOC was monitored on a Shimadzu VPCN model carbon analyzer (combustion method) equipped with an autosampler. Residual H2O2 in the samples was determined titrimetically by employing the molibdate-catalyzed iodometric method (Official Methods of Analysis, 1980).

The amount of DOS and PB in the aqueous solutions was measured by high-performance liquid chromatography (HPLC, Agilent 1100 Series, USA). DOS was determined with a fluorescence detector (FLD; λex = 225 nm, λem = 295 nm) and a C8 column. 3 mM NaCl / CH3CN (80/20; v/v) served as the mobile phase (flow rate of 1.5 mL min-1) for DOS. The instrument detection limit for DOS (1.5 mg L-1) was determined as the lowest injected standard that gave a signal-to-noise ratio of at least 3 and an accuracy of 80-95%.

PB abatement was monitored with a diode array detector (DAD; λ = 276 nm) and a Novapack C18 (Waters, USA) column. CH3CN/H2O (60/40; v/v) was used as mobile phase for the analysis of PB at a flow rate of 1 mL min-1. The column temperature was set as 30C for all measurements and the injection volume was selected as 30 µL. The instrument detection limit (1.5 mg L-1) for PB was determined as the lowest injected standard that gave a signal-to-noise ratio of at least 3 and an accuracy of 80-95%. The limit of quantification was calculated as 10 times of the signal-to-noise ratio as 5 mg L-1.

In case of Photo-Fenton Fe2+/H2O2/UV-C experiments, all analytical measurements were performed after filtration of the treated samples though a 0.45 µm membrane filter (Sartorius) in order to remove the settled Fe(OH)3 sludge. The pH was measured with a Thermo Orion model 720 pH-meter at any stage of the experiments.

#### **2.6 Kinetic analysis**

DOS and PB abatements during H2O2/UV-C and Photo-Fenton treatment processes followed pseudo-first order kinetics with respect to the model pollutant (DOS and PB). Hence, the kinetic rates could be expressed as follows;

$$
\ln \mathbb{C} / \mathbb{C}\_0 = -\mathbf{k}\_{\mathbb{C}} \times \mathbf{t} \tag{17}
$$

where kC is the pseudo-first order rate constant (in min-1), t stands for photochemical treatment time (in min), Co and C are the initial and final concentrations of model pollutant (DOS and PB, in mg L-1).

#### **2.7 Electrical energy requirements of the photochemical treatment processes**

The photodegradation of aqueous organic pollutants is an energy-intensive process, thus consideration of figures-of-merit definition based on the electrical energy input is appropriate. In order to compare treatment efficiencies for photochemical oxidation processes, it is important to use a general expression and relate the treatment efficiencies directly to operating costs. For this purpose, the electrical energy per order (EE/O) defined as the kWh electrical energy required to degrade a contaminant or group parameter by one order of magnitude in m3 of contaminant water or wastewater, can be calculated for treatment process (Bolton et al., 2001). Since the EE/O (kWh m-3 order-1) concept assumes pseudo-first order degradation kinetics with respect to pollutant concentration can be determined from the following formula;

$$\text{EE/O (kWh m}^{-3}\text{order}^{-1}) = (38.4 \times \text{P}) / (\text{V} \times \text{k}\_{\odot}) \tag{18}$$

where P is the lamp power (0.04 kW in the present work), V is the reactor volume (3.25 L in the present work), and kC is the pseudo-first-order rate constant of DOS and PB degradation (min−1).

#### **3. Results and discussion**

#### **3.1 Activated sludge inhibition experiments with untreated DOS and PB**

The ISO 8192 (2007) specifies a method for assessing the inhibitory or stimulatory effects of substances, mixtures or wastewaters to activated sludge (Paixão et al., 2002). Information generated by this method may be helpful in estimating the effect of a test material on bacterial communities in the aquatic environment, especially in aerobic biological treatment systems (Paixão et al., 2002). In the first part of this experimental study, untreated (original) aqueous DOS and PB solutions were individually exposed to activated sludge inhibition experiments. The obtained semi-logarithmic IOUR (in %) versus DOS and PB concentrations (log DOS and log PB) were presented in Figure 1 (a) and (b), respectively for an incubation period of 15 min.

for all measurements and the injection volume was selected as 30 µL. The instrument detection limit (1.5 mg L-1) for PB was determined as the lowest injected standard that gave a signal-to-noise ratio of at least 3 and an accuracy of 80-95%. The limit of quantification was

In case of Photo-Fenton Fe2+/H2O2/UV-C experiments, all analytical measurements were performed after filtration of the treated samples though a 0.45 µm membrane filter (Sartorius) in order to remove the settled Fe(OH)3 sludge. The pH was measured with a

DOS and PB abatements during H2O2/UV-C and Photo-Fenton treatment processes followed pseudo-first order kinetics with respect to the model pollutant (DOS and PB).

where kC is the pseudo-first order rate constant (in min-1), t stands for photochemical treatment time (in min), Co and C are the initial and final concentrations of model pollutant

The photodegradation of aqueous organic pollutants is an energy-intensive process, thus consideration of figures-of-merit definition based on the electrical energy input is appropriate. In order to compare treatment efficiencies for photochemical oxidation processes, it is important to use a general expression and relate the treatment efficiencies directly to operating costs. For this purpose, the electrical energy per order (EE/O) defined as the kWh electrical energy required to degrade a contaminant or group parameter by one order of magnitude in m3 of contaminant water or wastewater, can be calculated for treatment process (Bolton et al., 2001). Since the EE/O (kWh m-3 order-1) concept assumes pseudo-first order degradation kinetics with respect to pollutant concentration can be

where P is the lamp power (0.04 kW in the present work), V is the reactor volume (3.25 L in the present work), and kC is the pseudo-first-order rate constant of DOS and PB

The ISO 8192 (2007) specifies a method for assessing the inhibitory or stimulatory effects of substances, mixtures or wastewaters to activated sludge (Paixão et al., 2002). Information generated by this method may be helpful in estimating the effect of a test material on bacterial communities in the aquatic environment, especially in aerobic biological treatment systems (Paixão et al., 2002). In the first part of this experimental study, untreated (original) aqueous DOS and PB solutions were individually exposed to activated sludge inhibition experiments. The obtained semi-logarithmic IOUR (in %) versus DOS and PB concentrations (log DOS and log PB) were presented in Figure 1 (a) and (b), respectively for an incubation period of 15 min.

**3.1 Activated sludge inhibition experiments with untreated DOS and PB**

**2.7 Electrical energy requirements of the photochemical treatment processes**

ln C C⁄ � = − k� × t (17)

EE O⁄ (kWh m��order��) = (38.4 × P) ⁄ (V × k�) (18)

calculated as 10 times of the signal-to-noise ratio as 5 mg L-1.

Hence, the kinetic rates could be expressed as follows;

**2.6 Kinetic analysis**

(DOS and PB, in mg L-1).

degradation (min−1).

**3. Results and discussion**

determined from the following formula;

Thermo Orion model 720 pH-meter at any stage of the experiments.

From Figure 1 (a) it is evident that increasing the DOS concentration resulted in an accelerated inhibition of the OUR value measured relative to the sample blank (synthetic sewage). The IOUR values ranged between 4%-53% from the lowest to highest studied concentrations, respectively, enabling the determination of EC30 and EC50 values (effective concentrations causing 30 and 50% inhibition in heterotrophic biomass relative to the blank) for DOS, as shown in the insert of Figure 1 (a). These values were established as 338 and 544 mg L-1 corresponding to the respirometric inhibition levels of 30% and 50%, respectively. This information implied that DOS is an acutely toxic compound at concentrations encountered in the textile preparation process. Hence its treatment after discharge becomes essential. On the other hand, Figure 1 (b) reveals that, although the studied concentration range for PB is appreciably higher, activated sludge inhibition rates always remained below 20%. From the obtained experimental results it can be concluded that PB is not exhibiting an inhibitory/toxic effect on heterotrophic biomass relative to synthetic sewage. However, it should be noted that the inert COD content of PB creates a significant effluent discharge penalty problem for dye synthesis effluent. For both chemicals it is obvious that efficient treatment employing advanced oxidation process is necessary to comply with the environmental regulations and to minimize pollution loads.

Fig. 1. Log(concentration) versus IOUR (%) curves obtained from the activated sludge inhibition experiments for aqueous DOS (a) and PB (b). DOS concentration range: 97-584 mg L-1; PB concentration range: 144-721 mg L-1. The selected incubation time is 15 min.

#### **3.2 H2O2/UV-C and Photo-Fenton treatment of DOS and PB**

Figure 2 presents the H2O2/UV-C and Photo-Fenton treatment results for aqueous DOS (a, H2O2/UV-C; b, Photo-Fenton) and PB (c, H2O2/UV-C; d, Photo-Fenton) solutions in terms of DOS-PB, COD and TOC abatement rates as a function of photochemical treatment time. In the Photo-Fenton experiments all experimental conditions were kept identical with the exception of 0.5 mM Fe2+ addition to the reaction solution. The initial COD of the DOS-PB solutions were selected as 450 mg L-1 in these experiments, which is an average, typical COD for effluents containing these model compounds as the main organic pollutant.

As it can be seen from Figure 2 (a) and (b) complete DOS removal was obtained after 60 min for both treatment processes. In terms of the COD parameter, 84% and 92% removals were achieved after 100 min H2O2/UV-C and Photo-Fenton treatment, respectively. TOC removals were also close to another; these were obtained as 89% after 100 min H2O2/UV-C treatment and 90% at the end of 100 min Photo-Fenton treatment. From these findings it can be concluded that DOS, COD and TOC removal efficiencies were not significantly different or higher for the Photo-Fenton process as compared to H2O2/UV-C treatment. Considering that the addition of Fe2+ as a photocatalyst and extra chemical as well as the requirement of acidic reaction conditions for iron-based AOPs will affect the operating costs for DOS treatment considerably, it was decided to keep the H2O2/UV-C process for DOS treatment. Kinetic rate coefficients and electrical energy calculations shown in Table 2 and explained later on supported this argument.

Figure 2 (c) and (d) depict the results of H2O2/UV-C and Photo-Fenton treatment of aqueous PB solution, respectively. As expected, the addition of 0.5 mM Fe2+ had a significant positive effect in terms of PB as well as COD and TOC abatement rates as compared with the H2O2/UV-C process (see Figure 2 (c) and (d)). Total PB elimination was realized only after 30 min Photo-Fenton treatment, whereas at least 180 min were needed when PB was subjected to H2O2/UV-C treatment. Moreover, COD and TOC were only poorly removed via mere H2O2/UV-C treatment. For instance, parallel to the slow COD abatement, TOC practically remained unchanged indicating that PB, the mother compound, was converted to different advanced oxidation intermediates but not ultimately oxidized to mineralization end products in the absence of Fe2+ ions. At the end of the 120 min treatment period, COD and TOC removals were obtained as 49% (H2O2/UV-C treatment) and 98% (Photo-Fenton treatment) as well as 33% (H2O2/UV-C treatment) and 100% (Photo-Fenton treatment), respectively. Considering that H2O2/UV-C treatment of PB resulted in very poor and insufficient results, Photo-Fenton oxidation appears by far the better option for the efficient treatment of PB.

The results obtained for DOS and PB could be partially attributed to the high absorbance (optical density) of PB at the maximum emission band of the UV-C light source, hindering its effective absorption by H2O2, whereas DOS did not compete with H2O2 for UV-C light absorption at 254 nm wavelength that is required for the production of HO● radicals via UV-C.

Table 2 summarizes the calculated kC and EE/O values for the photochemical treatment of DOS and PB in the absence and presence of Fe2+ catalyst. From Table 2 it is obvious that for DOS treatment, Fe2+ is not essential, whereas the positive effect of the catalyst was very pronounced for PB abatement. Hence, in the forthcoming experiments where changes in the acute toxicity of DOS and PB during photochemical treatment were examined, results were presented only for the selected AOPs (e.g. H2O2/UV-C for DOS and Photo-Fenton for PB model pollutants).

Figure 2 presents the H2O2/UV-C and Photo-Fenton treatment results for aqueous DOS (a, H2O2/UV-C; b, Photo-Fenton) and PB (c, H2O2/UV-C; d, Photo-Fenton) solutions in terms of DOS-PB, COD and TOC abatement rates as a function of photochemical treatment time. In the Photo-Fenton experiments all experimental conditions were kept identical with the exception of 0.5 mM Fe2+ addition to the reaction solution. The initial COD of the DOS-PB solutions were selected as 450 mg L-1 in these experiments, which is an average, typical COD

As it can be seen from Figure 2 (a) and (b) complete DOS removal was obtained after 60 min for both treatment processes. In terms of the COD parameter, 84% and 92% removals were achieved after 100 min H2O2/UV-C and Photo-Fenton treatment, respectively. TOC removals were also close to another; these were obtained as 89% after 100 min H2O2/UV-C treatment and 90% at the end of 100 min Photo-Fenton treatment. From these findings it can be concluded that DOS, COD and TOC removal efficiencies were not significantly different or higher for the Photo-Fenton process as compared to H2O2/UV-C treatment. Considering that the addition of Fe2+ as a photocatalyst and extra chemical as well as the requirement of acidic reaction conditions for iron-based AOPs will affect the operating costs for DOS treatment considerably, it was decided to keep the H2O2/UV-C process for DOS treatment. Kinetic rate coefficients and electrical energy calculations shown in Table 2 and explained

Figure 2 (c) and (d) depict the results of H2O2/UV-C and Photo-Fenton treatment of aqueous PB solution, respectively. As expected, the addition of 0.5 mM Fe2+ had a significant positive effect in terms of PB as well as COD and TOC abatement rates as compared with the H2O2/UV-C process (see Figure 2 (c) and (d)). Total PB elimination was realized only after 30 min Photo-Fenton treatment, whereas at least 180 min were needed when PB was subjected to H2O2/UV-C treatment. Moreover, COD and TOC were only poorly removed via mere H2O2/UV-C treatment. For instance, parallel to the slow COD abatement, TOC practically remained unchanged indicating that PB, the mother compound, was converted to different advanced oxidation intermediates but not ultimately oxidized to mineralization end products in the absence of Fe2+ ions. At the end of the 120 min treatment period, COD and TOC removals were obtained as 49% (H2O2/UV-C treatment) and 98% (Photo-Fenton treatment) as well as 33% (H2O2/UV-C treatment) and 100% (Photo-Fenton treatment), respectively. Considering that H2O2/UV-C treatment of PB resulted in very poor and insufficient results, Photo-Fenton oxidation

The results obtained for DOS and PB could be partially attributed to the high absorbance (optical density) of PB at the maximum emission band of the UV-C light source, hindering its effective absorption by H2O2, whereas DOS did not compete with H2O2 for UV-C light absorption at 254 nm wavelength that is required for the production of HO● radicals via

Table 2 summarizes the calculated kC and EE/O values for the photochemical treatment of DOS and PB in the absence and presence of Fe2+ catalyst. From Table 2 it is obvious that for DOS treatment, Fe2+ is not essential, whereas the positive effect of the catalyst was very pronounced for PB abatement. Hence, in the forthcoming experiments where changes in the acute toxicity of DOS and PB during photochemical treatment were examined, results were presented only for the selected AOPs (e.g. H2O2/UV-C for DOS and Photo-Fenton for PB

for effluents containing these model compounds as the main organic pollutant.

**3.2 H2O2/UV-C and Photo-Fenton treatment of DOS and PB**

appears by far the better option for the efficient treatment of PB.

later on supported this argument.

UV-C.

model pollutants).

**Photochemical Treatment Time (min)**

Fig. 2. DOS, PB, COD and TOC abatement rates observed during photochemical treatment of DOS (a, H2O2/UV-C; b, Photo-Fenton) and PB (c, H2O2/UV-C; d, Photo-Fenton). For H2O2/UV-C experiments: Initial COD= 450 mg L-1; H2O2 = 30 mM (DOS) and 60 mM (PB); pH = 10.5 (DOS) and 3.0 (PB). For Fe2+/H2O2/UV-C experiments: Initial COD= 450 mg L-1; H2O2 = 30 mM; Fe2+ = 0.5 mM; pH = 3.0.

Table 2. First-order abatement rate constants and EE/O values derived for the treatment of aqueous DOS and PB solutions with the H2O2/UV-C and Fe2+/H2O2/UV-C processes.

#### **3.3 Activated sludge inhibition results for photochemically treated DOS and PB**

The possibility exists that during the applicability of AOPs, more toxic and/or less bioamenable intermediates and/or end products form and accumulate in the reaction solution. In order to evaluate the acute toxicity and bioamenability of photochemical degradation products towards heterotrophic biomass, activated sludge inhibition experiments were conducted during H2O2/UV-C of DOS and Photo-Fenton oxidation of PB (Figure 3).

For the activated sludge experiments the OUR of the blank samples (containing only synthetic sewage) was adjusted to around 100 mg L-1 h-1. The OURs of the sample blanks were also given in Figure 3. Figure 3 indicated that a slight increase in the OUR values is observed during photochemical treatment of DOS starting with 71 mg L-1 h-1 and reaching its highest value after 40 min photochemical oxidation. Thereafter, the OUR value dropped down even below its original value indicating that extended oxidation did not improve but reduce the quality of DOS bearing effluent most probably due to the formation of less biodegradable and hence more inhibitory end products. The trend observed for the OUR values of the PB samples exposed to Photo-Fenton was different; the OUR values increased throughout the entire treatment period from 98 mg L-1 h-1 at the beginnig up to 120 mg L-1 h-1 after 80 min Photo-Fenton treatment. These results demonstrate that PB does not cause any inibitory effect of heterotrophic biomass and the treated reaction solution was even accepted as an extra carbon source for the bioculture as advanced oxidation progressed. The OUR results implied that no acutely toxic/bioinhibitory advanced oxidation products were formed during Photo-Fenton treatment of PB.

Fig. 3. OUR values obtained during photochemical treatment of DOS via H2O2/UV-C (H2O2=42 mM; pH=10.5) and PB via Photo-Fenton (Fe2+= 0.72 mM; H2O2=40 mM; pH=3.0) processes. Experimental conditions: Initial COD = 450 mg L-1; Incubation period = 15 min; OURblank for DOS = 92 mg L-1 h-1; OURblank for PB = 118 mg L-1 h-1.

#### **4. Conclusion**

84 Advances in Treating Textile Effluent

Fig. 2. DOS, PB, COD and TOC abatement rates observed during photochemical treatment of DOS (a, H2O2/UV-C; b, Photo-Fenton) and PB (c, H2O2/UV-C; d, Photo-Fenton). For H2O2/UV-C experiments: Initial COD= 450 mg L-1; H2O2 = 30 mM (DOS) and 60 mM (PB); pH = 10.5 (DOS) and 3.0 (PB). For Fe2+/H2O2/UV-C experiments: Initial COD= 450 mg L-1;

**0 20 40 60 80 100 120**

**Photochemical Treatment Time (min)**

**Fe2+/H2O2/UV-C**  Initial COD= 450 mg L-1 H2O2 = 30 mM; Fe2+ = 0.5 mM

**PB COD TOC**

EE/O

(kWh m-3 order-1)

pH = 3.0

kC (min-1)

H2O2 = 30 mM; Fe2+ = 0.5 mM; pH = 3.0.

**0**

**100**

**200**

**PB, COD and TOC**

 **(mg L-1)**

**300**

**400**

**500**

**(d)**

kC (min-1)

**H2O2/UV-C**

Initial COD= 450 mg L-1

pH = 10.5 (DOS) and 3.0 (PB)

EE/O

H2O2 = 30 mM (DOS) and 60 mM (PB)

DOS 0.1500 3.15 0.1016 4.65 PB 0.0230 20.55 0.2084 2.27

(kWh m-3 order-1)

Table 2. First-order abatement rate constants and EE/O values derived for the treatment of aqueous DOS and PB solutions with the H2O2/UV-C and Fe2+/H2O2/UV-C processes.

For the activated sludge experiments the OUR of the blank samples (containing only synthetic sewage) was adjusted to around 100 mg L-1 h-1. The OURs of the sample blanks were also given in Figure 3. Figure 3 indicated that a slight increase in the OUR values is observed during photochemical treatment of DOS starting with 71 mg L-1 h-1 and reaching its highest value after 40 min photochemical oxidation. Thereafter, the OUR value dropped down even below its original value indicating that extended oxidation did not improve but reduce the quality of DOS bearing effluent most probably due to the formation of less

**3.3 Activated sludge inhibition results for photochemically treated DOS and PB**  The possibility exists that during the applicability of AOPs, more toxic and/or less bioamenable intermediates and/or end products form and accumulate in the reaction solution. In order to evaluate the acute toxicity and bioamenability of photochemical degradation products towards heterotrophic biomass, activated sludge inhibition experiments were conducted during H2O2/UV-C of DOS and Photo-Fenton oxidation of PB

**Model Pollutants**

(Figure 3).

The present study aimed at exploring the effect of photochemical advanced oxidation processes (H2O2/UV-C and Photo-Fenton treatment) on the treatability and bioamenability of two frequently used industrial chemicals; i.e. an anionic textile surfactant (called DOS herein) and a vinylsuphone dye precursor (called PB herein). The following experimental results were obtained in this work;

mere H2O2/UV-C treatment in the absence of Fe2+ photocatalyst was sufficient for complete DOS and associated organic carbon content abatements. The main advantage of the H2O2/UV-C treatment process over Photo-Fenton treatment is its robustness towards reaction pH below pH = 11 hence not requiring a pH adjustment to acidic values.

 For aromatic compounds that strongly absorp UV-C light at 254 nm thus hindering its efficient use by H2O2 for HO● production, Photo-Fenton treatment improves the degradation rate considerably. PB abatement via H2O2/UV-C treatment process resulted in relatively slow and incomplete degradation rates. In the presence of Fe2+, however, PB degradation was enhanced by a factor of 10 and its removal was complete after 30-40 min Photo-Fenton treatment. These results suggest that the H2O2/UV-C process is not capable of effectively degrading the complicated aromatic structure of PB and under these circumstances; Photo-Fenton treatment is the treatment process of choice.

According to the activated sludge inhibition tests no toxic/bioinhibitory degradation intermediates and/or oxidation end products were formed during and after photochemical treatment of DOS (60 min H2O2/UV-C treatment at an initial pH of 10.5 with 42 mM H2O2) and PB (Photo-Fenton treatment for 80 min at pH 3; in the presence of 0.72 mM Fe2+and 40 mM H2O2). The obtained findings indicated that photochemical advanced oxidation processes are potential, promising options for the effective treatment of refractory and toxic industrial pollutants.

#### **5. Acknowledgments**

The authors acknowledge the financial support of the Scientific and Technological Research Council of Turkey (TUBITAK) under project number 108Y051 and Eksoy Chemicals for the DOS and PB gift samples. The authors also thank Ms. Zeynep Kartal and Ms. Elif Dedetas for their help in the photochemical treatment experiments.

#### **6. References**

 For aromatic compounds that strongly absorp UV-C light at 254 nm thus hindering its efficient use by H2O2 for HO● production, Photo-Fenton treatment improves the degradation rate considerably. PB abatement via H2O2/UV-C treatment process resulted in relatively slow and incomplete degradation rates. In the presence of Fe2+, however, PB degradation was enhanced by a factor of 10 and its removal was complete after 30-40 min Photo-Fenton treatment. These results suggest that the H2O2/UV-C process is not capable of effectively degrading the complicated aromatic structure of PB and under these circumstances; Photo-Fenton treatment is the treatment process of

According to the activated sludge inhibition tests no toxic/bioinhibitory degradation intermediates and/or oxidation end products were formed during and after photochemical treatment of DOS (60 min H2O2/UV-C treatment at an initial pH of 10.5 with 42 mM H2O2) and PB (Photo-Fenton treatment for 80 min at pH 3; in the presence of 0.72 mM Fe2+and 40 mM H2O2). The obtained findings indicated that photochemical advanced oxidation processes are potential, promising options for the effective treatment of refractory and toxic

The authors acknowledge the financial support of the Scientific and Technological Research Council of Turkey (TUBITAK) under project number 108Y051 and Eksoy Chemicals for the DOS and PB gift samples. The authors also thank Ms. Zeynep Kartal and Ms. Elif Dedetas

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## **Textile Dyeing Wastewater Treatment**

Zongping Wang, Miaomiao Xue, Kai Huang and Zizheng Liu *Huazhong University of Science and Technology China*

#### **1. Introduction**

90 Advances in Treating Textile Effluent

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> Textile industry can be classified into three categories viz., cotton, woolen, and synthetic fibers depending upon the used raw materials. The cotton textile industry is one of the oldest industries in China.

> The textile dyeing industry consumes large quantities of water and produces large volumes of wastewater from different steps in the dyeing and finishing processes. Wastewater from printing and dyeing units is often rich in color, containing residues of reactive dyes and chemicals, such as complex components,many aerosols,high chroma,high COD and BOD concentration as well as much more hard-degradation materials. The toxic effects of dyestuffs and other organic compounds, as well as acidic and alkaline contaminants, from industrial establishments on the general public are widely accepted. At present, the dyes are mainly aromatic and heterocyclic compounds, with color-display groups and polar groups. The structure is more complicated and stable, resulting in greater difficulty to degrade the printing and dyeing wastewater (Shaolan Ding et al.,2010).

> According to recent statistics, China's annual sewage has already reached 390 million tons, including 51% of industrial sewage, and it has been increasing with the rate of 1% every year. Each year about 70 billion tons of wastewater from textile and dyeing industry are produced and requires proper treatment before being released into the environment (State Environmental Protection Administration ,1994).

> Therefore, understanding and developing effective printing-dye industrial wastewater treatment technology is environmentally important.

#### **1.1 Textile printing and dyeing process**

Textile Printing and dyeing processes include pretreatment, dyeing / printing, finishing and other technologies.

Pre-treatment includes desizing, scouring, washing, and other processes. Dyeing mainly aims at dissolving the dye in water, which will be transferred to the fabric to produce colored fabric under certain conditions. Printing is a branch of dyeing which generally is defined as 'localized dyeing' i.e. dyeing that is confirmed to a certain portion of the fabric that constitutes the design. It is really a form of dyeing in which the essential reactions involved are the same as those in dyeing. In dyeing, color is applied in the form of solutions, whereas color is applied in the form of a thick paste of the dye in printing. Both natural and synthetic textiles are subjected to a variety of finishing processes. This is done to improve specific properties in the finished fabric and involves the use of a large number of finishing agents for softening, crosslinking, and waterproofing. All of the finishing processes contribute to water pollution. In

Raw wool, cotton Man-made staple fibers Fiber preparation Spinning Knitting W arping Slashing Man-made staple fibers Texturizing Texturizing W eaving Preparation De-sizing Bleaching Dyeing, printing Finishing M ercerizing Yarn formation Yarn formation

addition, in different circumstances, the singeing, mercerized, base reduction, and other processes should have been done before dyeing/printing.

Fig. 1. Various steps involved in processing textile in a cotton mill (B. Ramesh Babu et al.,2007).

In the textile dyeing industry, bleaching is an important process. It has three technologies: sodium hypochlorite bleaching; hydrogen peroxide bleaching and sodium chlorite bleaching. Sodium hypochlorite bleaching and sodium chlorite bleaching are the most commonly used processes. Normal concentration of chlorine dioxide in bleaching effluent is 10-200 mg/L. As chlorine dioxide is a strong oxidant, it is very corrosive and toxic as well.

The typical printing and dyeing process is shown in Fig. 1 and the main used fiber dyes at present have been shown in Table 1 (Kelu Yan,2005).

Table 1. The varieties of common used fiber

addition, in different circumstances, the singeing, mercerized, base reduction, and other

Fiber preparation

Spinning

Knitting

M ercerizing

Fig. 1. Various steps involved in processing textile in a cotton mill (B. Ramesh Babu et al.,2007). In the textile dyeing industry, bleaching is an important process. It has three technologies: sodium hypochlorite bleaching; hydrogen peroxide bleaching and sodium chlorite bleaching. Sodium hypochlorite bleaching and sodium chlorite bleaching are the most commonly used processes. Normal concentration of chlorine dioxide in bleaching effluent is 10-200 mg/L. As chlorine dioxide is a strong oxidant, it is very corrosive and toxic as well. The typical printing and dyeing process is shown in Fig. 1 and the main used fiber dyes at

The variety of fiber The commonly used dyes

Wool Acid dyes Silk Direct dyes,Acid dyes Polyester Azo dyes,Disperse dyes Polyester-cotton Disperse / Vat dyes,Disperse / Insoluble dye

polyacrylonitrile fiber Cationic dyes, Disperse dyes polyacrylonitrile fiber -wool Cationic dyes,Acid dyes

Cellulose fiber Direct dyes, Reactive dyes,Vat dyes,

vinylon Direct dyes,Vat dyes,Sulfur dyes,Acid dyes,

Sulfide dyes, Azo dyes

Man-made staple fibers

W arping

Slashing

W eaving

Preparation

De-sizing

Bleaching

Dyeing, printing

Finishing

present have been shown in Table 1 (Kelu Yan,2005).

Table 1. The varieties of common used fiber

Raw wool, cotton

> Yarn formation

> Yarn formation

processes should have been done before dyeing/printing.

Man-made staple fibers

Texturizing

Texturizing

#### **1.2 Production of textile industry pollution**

Textile Printing and dyeing processes include pre-treatment, dyeing and printing, finishing. The main pollutants are organic matters which come from the pre-treatment process of pulp, cotton gum, cellulose, hemicellulose and alkali, as well as additives and dyes using in dyeing and printing processes. Pre-treatment wastewater accounts for about 45% of the total, and dyeing/printing process wastewater accounts for about 50%~55%, while finishing process produces little.

In China, chemical fiber accounts for about 69% of total in which polyester fibers accounts for more than 80%. Cotton accounts for 80% of the natural fiber production. Therefore, the dyeing wastewater analysis of production and pollution is based on these two fibers.

Pre-treatment of cotton includes desizing and scouring. The main pollutants are the impurities in the cotton, cotton gum, hemicellulose and the slurry, alkali in weaving process. The current average COD concentration in the pre-treatment is 3000 mg/L. The main pollutants in dyeing/printing are auxiliaries and the residual dyes. The average concentration of COD is 1000 mg/L and the total average concentration is 2000 mg/L after mixing.

Pre-treatment of polyester fibers mainly involves in the reduction with alkali. The so-called reduction is treating the polyester fabric with 8% of sodium hydroxide at 90 C for about 45 minutes. Some polyester fabrics will peel off and decompose into terephthalic acid and ethylene glycol so that a thin polyester fabric will have the feel of silk. This process can be divided into continuous and batch type. Taking the batch type as an example, the concentration of COD is up to 20000 mg/L-60000 mg/L. The wastewater from reduction process may account for only 5% of the volume of wastewater, while COD accounts for 60% or more in the conventional dyeing and finishing business.

The chroma is one pollutant of the wastewater which causes a lot of concerns. In the dyeing process, the average dyeing rate is more than 90%. It means that the residual dyeing rate in finishing wastewater is about 10%, which is the main reason of contamination. According to the different dyes and process, the chroma is 200 to 500 times higher than before.

pH is another factor of the dyeing wastewater. Before the printing and the dyeing process, pH is another factor ,the pH of dyeing wastewater remains between 10 to 11 when treated by alkali at high temperature around 90C in the process of desizing, scouring and mercerization. Polyester base reduction process mainly uses sodium hydroxide, and the total pH is also 10 to 11. Therefore, most dyeing water is alkaline and the first process is to adjust the pH value of the textile dyeing wastewater.

The total nitrogen and ammonia nitrogen come from dyes and raw materials, which is not very high, about 10 mg/L. But the urea is needed while using batik techniques. Its total nitrogen is 300 mg/L, which is hard to treat. The phosphorus in the wastewater comes from the phosphor detergents. Considering the serious eutrophication of surface water, it needs to be controlled. Some enterprises use trisodium phosphate so that the concentration of phosphorus will reach 10 mg/L. So, this phosphorus must be removed in the pre-treatment.

In the production process, suspended substance comes from fiber scrap and undissolved raw materials. It will be removed through the grille, grid, etc. The suspended solids (SS) in the outflow mainly come from the secondary sedimentation tank, whose sludge has not been separated completely which will reach 10-100 mg/L as usual.

Sulfide mainly comes from the sulfur, which is a kind of cheap and qualified dye. Due to its toxicity, it has been forbidden in developed countries. However, in China, some enterprises are still using it, so it has been included in the wastewater standards. The sulfide in the wastewater is about 10 mg/L. There are two main sources of hexavalent chromium. Cylinder engraving makes the wastewater containing hexavalent chromium. However, this technology has not been used. Another possible source is the use of potassium dichromate additive in hair dyeing process. Aniline mainly comes from the dyes. The color of the dye comes from the chromophore. Some dyes have a benzene ring, amino, etc., which will be decomposed in the wastewater treatment process.

The potential specific pollutants from textile Printing and dyeing is shown in Table 2 (C. All`egre et al.,2006).

Table 2. Specific pollutants from textile and dyeing processing operations

#### **1.3 Textile dyeing wastewater risk**

Discharged wastewater by some industries under uncontrolled and unsuitable conditions is causing significant environmental problems. The importance of the pollution control and treatment is undoubtedly the key factor in the human future. If a textile mill discharges the wastewater into the local environment without any treatment, it will has a serious impact on natural water bodies and land in the surrounding area. High values of COD and BOD5, presence of particulate matter and sediments, and oil and grease in the effluent causes depletion of dissolved oxygen, which has an adverse effect on the aquatic ecological system.

Effluent from textile mills also contains chromium, which has a cumulative effect, and higher possibilities for entering into the food chain. Due to usage of dyes and chemicals, effluents are dark in color, which increases the turbidity of water body. This in turn hampers the photosynthesis process, causing alteration in the habitat (Joseph Egli,2007).

#### **1.4 The textile industry standards for water pollutants**

As the wastewater is harmful to the environment and people, there are strict requirements for the emission of the wastewater. However, due to the difference in the raw materials, products, dyes, technology and equipment, the standards of the wastewater emission have too much items. It is developed by the national environmental protection department according to the local conditions and environmental protection requirements which is not fixed. It varies according to the situation in different regions. Therefore, the nature of emission targets is priorities of the points.

are still using it, so it has been included in the wastewater standards. The sulfide in the wastewater is about 10 mg/L. There are two main sources of hexavalent chromium. Cylinder engraving makes the wastewater containing hexavalent chromium. However, this technology has not been used. Another possible source is the use of potassium dichromate additive in hair dyeing process. Aniline mainly comes from the dyes. The color of the dye comes from the chromophore. Some dyes have a benzene ring, amino, etc., which will be

The potential specific pollutants from textile Printing and dyeing is shown in Table 2 (C.

Desizing Sizes, enzymes, starch, waxes, ammonia.

Bleaching H2O2, AOX, sodium silicate or organic stabiliser, high pH.

Dyeing Colour, metals, salts, surfactants, organic processing assistants,

Finishing Resins, waxes, chlorinated compounds, acetate, stearate, spent

Discharged wastewater by some industries under uncontrolled and unsuitable conditions is causing significant environmental problems. The importance of the pollution control and treatment is undoubtedly the key factor in the human future. If a textile mill discharges the wastewater into the local environment without any treatment, it will has a serious impact on natural water bodies and land in the surrounding area. High values of COD and BOD5, presence of particulate matter and sediments, and oil and grease in the effluent causes depletion of dissolved oxygen, which has an adverse effect on the aquatic

Effluent from textile mills also contains chromium, which has a cumulative effect, and higher possibilities for entering into the food chain. Due to usage of dyes and chemicals, effluents are dark in color, which increases the turbidity of water body. This in turn hampers the photosynthesis process, causing alteration in the habitat (Joseph Egli,2007).

As the wastewater is harmful to the environment and people, there are strict requirements for the emission of the wastewater. However, due to the difference in the raw materials, products, dyes, technology and equipment, the standards of the wastewater emission have too much items. It is developed by the national environmental protection department according to the local conditions and environmental protection requirements which is not fixed. It varies according to the situation in different regions. Therefore, the nature of

Disinfectants and insecticides residues, NaOH, surfactants, soaps, fats, waxes, pectin, oils, sizes, anti-static agents, spent solvents, enzymes.

sulphide, acidity/alkalinity, formaldehyde.

solvents, softeners.

Process Compounds

Mercerizing High pH, NaOH

Table 2. Specific pollutants from textile and dyeing processing operations

**1.4 The textile industry standards for water pollutants**

emission targets is priorities of the points.

Printing Urea, solvents, colour, metals.

decomposed in the wastewater treatment process.

All`egre et al.,2006).

Scouring

**1.3 Textile dyeing wastewater risk**

ecological system.

Table 3. "Textile industry standards for water pollutants"

For printing and dyeing wastewater, the first consideration is the organic pollutants, color and heavy metal ions. Recently, as the lack of water, the recovery of wastewater should be considered. So the decolorization of the printing and dyeing wastewater increased heavily. The standards of printing and dyeing are different in different countries.Through access to the relevant information, the textile industry standards for water pollutants in China, Germany, U.S have been found.

#### **1.4.1 Textile industry standards for water pollutants in China**

For the emission standards of the textile dyeing wastewater in China, it is the very stringentstandards in the world. The emission standards for different indicators in textile industry standards for water pollutants in China have been shown in table 3 ("Discharge standard of water pollutants for dyeing and finishing of textile industry").

#### **1.4.2 Textile industry standards for water pollutants in Germany**

The emission standards for different indicators in textile industry standards for water pollutants in Germany have been shown in table 4 ("Discharge standard of water pollutants for dyeing and finishing of textile industry").

Table 4. Textile industry standards for water pollutants

The requirements of ammonia and total nitrogen are adjusted for the biochemical outflow at 12°C or above. Besides, the standard has also made the following emission requirements for the wastewater at the production stain.

There must not be in the wastewater:

#### **1.4.3 Textile industry standards for water pollutants in U.S.**

#### **Printing and dyeing wastewater**

It is the order for the printing and dyeing, including rinsing, dyeing, bleaching, washing, drying and other similar processes.

The requirements using BPT (best practical control tech.)which is published by EPA has shown in Table 5 ("Discharge standard of water pollutants for dyeing and finishing of textile industry").

#### **Fabric printing and dyeing wastewater**

It is adjusted for the fabric printing and dyeing wastewater, including bleaching, mercerization, dyeing, resin processing, washing, drying and so on.

The requirements using BPT (best practical control tech.) to treat the fabric printing and dyeing wastewater has been shown in Table 6 ("Discharge standard of water pollutants for dyeing and finishing of textile industry").

Table 5. Emission standards for gross printing and dyeing wastewater

#### **Yarn printing and dyeing wastewater**

It is adjusted for the Yarn printing and dyeing wastewater, including washing, mercerizati on, resin processing, dyeing and special finishing.

The requirements of ammonia and total nitrogen are adjusted for the biochemical outflow at 12°C or above. Besides, the standard has also made the following emission requirements for

2. Separation of chlorine bleach materials, except the sodium chlorite from the bleached

It is the order for the printing and dyeing, including rinsing, dyeing, bleaching, washing,

The requirements using BPT (best practical control tech.)which is published by EPA has shown in Table 5 ("Discharge standard of water pollutants for dyeing and finishing of

It is adjusted for the fabric printing and dyeing wastewater, including bleaching,

The requirements using BPT (best practical control tech.) to treat the fabric printing and dyeing wastewater has been shown in Table 6 ("Discharge standard of water pollutants for

1 BOD5 22.4 11.2 2 COD 163.0 81.5 3 TSS 35.2 17.6 4 S 0.28 0.14 5 Phenol 0.14 0.07 6 Cr 0.14 0.07 7 pH 6.0~9.0 6.0~9.0

It is adjusted for the Yarn printing and dyeing wastewater, including washing, mercerizati

BPT Maximum Average of 30 days Kg/t(Fabric)

the wastewater at the production stain. There must not be in the wastewater:

4. Arsenic, mercury and their mixtures

**Printing and dyeing wastewater**

drying and other similar processes.

**Fabric printing and dyeing wastewater**

dyeing and finishing of textile industry").

Serial number Parameters

**Yarn printing and dyeing wastewater**

on, resin processing, dyeing and special finishing.

textile industry").

synthetic fibers

1. Organic chlorine carriers (dyed acceleration)

3. The free chlorine after using sodium chlorite

5. Alkyl phenol as a bleaching agent(APEO)

6. Cr6 + compounds in the oxidizing of sulfur dyes and vat dyes 7. EDTA, DTPA, and phosphate in the water treatment softeners 8. Accumulation of chemicals, dyes and textile auxiliaries

**1.4.3 Textile industry standards for water pollutants in U.S.**

mercerization, dyeing, resin processing, washing, drying and so on.

Table 5. Emission standards for gross printing and dyeing wastewater

The requirements using BPT (best practical control tech.) to treat the yarn printing and dyeing wastewater has been shown in Table 7 ("Discharge standard of water pollutants for dyeing and finishing of textile industry").

Table 6. Emission standards for fabric printing and dyeing wastewater

Table 7. Emission standards for yarn printing and dyeing wastewater

#### **2. Textile dyeing wastewater treatment processes**

The textile dyeing wastewater has a large amount of complex components with high concentrations of organic, high-color and changing greatly characteristics. Owing to their high BOD/COD, their coloration and their salt load, the wastewater resulting from dyeing cotton with reactive dyes are seriously polluted. As aquatic organisms need light in order to develop, any deficit in this respect caused by colored water leads to an imbalance of the ecosystem. Moreover, the water of rivers that are used for drinking water must not be colored, as otherwise the treatment costs will be increased. Obviously, when legal limits exist (not in all the countries) these should be taken as justification. Studies concerning the feasibility of treating dyeing wastewater are very important (C. All`egre et al.,2006).

In the past several decades, many techniques have been developed to find an economic and efficient way to treat the textile dyeing wastewater, including physicochemical, biochemical, combined treatment processes and other technologies. These technologies are usually highly efficient for the textile dyeing wastewater.

#### **2.1 Physicochemical wastewater treatment**

Wastewater treatment is a mixture of unit processes, some physical, others chemical or biological in their action. A conventional treatment process is comprised of a series of individual unit processes, with the output (or effluent) of one process becoming the input (influent) of the next process. The first stage will usually be made up of physical processes. Physicochemical wastewater treatment has been widely used in the sewage treatment plant which has a high removal of chroma and suspended substances, while it has a low removal of COD. The common physicochemical methods are shown as followed.

#### **2.1.1 Equalization and homogenization**

Because of water quality highly polluted and quantity fluctuations, complex components, textile dyeing wastewater is generally required pretreatment to ensure the treatment effect and stable operation.

In general, the regulating tank is set to treat the wastewater. Meantime,to prevent the lint, cotton seed shell, and the slurry Settle to the bottom of the tank, it's usually mixed the wastewater with air or mechanical mixing equipment in the tank. The hydraulic retention time is generally about 8 h.

#### **2.1.2 Floatation**

The floatation produces a large number of micro-bubbles in order to form the three-phase substances of water, gas, and solid. Dissolved air under pressure may be added to cause the formation of tiny bubbles which will attach to particles. Under the effect of interfacial tension, buoyancy of bubble rising, hydrostatic pressure and variety of other forces, the microbubble adheres to the tiny fibers. Due to its low density, the mixtures float to the surface so that the oil particles are separated from the water. So, this method can effectively remove the fibers in wastewater.

#### **2.1.3 Coagulation flocculation sedimentation**

Coagulation flocculation sedimentation is one of the most used methods, especially in the conventional treatment process.

Active on suspended matter, colloidal type of very small size, their electrical charge give repulsion and prevent their aggregation. Adding in water electrolytic products such as aluminum sulphate, ferric sulphate, ferric chloride, giving hydrolysable metallic ions or organic hydrolysable polymers (polyelectrolyte) can eliminate the surface electrical charges of the colloids. This effect is named coagulation. Normally the colloids bring negative charges,so the coagulants are usually inorganic or organic cationic coagulants (with positive charge in water).

The metallic hydroxides and the organic polymers, besides giving the coagulation, can help the particle aggregation into flocks, thereby increasing the sedimentation. The combined action of coagulation, flocculation and settling is named clariflocculation.

Settling needs stillness and flow velocity, so these three processes need different reactions tanks. This processes use mechanical separation among heterogeneous matters, while the dissolved matter is not well removed (clariflocculation can eliminate a part of it by absorption into the flocks). The dissolved matter can be better removed by biological or by other physical chemical processes (Sheng.H et al.,1997) .

But additional chemical load on the effluent (which normally increases salt concentration) increases the sludge production and leads to the uncompleted dye removal.

#### **2.1.4 Chemical oxidation**

98 Advances in Treating Textile Effluent

Wastewater treatment is a mixture of unit processes, some physical, others chemical or biological in their action. A conventional treatment process is comprised of a series of individual unit processes, with the output (or effluent) of one process becoming the input (influent) of the next process. The first stage will usually be made up of physical processes. Physicochemical wastewater treatment has been widely used in the sewage treatment plant which has a high removal of chroma and suspended substances, while it has a low removal

Because of water quality highly polluted and quantity fluctuations, complex components, textile dyeing wastewater is generally required pretreatment to ensure the treatment effect

In general, the regulating tank is set to treat the wastewater. Meantime,to prevent the lint, cotton seed shell, and the slurry Settle to the bottom of the tank, it's usually mixed the wastewater with air or mechanical mixing equipment in the tank. The hydraulic retention

The floatation produces a large number of micro-bubbles in order to form the three-phase substances of water, gas, and solid. Dissolved air under pressure may be added to cause the formation of tiny bubbles which will attach to particles. Under the effect of interfacial tension, buoyancy of bubble rising, hydrostatic pressure and variety of other forces, the microbubble adheres to the tiny fibers. Due to its low density, the mixtures float to the surface so that the oil particles are separated from the water. So, this method can effectively

Coagulation flocculation sedimentation is one of the most used methods, especially in the

Active on suspended matter, colloidal type of very small size, their electrical charge give repulsion and prevent their aggregation. Adding in water electrolytic products such as aluminum sulphate, ferric sulphate, ferric chloride, giving hydrolysable metallic ions or organic hydrolysable polymers (polyelectrolyte) can eliminate the surface electrical charges of the colloids. This effect is named coagulation. Normally the colloids bring negative charges,so the coagulants are usually inorganic or organic cationic coagulants (with

The metallic hydroxides and the organic polymers, besides giving the coagulation, can help the particle aggregation into flocks, thereby increasing the sedimentation. The combined

Settling needs stillness and flow velocity, so these three processes need different reactions tanks. This processes use mechanical separation among heterogeneous matters, while the dissolved matter is not well removed (clariflocculation can eliminate a part of it by absorption into the flocks). The dissolved matter can be better removed by biological or by

But additional chemical load on the effluent (which normally increases salt concentration)

action of coagulation, flocculation and settling is named clariflocculation.

increases the sludge production and leads to the uncompleted dye removal.

other physical chemical processes (Sheng.H et al.,1997) .

of COD. The common physicochemical methods are shown as followed.

**2.1 Physicochemical wastewater treatment**

**2.1.1 Equalization and homogenization**

and stable operation.

**2.1.2 Floatation**

time is generally about 8 h.

remove the fibers in wastewater.

conventional treatment process.

positive charge in water).

**2.1.3 Coagulation flocculation sedimentation**

Chemical operations, as the name suggests, are those in which strictly chemical reactions occur, such as precipitation. Chemical treatment relies upon the chemical interactions of the contaminants we wish to remove from water, and the application of chemicals that either aid in the separation of contaminants from water, or assist in the destruction or neutralization of harmful effects associated with contaminants. Chemical treatment methods are applied both as stand-alone technologies and as an integral part of the treatment process with physical methods (K.Ranganathan et al.,2007).Chemical operations can oxidize the pigment in the printing and dyeing wastewater as well as bleaching the effluent. Currently, Fenton oxidation and ozone oxidation are often used in the wastewater treatment.

#### **<IP_ADDRESS> Fenton reaction**

Oxidative processes represent a widely used chemical method for the treatment of textile effluent, where decolourisation is the main concern. Among the oxidizing agents, the main chemical is hydrogen peroxide (H2O2), variously activated to form hydroxyl radicals, which are among the strongest existing oxidizing agents and are able to decolourise a wide range of dyes.

A first method to activate hydroxyl radical formation from H2O2 is the so called Fenton reaction, where hydrogen peroxide is added to an acidic solution (pH=2-3) containing Fe2+ ions.

Fenton reaction is mainly used as a pre-treatment for wastewater resistant to biological treatment or/and toxic to biomass. The reaction is exothermic and should take place at temperature higher than ambient. In large scale plants, however, the reaction is commonly carried out at ambient temperature using a large excess of iron as well as hydrogen peroxide. In such conditions ions do not act as catalyst and the great amount of total COD removed has to be mainly ascribed to the Fe(OH)3 co-precipitation. The main drawbacks of the method are the significant addition of acid and alkali to reach the required pH, the necessity to abate the residual iron concentration, too high for discharge in final effluent, and the related high sludge production (Sheng.H et al.,1997).

#### **<IP_ADDRESS> Ozone oxidation**

It is a very effective and fast decolourising treatment, which can easily break the double bonds present in most of the dyes. Ozonation can also inhibit or destroy the foaming properties of residual surfactants and it can oxidize a significant portion of COD.

Moreover, it can improve the biodegradability of those effluents which contain a high fraction of nonbiodegradable and toxic components through the conversion (by a limited oxidation) of recalcitrant pollutants into more easily biodegradable intermediates. As a further advantage, the treatment does increase neither the volume of wastewater nor the sludge mass.

Full scale applications are growing in number, mainly as final polishing treatment, generally requiring up-stream treatments such as at least filtration to reduce the suspended solids contents and improve the efficiency of decolourisation. Sodium hypochlorite has been widely used in the past as oxidizing agent. In textile effluent it initiates and accelerates azo bond cleavage. The negative effect is the release of carcinogenic aromatic amines and otherwise toxic molecules and, therefore, it should not be used (Sheng.H et al.,1997).

#### **2.1.5 Adsorption**

Adsorption is the most used method in physicochemical wastewater treatment, which can mix the wastewater and the porous material powder or granules, such as activated carbon and clay, or let the wastewater through its filter bed composed of granular materials. Through this method, pollutants in the wastewater are adsorbed and removed on the surface of the porous material or filter.

Commonly used adsorbents are activated carbon, silicon polymers and kaolin. Different adsorbents have selective adsorption of dyes. But, so far, activated carbon is still the best adsorbent of dye wastewater. The chroma can be removed 92.17% and COD can be reduced 91.15% in series adsorption reactors, which meet the wastewater standard in the textile industry and can be reused as the washing water. Because activated carbon has selection to adsorb dyes, it can effectively remove the water-soluble dyes in wastewater, such as reactive dyes, basic dyes and azo dyes, but it can't adsorb the suspended solids and insoluble dyes. Moreover, the activated carbon can not be directly used in the original textile dyeing wastewater treatment, while generally used in lower concentration of dye wastewater treatment or advanced treatment because of the high cost of regeneration.

#### **2.1.6 Membrane separation process**

Membrane separation process is the method that uses the membrane's micropores to filter and makes use of membrane's selective permeability to separate certain substances in wastewater. Currently, the membrane separation process is often used for treatment of dyeing wastewater mainly based on membrane pressure, such as reverse osmosis, ultrafiltration, nanofiltration and microfiltration. Membrane separation process is a new separation technology, with high separation efficiency, low energy consumption, easy operation, no pollution and so on. However, this technology is still not large-scale promoted because it has the limitation of requiring special equipment, and having high investment and the membrane fouling and so on (K.Ranganathan et al.,2007) .

#### **<IP_ADDRESS> Reverse osmosis**

Reverse osmosis membranes have a retention rate of 90% or more for most types of ionic compounds and produce a high quality of permeate. Decolorization and elimination of chemical auxiliaries in dye house wastewater can be carried out in a single step by reverse osmosis. Reverse osmosis permits the removal of all mineral salts, hydrolyzed reactive dyes, and chemical auxiliaries. It must be noted that higher the concentration of dissolved salt, the more important the osmotic pressure becomes; therefore, the greater the energy required for the separation process (B. Ramesh Babu et al.,2007) .

#### **<IP_ADDRESS> Nanofiltration**

Nanofiltration has been applied for the treatment of colored effluents from the textile industry. Its aperture is only about several nanometers, the retention molecular weaght by which is about 80-1000da, A combination of adsorption and nanofiltration can be adopted for the treatment of textile dye effluents. The adsorption step precedes nanofiltration, because this sequence decreases concentration polarization during the filtration process, which increases the process output. Nanofiltration membranes retain low molecular weight organic compounds, divalent ions, large monovalent ions, hydrolyzed reactive dyes, and dyeing auxiliaries. Harmful effects of high concentrations of dye and salts in dye house effluents have frequently been reported. In most published studies concerning dye house effluents, the concentration of mineral salts does not exceed 20 g/L, and the concentration of dyestuff does not exceed 1.5 g/L. Generally, the effluents are reconstituted with only one dye, and the volume studied is also low. The treatment of dyeing wastewater by nanofiltration represents one of the rare applications possible for the treatment of solutions with highly c oncentrated and complex solutions (B. Ramesh Babu et al.,2007).

A major problem is the accumulation of dissolved solids, which makes discharging the treated effluents into water streams impossible. Various research groups have tried to develop economically feasible technologies for effective treatment of dye effluents. Nanofiltration treatment as an alternative has been found to be fairly satisfactory. The technique is also favorable in terms of environmental regulating.

#### **<IP_ADDRESS> Ultrafiltration**

100 Advances in Treating Textile Effluent

Through this method, pollutants in the wastewater are adsorbed and removed on the

Commonly used adsorbents are activated carbon, silicon polymers and kaolin. Different adsorbents have selective adsorption of dyes. But, so far, activated carbon is still the best adsorbent of dye wastewater. The chroma can be removed 92.17% and COD can be reduced 91.15% in series adsorption reactors, which meet the wastewater standard in the textile industry and can be reused as the washing water. Because activated carbon has selection to adsorb dyes, it can effectively remove the water-soluble dyes in wastewater, such as reactive dyes, basic dyes and azo dyes, but it can't adsorb the suspended solids and insoluble dyes. Moreover, the activated carbon can not be directly used in the original textile dyeing wastewater treatment, while generally used in lower concentration of dye wastewater

Membrane separation process is the method that uses the membrane's micropores to filter and makes use of membrane's selective permeability to separate certain substances in wastewater. Currently, the membrane separation process is often used for treatment of dyeing wastewater mainly based on membrane pressure, such as reverse osmosis, ultrafiltration, nanofiltration and microfiltration. Membrane separation process is a new separation technology, with high separation efficiency, low energy consumption, easy operation, no pollution and so on. However, this technology is still not large-scale promoted because it has the limitation of requiring special equipment, and having high investment

Reverse osmosis membranes have a retention rate of 90% or more for most types of ionic compounds and produce a high quality of permeate. Decolorization and elimination of chemical auxiliaries in dye house wastewater can be carried out in a single step by reverse osmosis. Reverse osmosis permits the removal of all mineral salts, hydrolyzed reactive dyes, and chemical auxiliaries. It must be noted that higher the concentration of dissolved salt, the more important the osmotic pressure becomes; therefore, the greater the energy required for

Nanofiltration has been applied for the treatment of colored effluents from the textile industry. Its aperture is only about several nanometers, the retention molecular weaght by which is about 80-1000da, A combination of adsorption and nanofiltration can be adopted for the treatment of textile dye effluents. The adsorption step precedes nanofiltration, because this sequence decreases concentration polarization during the filtration process, which increases the process output. Nanofiltration membranes retain low molecular weight organic compounds, divalent ions, large monovalent ions, hydrolyzed reactive dyes, and dyeing auxiliaries. Harmful effects of high concentrations of dye and salts in dye house effluents have frequently been reported. In most published studies concerning dye house effluents, the concentration of mineral salts does not exceed 20 g/L, and the concentration of dyestuff does not exceed 1.5 g/L. Generally, the effluents are reconstituted with only one dye, and the volume studied is also low. The treatment of dyeing wastewater by nanofiltration represents one of the rare applications possible for the treatment of solutions

with highly c oncentrated and complex solutions (B. Ramesh Babu et al.,2007).

treatment or advanced treatment because of the high cost of regeneration.

and the membrane fouling and so on (K.Ranganathan et al.,2007) .

the separation process (B. Ramesh Babu et al.,2007) .

surface of the porous material or filter.

**2.1.6 Membrane separation process**

**<IP_ADDRESS> Reverse osmosis**

**<IP_ADDRESS> Nanofiltration**

Ultrafiltration whose aperture is only about 1nm-0.05μm, enables elimination of macromolecules and particles, but the elimination of polluting substances, such as dyes, is never complete. Even in the best of cases, the quality of the treated wastewater does not permit its reuse for sensitive processes, such as dyeing of textile. So the retention molecular weaght is range from 1000-300000da, . Rott and Minke (1999) emphasize that 40% of the water treated by ultrafiltration can be recycled to feed processes termed "minor" in the textile industry (rinsing, washing) in which salinity is not a problem. Ultrafiltration can only be used as a pretreatment for reverse osmosis or in combination with a biological reactor (B. Ramesh Babu et al.,2007) .

#### **<IP_ADDRESS> Microfiltration**

Microfiltration whose aperture is about 0.1-1μm is suitable for treating dye baths containing pigment dyes, as well as for subsequent rinsing baths. The chemicals used in dye bath, which are not filtered by microfiltration, will remain in the bath. Microfiltration can also be used as a pretreatment for nanofiltration or reverse osmosis (B. Ramesh Babu et al.,2007).

Textile wastewater contains large amounts of difficult biodegradable organic matter and inorganic. At present,many factories have adopted physicochemical treatment process. Some typical physicochemical treatment process is shown in Table 8 and Fig. 2 (Kangmei Zeng et al.,2005).

#### **2.2 Biological wastewater treatment method**

The biological process removes dissolved matter in a way similar to the self depuration but in a further and more efficient way than clariflocculation. The removal efficiency depends upon the ratio between organic load and the bio mass present in the oxidation tank, its temperature, and oxygen concentration.

The bio mass concentration can increase, by aeration the suspension effect but it is important not to reach a mixing energy that can destroy the flocks, because it can inhibit the following settling.

Normally, the biomass concentration ranges between 2500-4500 mg/l, oxygen about 2 mg/l. With aeration time till 24 hours the oxygen demand can be reduced till 99%.

According to the different oxygen demand, biological treatment methods can be divided into aerobic and anaerobic treatment. Because of high efficiency and wide application of the aerobic biological treatment, it naturally becomes the mainstream of biological treatment.

#### **2.2.1 Aerobic biological treatment**

According to the oxygen requirements of the different bacteria, the bacteria can be divided into aerobic bacteria, anaerobic bacteria and facultative bacteria. Aerobic biological treatment can purify the water with the help of aerobic bacteria and facultative bacteria in the aerobic environment. Aerobic biological treatment can be divided into two major categories: activated sludge process and biofilm process.

Table 8. Physicochemical treatment instance of textile dyeing wastewater

Table 8. Physicochemical treatment instance of textile dyeing wastewater

#### **<IP_ADDRESS> Activated sludge process**

Activated sludge is a kind of floc which is mainly comprised of many microorganisms, which has strong decomposition and adsorption of the organics, so it is called "activated sludge". The wastewater can be clarified and purified after the separation of activated sludge. Activated sludge process is based on the activated sludge whose main structure is the aeration tank.

Activated sludge Process is an effective method. As long as according to the scientific laws, after getting some experiences, higher removal efficiency can be got. In present treatment, the oxidation ditch and SBR process are the commonly used activated sludge process.

#### **Oxidation ditch process**

Oxidation ditch is a kind of biological wastewater treatment technology, which developed by the netherlands health engineering research institute in the 50's of last century. It is a variant of activated sludge, which is a special form of extended aeration. The oxidation ditch plan was shown in Fig. 3. And the biological sewage treatment process which mainly composed by the oxidation ditch, as was shown in Fig. 4.

#### Fig. 3. Oxidation ditch plan

Fig. 4. The oxidation ditch for the biological treatment process

FeCl4 H2O2 FeSO4

regulating tank pump Coagulation tank Sedimentation tank Oxidation tank Sedimentation tank

(e)

Activated sludge is a kind of floc which is mainly comprised of many microorganisms, which has strong decomposition and adsorption of the organics, so it is called "activated sludge". The wastewater can be clarified and purified after the separation of activated sludge. Activated sludge process is based on the activated sludge whose main structure is

Activated sludge Process is an effective method. As long as according to the scientific laws, after getting some experiences, higher removal efficiency can be got. In present treatment, the oxidation ditch and SBR process are the commonly used activated sludge process.

Oxidation ditch is a kind of biological wastewater treatment technology, which developed by the netherlands health engineering research institute in the 50's of last century. It is a variant of activated sludge, which is a special form of extended aeration. The oxidation ditch plan was shown in Fig. 3. And the biological sewage treatment process which mainly

Surface aerator

Effluent

Dehydration Equipment

Rotary Brush Aerator

settling tank)

Effluent(To the Secondary

Sludge pumping station

Secondary sedimentation tank

Fig. 2. The physicochemical treatment process in textile dyeing wastewater

composed by the oxidation ditch, as was shown in Fig. 4.

wastewater Rotary Brush Aerator

Fig. 4. The oxidation ditch for the biological treatment process

wastewater

Neutralisation effluent

wastewater

**<IP_ADDRESS> Activated sludge process**

the aeration tank.

**Oxidation ditch process**

Fig. 3. Oxidation ditch plan

The oxidation ditch is generally consisted of the ditch body, aeration equipment, equipment of the water in or out, diversion and mixing equipment. The shape of the ditch body is usually ring; it also can be rectangular, L-shaped, round or other shape. The side shape of the oxidation ditch generally is rectangular or trapezoidal. Wastewater, activated sludge, and various microorganisms are mixed in a continuous loop ditch in order to complete nitrification and denitrification. Oxidation ditch has the characteristics of completely mixed, plug-flow and oxidation tank. Since oxidation ditch has long hydraulic retention time(HRT), low organic loading and long sludge age, compared to conventional activated sludge process, the equalization tank, primary sedimentation tank, sludge digestion tank can be omitted. The secondary sedimentation tank also can be omitted in some processes. It has many characteristics, such as high degree of purification, impact resistance, stable, reliable, simple, easy operation and management, easy maintenance, low investment and energy consumption. Oxidation ditch forms aerobic zone, anoxic zone and the anaerobic zone in space, which has a good function of denitrification.

#### **Sequencing Batch Reactor Activated Sludge process**

Sequencing Batch Reactor Activated Sludge Process (SBR Process) is a reform process from activated sludge, which is a new operating mode. Its operation is mainly composed of five processes: ①inflow; ②reaction; ③sedimentation; ④outflow; ⑤standby. The reaction process plan has been showed in Fig. 5.

SBR treatment process not only has a high removal rate of COD, but also has a high removal efficiency of color. Compared to the traditional methods, using SBR process has the following advantages (Li-yan Fu et al.,2001):

i. It has great resistance to shock loading. The reserved water can effectively resist the impact of water and organic matter.

Fig. 5. Reaction process of SBR

Table 9. The removal efficiency of various indicators in this dyeing factory

There is a dyeing factory in Jiangsu, which produces 60,000 tons wastewater each year. The wastewater mainly comes from dyeing and rinse. The major pollutant in wastewater is seriously exceeded, such as COD, color and SS. The wastewater treatment process which is used in this factory has been shown in Fig.2.5. Through this process, its various indicators have been shown in table 9(Haidong Jia,2003).

Fig. 6. The wastewater treatment process based on SBR

#### **<IP_ADDRESS> Biofilm**

The biofilm process is a kind of biological treatment that making the numerous microorganisms to attach to some fixed object surface, while letting the wastewater flow on its surface to purify it by contact. The main types of the biofilm process are biological contact oxidation, rotating biological contactors and biological fluidized bed.

#### **Biological contact oxidation**

Biological contact oxidation is widely used in the dyeing wastewater treatment. The main feature of the process is to set fillers in the aeration tanks, so that it has the characteristics of activated sludge and biofilm. The wastewater in oxidation tank contains a certain amount of activated sludge, while the fillers are covered with a large number of biofilm. When the wastewater contact with the fillers, it can be purified under the function of aerobic microorganisms.