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Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent
Arabian Journal of Chemistry (2017) xxx, xxx–xxx
King Saud University
Arabian Journal of Chemistry www.ksu.edu.sa www.sciencedirect.com
ORIGINAL ARTICLE
Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent Raja Kumar a, Alok Sinha a,*, Gautam C. Mondal b, Reginald E. Masto c a
Department of Environmental Science & Engineering, Indian Institute of Technology (ISM), Dhanbad 826004, Jharkhand, India Central Institute of Mining and Fuel Research (CSIR-CIMFR), Barwa Road, Dhanbad 826001, Jharkhand, India c Environmental Management Division, Central Institute of Mining and Fuel Research (Digwadih Campus), PO: FRI, Dhanbad 828108, India b
Received 11 November 2016; accepted 4 March 2017
KEYWORDS Adsorption; Biodegradability; Direct green 1 (DG1); Reduction; Scrap cast iron particles (SIP)
Abstract This study proposed that hybrid scrap cast iron particles (SIP)-aerobic biodegradation technology could enhance the biodegradability of toxic wastewater. SIP cleaved the azo linkages of Direct Green1 dye to form benzidine, 4-aminophenol, aniline and 1,2,7-triamino-8-hydroxynap thalene-3,6-disulfonic acid. SIP-mediated dye reduction was effective at wide pH range; however, kinetic analysis revealed fastest pseudo-first order dye reduction rate at acidic pH 3 (kd = 0.549 min1) followed by pH 9 (kd = 0.383 min1) and pH 7 (kd = 0.318 min1). The daughter aromatic amines produced were partially adsorbed onto the SIP surface and maximally at neutral pH. The adsorption process followed pseudo-second order adsorption kinetics and Langmuir isotherm. Benzidine was adsorbed more than 4-aminophenol and aniline. BOD5 of the SIP-treated effluent increased from 0.93 to 12 mg/L showing improved biodegradability. The daughter amines were rapidly mineralized in the aerobic bioreactor within 6 h. Cost-effective SIP pre-treatment could accelerate mineralization and detoxification of recalcitrant wastewater. Ó 2017 The Authors. Production and hosting by Elsevier B.V. on behalf of King Saud University. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
1. Introduction * Corresponding author. E-mail addresses: [email protected] (R. Kumar), [email protected] (A. Sinha), [email protected] (G.C. Mondal), [email protected] (R.E. Masto). Peer review under responsibility of King Saud University.
Production and hosting by Elsevier
Synthetic dyes have been extensively used in textile, power loom and dying industries. Accumulation of these dyestuff and dye wastewater creates not only environmental pollution, but also medical and aesthetic problems. More than 500 azo dyes that are based on carcinogenic amines are also recognized as hazardous contaminants and must be removed from waste effluents prior to discharge (Bafana et al., 2008). Besides, dyes wastes can also generate dangerous aromatic amine by-products through chemical transformation taking place in the aquatic-phase which are often more harmful to the environment than the parent dye molecule (Alves de Lima et al., 2007). These
http://dx.doi.org/10.1016/j.arabjc.2017.03.001 1878-5352 Ó 2017 The Authors. Production and hosting by Elsevier B.V. on behalf of King Saud University. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/). Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
2 amines have potential to persist and their physical properties make them highly mobile contaminants in reducing aquifers and sediments (Elovitz and Weber, 1999). Several aromatic amines, such as aniline (AN), 4-aminophenol (4-AP) and benzidine (BZN), are highly toxic chemicals which can lead to carcinogenesis, teratogenesis, and mutagenesis (Jurado-Sanchez et al., 2012). Many of them have been listed as priority pollutants by the US EPA, EU, and China EPA. Hence, it is necessary to contain azo dyes and their daughter amines, either by physico-chemical or by microbial routes. Available literature reports scores of physical, chemical and biological technologies evolved so far for azo dye effluent treatment. Unfortunately, the practical application of most of the reported technologies is limited by low efficiency, longer retention time, difficulty of operation, unfavourable economics, production of large volume of sludge, generation of secondary waste products and environmental incompatibility (Robinson et al., 2001; Crini, 2006). Biological treatment is considered as cheapest alternative for remediation of most of the wastewaters. However, aerobic biodegradation systems are not successful for decolorizing the azo dye-containing effluents because of their strong electron-withdrawing azo linkages (AN‚NA), which protects them from attack by the oxygenases (Singh and Arora, 2011). Anaerobic biological treatment of azo dye has been investigated by several researchers (Fernando et al., 2013; Quan et al., 2015) but this process can only decolorize azo dye (by splitting the azo bond) while leaving behind potentially toxic aromatic amines generated from the reaction. The fact that these aromatic amines can be degraded by aerobic microflora (Fernando et al., 2014) has drawn attention towards applicability of combined anaerobic–aerobic biological methods for the treatment of azo dye-containing effluent. However, a major limitation to any biological technique is longer treatment time. To overcome this problem a rapid dye degradation technology is sought after which should be cheap and environmentally sustainable. Zero valent iron (ZVI) seems to be the most appropriate replacement for the anaerobic biological pre-treatment. ZVI is an electron donor, a strong reducing agent, and the spent agent, Fe2+ is environmentally innocuous. Azo dye decolourization studies performed using ZVI have shown very rapid dye reduction at varied experimental conditions (Mielczarski et al., 2005; Kumar and Sinha, 2013). Combination of ZVI based dye reduction technology with aerobic biological treatment can achieve complete mineralization of azo dye solutions, that too in much quicker time. Therefore, in the present study, a hybrid method combining the reducing effect of ZVI to discolour the azo dye effluent, making it amenable for subsequent aerobic biodegradation step, was investigated. The form of ZVI used was scrap cast iron particles (SIPs) which are available in plenty at much lower cost compared to pure electrolytic iron. This study explores the efficiency of SIPs in decolourization of azo dye. Available literature is deficient of information on the interaction of azo dyes and their reductive transformation products with metallic iron. Both qualitative and quantitative analyses of formation and adsorption of daughter aromatic amines have been done. The iron-treated synthetic solution has been screened for amenability to biological treatment. To this end, this study may provide new insight into the interaction of metallic iron with aqueous contaminants and may help in the development of an integrated detoxification process for complete remediation of textile dye-bath effluent.
R. Kumar et al. without any purification. Analytical grade benzidine (BZN), aniline (ANI) and 4-aminophenol (4-AP) were procured from Sigma–Aldrich (Germany), Merck (USA), and Acros Organics (Belgium), respectively. HPLC grade methanol used as an HPLC eluent was purchased from Merck. Deionised water was used throughout the whole experimental process. 2.2. Preparation and characterization of SIPs The SIPs used in this research were prepared using cast iron chips obtained from a local foundry. Cast iron chips are generated while cutting the edges of casted billets in order to make its surface smooth. The collected cast iron chips were ground into smaller iron particles in a dough-size ball mill. The particles were sieved and those passing through sieve #40 and remaining on sieve #80 were retained for use. These particles were washed 5 times with N2-sparged 1 N HCl with periodic shaking. The SIPs, thus obtained, were rinsed 8–10 times with N2-sparged deionised water; instantly rinsed twice with methanol and dried for an hour under anaerobic condition. The SIPs were then stored in a dessicator filled with nitrogen gas. X-ray diffraction (XRD) analyses (data not shown) of the SIP surface carried out using X-ray Powder Diffractometer (Bruker Advanced X-ray Solutions, GMBH, Germany) indicated that the SIP surface was devoid of any oxide coating. The carbon content (per cent weight basis) of the SIPs as determined by LECO Induction Furnace Instrument (LECO Corporation, St. Joseph, USA) was found to be 3.08%. Specific surface area of the SIPs was determined by BET (Brunauer Emmett Teller) surface area analyzer (NOVA 4000e, Quantachrome Instruments, USA) to be 2.16 m2/g. 2.3. Batch reduction experiments The reduction in Direct Green 1 (DG1) azo dye by SIPs in aqueous solution was investigated thoroughly in batch and column experiments. All the experiments were carried out in 35 mL capacity borosilicate glass vials bearing screw caps (Borosil, India). The concentration of DG1 dye and SIPs was pre-optimized and was kept uniform at 100 mg/L and 57.12 g/L, respectively, for all the experiments. The desired initial pH value in aqueous solution was adjusted with 0.1 M HCl or 0.1 M NaOH. The vials were properly shaken on an end-toend test tube rotator (Rotospin, Tarsons, India) at 30 rpm in room temperature (25 ± 2 C). With every test, a SIP-free dye solution was run as a control. The vials were periodically withdrawn from the rotator in duplicate along with a control for sampling and analysis, and were put on a magnet to allow iron particles to settle down quickly. The supernatants from each vial were filtered through GF/C filter paper (1.2 mm nominal pore size, Whatman, Springfield Mill, England) and the filtrate was quantified for colour and reduction by-products.
2. Materials and methods 2.4. Desorption of reduction by-products adsorbed onto the SIPs 2.1. Chemicals and reagents Commercially available tris-azo dye C.I. Direct Green 1 (2,7Naphthalenedisulfonic acid, 4-amino-5-hydroxy-3-((40 -((4-hyd roxyphenyl) azo) (1,10 -biphenyl)-4-yl)azo)-6-(phenylazo)-, disodium salt; M.F.: C34H23N7Na2O8S2; M.W.: 767.70) was obtained from Mac Dye Chem, Ahmedabad (India), and used
To extract the DG1 reduction by-products adsorbed to the SIP surface, the treated dye solution of the 35 mL reaction vial was transferred to another vial. To the vial containing the used SIPs, 5 mL methanol was added and the contents were vortex mixed for 10 min. The solvent along with desorbed compounds was transferred to another vial. Again 5 mL of methanol was
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
Complete mineralization of benzidine based azo dye effluent added to the solid phase and the contents were mixed for another 10 min to ensure desorption of any residual compounds. This solvent was also transferred to the same vial. The 10 mL of methanol, thus collected, was filtered through 0.22 mm filter disc. The filtrate was used for qualitative and quantitative analysis of adsorbed daughter products. 2.5. Hybrid scrap iron-aerobic biodegradation treatment system To evaluate the feasibility of treatment of DG1 dye-containing wastewater in a continuous flow through system and to assess the biodegradability of daughter products after SIP pretreatment, a hybrid textile effluent treatment system (as shown in Fig. S1, Supplementary Information) was constructed. The pre-treatment column was fabricated using 4 mm thick transparent acrylic plastic tubing (60 cm L 2.5 cm i.d.) provided with an inlet and an outlet. The column was loaded with SIPs and well rinsed sterilized sand (20–30 mesh size) mixture in 1:1 ratio. Column end caps, valves, and tubing were made of polypropylene and a variable speed peristaltic pump (Rivotek, India) was used to pump the solution through the columns at a flow rate of 10 mL/min in an up-flow mode to prevent gas entrainment. Aerobic biodegradation studies were carried out in 4 L batch reactors, with a working volume of 2 L, where constant air flow was maintained using aquarium air pumps and porous stone diffusers. The bioreactor was seeded with activated sludge collected from Durgapur Coke oven effluent Treatment Plant, West Bengal, India. The reactors were initially fed with synthetic wastewater containing COD concentration of 1000 mg/L along with 1 mL/L trace metal solution (Sharan, 2011) (The detailed characteristics of the feed and composition of trace metal solution are given in Tables A1 and A2, respectively, in Supplementary Information). Both the reactors were operated for 3–4 weeks under room temperature (25 ± 2 °C) until the steady state conditions were established. Before exposing the microbial flora to the SIP-treated effluent, the bioreactors were acclimatized to the aromatic amines detected after reductive degradation of DG1. During operation, the treated effluent was collected from the exit port of the SIP/ sand column and was analysed by HPLC for the quantification of degradation by-products formed and then was transferred to the aerobic bioreactor. After 24 h of HRT in the bioreactor, the hybrid system effluent was quantified. 2.6. Analytical procedure Preliminary experiments were carried out at solution pH 3 using 57.12 g/L SIP dose to understand DG1 degradation mechanism and to identify the reduction by-products. The surface morphology and composition of SIP surface during the course of DG1 degradation reaction were studied using FESEM (Field Emission Scanning Electron Microscopy) and Energy dispersive spectra (EDS) and FTIR, respectively. UV–Vis spectral analysis and LC-MS/MS were used for qualitative analysis of dye degradation and reduction by-products formed. For kinetic studies, colour removal and concentration of by-products were analysed using UV–Vis spectrophotometer and HPLC, respectively. The details of analytical procedure and conditions using each instrument are given in the Supplementary Information.
3 3. Results and discussion 3.1. Interaction of azo dye with SIP surface: characterization of fresh and used SIP surface Fig. 1(a) –(c) shows the SEM images of fresh SIP surface, SIP surface after 10 min of reaction and SIP surface after 120 min of reaction, respectively. Fresh SIP surface was free from rust; hence, no peak for oxygen was observed in its EDS spectra (Fig. S2-a, Supplementary information). Also, it showed highest iron content of 93.73% (see Table A3, Supplementary information) and was free from any surface deposition. SEM image of SIP taken after 10 min of reaction (Fig. 1b) shows large amount of deposition on its surface which could be adsorbed dye molecules. This statement is supported by the corresponding EDS spectra (Fig. S2-b, Supplementary information) which shows increase in carbon (17.63%) and sulphur content (1.41%) indicating towards the adsorption of sulphonated organic compound, that could be DG1 dye molecule and/or its reduction by-product. After 120 min of reaction, when decolourization was almost complete, the SIP surface exhibited rusted morphology with some amount of surface deposition (Fig. 1c). The carbon (9.74%) and sulphur (1%) content decreased marginally (Fig. S2-c, Supplementary information), but not completely, suggesting adsorption of degradation by-products onto the rusted surface. 3.1.1. FTIR spectra analysis of SIP surface Fig. 2 illustrates the FTIR spectra of fresh SIP surface and used SIP surface (sampled after 10 min and 120 min of reaction time). FTIR spectroscopy of SIPs performed after 10 min of reaction shows an intense band at 1514 cm1, attributable to the azo-chromophore (AN‚NA) bond vibrations of adsorbed unreduced dye molecule. This band was absent in CIP sampled after 120 min of reaction suggesting breaking of azo bonds. The broad peak at 3000–3500 cm1 is due to hydration of rust and is formed due to OAH stretching (Balasubramaniam and Kumar, 2000). The medium-weak, multiple bands between 1400 and 1600 cm1 were assigned to C‚C aromatic skeleton vibrations, while the band at 1403 cm1 was ascribed to the OAH bending vibrations and deformation (Bandara et al., 1999). Strong bands at 1104, 1156 and 1207 cm1 were assigned to the symmetric and asymmetric stretching of sulphonate anions (ASO 3 ) of naphthalene ring (Hou et al., 2007). The weak frequencies at 1360 cm1 were assigned to the CAN stretching due to the resonance effect of the aromatic ring and those between 1617 and 1680 cm1 were attributable to the in-plane bending vibrations for NAH of aromatic amines. These bands were found in all the used SIP samples. This suggests fast reduction in DG1 and adsorption of its reduction by-products bearing ASO3H and ANH2 functional groups onto the SIP surface. 3.2. Identification of reduced and adsorbed daughter products 3.2.1. UV–Vis spectral analysis The molecular structure of DG1 has three azo groups. The chromophore containing azo linkage in the DG1 dye molecule has absorption in the visible region (k = 603.5 nm), while the characteristic wavelength of benzene-ring (k = 234.5 nm) and
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
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Figure 1 Scanning electron micrographs of (a) fresh SIP surface, (b) SIP surface after 10 min of reaction, and (c) SIP surface after 120 min of reaction.
Figure 2 FTIR spectra of SIP surface employed for Direct Green 1 degradation.
naphthalene ring (k = 376 nm) in the UV region (Fig. 3a). During the decolourization process, the absorbance band at 603.5 nm disappeared completely within 20 min of the reduction reaction using SIPs. Also, the absorbance bands at 234.5 nm and 376 nm disappeared; instead, a new peak appeared at 280 nm which became more prominent with reaction progress. This peak is attributable to the aromatic ring structures of BZN and 4-AP (Min et al., 2007; Azetsu et al., 2011) and has been matched with their reference spectrum
(Fig. S3, Supplementary information). ANI was also expected to form due to breaking of one of the azo bonds but no separate peak was observed at 232 nm wavelength. This could be either due to adsorption of ANI onto SIPs and/or due to expectedly low stoichiometric formation of ANI. However, the secondary absorbance peak of ANI is also reported to form at 280 nm, which is observed in the spectra. On the basis of the sole UV–VIS spectra, no other species were detected. Fig. 3b shows the UV–VIS absorbance spectra of DG1 dye desorbed from SIPs using methanol at definite time intervals of 10 min, 30 min and 120 min. In this case, two peaks, i.e. at 603.5 nm and 376 nm, of the major peaks found in the parent dye molecule were found in the solution desorbed after first 10 min. The initial peak found earlier at 234.5 nm shifted towards shorter wavelength (at 227 nm) when the p bonds of characteristic conjugated chromophores in an azo dye molecule were broken (Skoog and Leary, 1992). The new peak at 280 nm was also observed to form in solution desorbed after 10 min of reductive treatment. All the peaks observed in parent compound vanished after 120 min whereas the new peak at 280 nm, which was attributed to aromatic amine formation, increased in height indicating towards increase in their adsorbed concentration. The peak for ANI was formed at the wavelength of 232 nm. The intensity of the peak was low due to the low concentration of adsorbed ANI. 3.2.2. Liquid chromatography and mass spectrometric analysis The HPLC chromatograms of SIP treated aqueous phase and methanol extracted solid phase showed three prominent peaks for daughter products at 3.2, 3.6, and 4.2 min retention times (Fig.S4, Supplementary information). Fig. 4(a) and (b) shows
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
Complete mineralization of benzidine based azo dye effluent
5
(b) Figure 3 UV–VIS spectra of (a) treated aqueous phase DG1 solution as a function of reaction time (b) methanol extracted solid phase at definite reaction time (Initial conditions: pH 3; dye conc. 100 mg/L; SIP dose 57.12 g/L; mixing speed 30 rpm).
the mass spectra of the SIP-treated DG1 aqueous solution recorded after 10 min and 120 min of reaction intervals, respectively. The spectra exhibit three major species with m/z ratio of 185, 109 and 94, of which former was the most abundant. These positive ions were attributable to BZN, 4-AP and ANI, respectively, which were formed by the reductive cleavage of the azo linkages of DG1 molecule. Their identification was further validated by the MS/MS fragmentation pattern of BZN, 4-AP and ANI (Fig.S5, Supplementary information) that matched their respective NIST library record. It is confirmed that only these three daughter products were formed after DG1 reduction. The other expected daughter product,
namely 1,2,7-triamino-8-hydroxynapthalene-3,6-disulfonic acid (TAHNDS), could not be identified. The mass spectrum of dye solution desorbed from SIP surface using methanol after 120 min also exhibits same three by-products (Fig. 4c) of which species with m/z 185 was the most abundant, indicating its higher degree of adsorption. LC-MS analysis could confirm only these three daughter amines formed after DG1 reduction. However, SEM-EDS and FTIR analysis reveals adsorption of a species bearing –SO3H functional group onto the SIP surface. This could be the other expected daughter product, namely TAHNDS. The proposed mechanism involved in the transformation of DG1 by SIP is represented in Fig. 4(d).
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
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Figure 4 ESI mass spectrum of solution during reductive degradation of DG1, (a) after 10 min, (b) after 120 min of reaction, (c) desorbed from SIP surface after 120 min of reaction (initial DG1 concentration; 100 mg/L; SIP dosage: 57.12 g/L; pH 3.0), (d) proposed pathway of Direct Green 1 reductive transformation (SIP:scrap iron particle; 4-AP: 4-aminophenol; ANI: aniline, BZN: benzidine, TAHNDS:1,2,7-triamino-8-hydroxynapthalene-3,6-disulfonic acid).
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
Complete mineralization of benzidine based azo dye effluent 3.3. Kinetics of DG1 reduction and formation of aromatic amines In water or wastewater treatment, pH of the aqueous solution is a complex parameter since it is related to the ionization state of the metal surface and that of the reactants and products. Hence, kinetics of degradation of DG1 and formation of its daughter by-products were investigated at three levels of pH 3, 7, and 9 (representing acidic, neutral, and alkaline conditions, respectively). Dye degradation by metallic iron is assumed to be pseudo-first order reaction; hence, the rate of degradation of DG1 and the rate of formation of daughter aromatic amines were evaluated by nonlinear regression of the experimental data using Eqs. (1) and (2), respectively. Ct ¼ C0 ekd t
ð1Þ
Ct ¼ C0 ð1 ekf t Þ
ð2Þ
where Ct (mg/L) is the concentration of analyte at time t (min), C0 (mg/L) is the initial or final concentration, kd (min1) is the pseudo-first order DG1 degradation rate coefficient and kf (min1) is the rate coefficient of formation of daughter aromatic amines. The temporal variation in the concentration of DG1 and simultaneous increase in the total concentration of its reduction by-products at pH 3,7 and 9 are exhibited in Fig. 5(a) (b) and (c), respectively, and the result of kinetic analysis is shown in Table 1. The final DG1 degradation efficiencies recorded after 120 min at pH 3, 7 and 9 were 99.82%, 99.35% and 99.50%, respectively. The respective final concentration of BZN, 4-APl and ANI was 13.82 mg/L and 8.57 mg/ L and 7.2 mg/L at pH 3; 11.87 mg/L, 7.25 mg/L and 7.12 mg/L at pH 7; 13.40 mg/L and 8.25 mg/L and 7.17 mg/L at pH 9. The rate of degradation of DG1 followed the order (from the highest to lowest) pH 3 > pH 9 > pH 7. At acidic pH 3, H+ would obviously promote the corrosion of SIPs providing more amount of Fe2+ to the reaction. This could account for a higher rate of DG1 degradation and appearance of daughter amines. At alkaline pH 9, OH would significantly enhance the formation of the iron hydroxides and oxy-hydroxides. It is reported that the iron oxide(s) can act as semi-conductor to relay electron transfer from iron to the contaminants (Huang and Zhang, 2006). Additionally, the oxides in Fe0/ H2O systems are good adsorbents for FeII originating from iron oxidation yielding adsorbed FeII (structural FeII). Structural FeII is a very strong reducing agent for inorganics and organics (Geetha and Surender, 1994). This could be the reason for the dye getting reduced at pH 9 and stoichiometrically higher concentration of daughter amines formed compared to that at pH 7. Reduction in DG1 dye at pH 9 and formation of high amount of by-products contradict the literature reports that claim removal of dye molecules from aqueous phase by adsorption and/or co-precipitation at alkaline pH, regardless of redox activity (Noubactep, 2009). Low dye reduction at pH 7 could be because of slower iron corrosion (Johnson and Tratnyek, 1995); hence, neither much of the rejuvenated iron surface and H+ were present (as in case of acidic conditions) nor was the oxide and hydroxide coated adsorptive surface area (as in case of alkaline conditions) (Kumar and Sinha, 2016). The rate of appearance of ANI and 4-AP was faster than that of BZN. This could be because these two amines were
7 formed due to splitting of single azo bonds present at the either end of DG1 molecule, which were supposed to lie in closer proximity to the iron surface thus facilitating faster sharing of electrons. On the other hand, the formation of BZN required breaking of two azo bonds one of which lied deep in the DG1 structure such that it’s splitting could get delayed. Mass balance (%) evaluated for these three by-products (considering their stoichiometric, aqueous and adsorbed concentration) at different solution pH was between 55% and 60% (see Table 1). The purity of DG1 used in this study was approximately 60%; therefore, the obtained mass balance was sustainable 3.4. Simultaneous adsorption of daughter aromatic amines: Isotherm and kinetics Considerable amount of daughter aromatic amines was found to adsorb onto the SIP surface. Hence, the interactive behaviour of aromatic amines with SIP was studied at different pH. The feasibility of the adsorption process is predicted using adsorption isotherm, which also predicts the distribution of adsorbate molecules between the liquid and solid phase at equilibrium. The experimental isotherm data were fitted using two most common adsorption models, i.e. the Langmuir model (corresponding to monolayer homogeneous adsorbent surface) and the Freundlich model (corresponding to heterogeneous adsorbent surface). The details of the adsorption isotherms are given in the Supplementary information. Fig. S6 (in Supplementary information) shows the applicability of both models at different pH and the Freundlich and Langmuir parameters as calculated using the nonlinear regression analysis are reported in Table 2. Based on the higher correlation coefficient (r2) and lower root mean square error (RMSE) value (Table 2), it is evident that Langmuir isotherm model shows better mathematical fit with experimental data. This suggests that the adsorption is monolayer at different pH. The value of Langmuir constant ‘KL’ followed the descending order: pH 7 > pH 9 > pH 3. This specifies the stronger affinity between the adsorption site and aromatic amines at neutral pH. In order to analyse the adsorption kinetics of aromatic amines, correlations between their adsorbed amounts and time were looked for, through the nonlinear regression of different mathematical expressions corresponding to pseudo-first-order equation and pseudo-second-order reaction. The details of the kinetic models are given in the Supporting Information. The kinetics data of aromatic amine adsorption studied at different pH (Fig. S7, Supporting Information) fitted best into pseudo-second-order model (based on better r2 and lower RMSE, see Table 3), which is based on the assumption that the rate limiting step may be chemisorption involving valency forces through sharing or exchange of electrons between the functional groups of the amines and the SIP surface. 3.5. Discussion on the adsorption of aromatic amines Comparative analysis of the data for sorption of aromatic amines onto the SIP surface reveals that sorption of BZN per gram SIP was the highest followed by 4-AP and ANI. This observation can be rationalized on the basis of aqueous solvation properties, structural and electrochemical properties of the
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
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Figure 5 Temporal variation in the concentration of DG1, daughter aromatic amines and stoichiometric mass balance for the daughter aromatic amines at (a) pH 3, (b) pH 7 and (c) pH 9.
three amines. The aqueous solubility of the three amines at 25 °C follows the descending order: BZN (520 mg/L) > 4-AP (15 g/L) > ANI (36 gm/L). Lesser the solubility of a compound, greater is the exclusion from bulk solution and higher is the extent of adsorption. However, the reported values of
octanol-water partition coefficients (log Kow) for the three amines follow a different trend, i.e. BZN (1.34) > ANI (0.90) > 4-AP (0.04). The value of log Kow increases as the compounds become more hydrophobic; therefore, highest BZN adsorption is primarily due to hydrophobicity. Although
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
Complete mineralization of benzidine based azo dye effluent Table 1
9
Kinetic parameters for the degradation of DG1 and formation of daughter aromatic amines.
pH
Dye/Amines
kd (1/min)
r2
3
DG1 BZN 4-AP ANI
0.549
0.988
DG1 BZN 4-AP ANI
0.318
DG1 BZN 4-AP ANI
0.383
7
9
kf (1/min)
r2
Mass balance (%)
0.284 0.347 0.342
0.990 0.984 0.966
58.10 59.70 59.88
0.187 0.230 0.228
0.989 0.944 0.935
57.98 57.68 59.70
0.211 0.286 0.265
0.967 0.948 0.941
56.50 59.18 59.22
0.985
0.993
Table 2
Parameters of Langmuir and Freundlich isotherms for adsorption of daughter aromatic amines.
pH
Amines
Langmuir isotherm
Freundlich isotherm 2
qm (mg/g)
KL (L/mg)
r
RMSE
KF (mg/g)
n
r2
RMSE
3
BZN 4-AP ANI
0.1080 0.0606 0.0535
0.111 0.243 0.298
0.993 0.996 0.990
0.775 0.851 2.186
0.015 0.012 0.010
0.580 0.557 0.560
0.991 0.989 0.981
1.225 1.406 2.235
7
BZN 4-AP ANI
0.0866 0.0550 0.0437
0.336 0.400 0.435
0.987 0.992 0.995
0.890 0.518 0.640
0.026 0.019 0.016
0.401 0.404 0.421
0.975 0.991 0.984
1.519 0.696 0.881
9
BZN 4-AP ANI
0.0831 0.0520 0.0415
0.204 0.334 0.236
0.992 0.988 0.975
1.020 0.823 1.680
0.024 0.017 0.011
0.419 0.471 0.489
0.982 0.985 0.969
1.257 0.956 1.994
Table 3
Kinetic parameters for the adsorption of daughter aromatic amines.
pH
Amines
Pseudo-first-order
Pseudo-second-order
K1 (1/min)
qe (mg/g)
r2
RMSE
K2 (g/mg /min)
qe (mg/g)
r2
RMSE
3
BZN 4-AP ANI
0.443 0.412 0.388
0.062 0.035 0.027
0.953 0.969 0.977
1.43 1.25 1.91
10.712 17.430 18.819
0.065 0.036 0.028
0.987 0.980 0.994
1.34 0.61 0.85
7
BZN 4-AP ANI
0.522 0.485 0.422
0.064 0.039 0.030
0.957 0.979 0.960
1.82 1.59 2.10
13.272 28.254 35.354
0.067 0.041 0.032
0.976 0.993 0.984
0.68 0.79 1.23
9
BZN 4-AP ANI
0.490 0.426 0.410
0.061 0.036 0.028
0.959 0.973 0.966
2.00 1.82 4.05
12.457 24.290 24.297
0.064 0.038 0.029
0.979 0.992 0.988
0.88 0.81 3.21
log Kow of ANI is greater than 4-AP, the observed adsorption of 4-AP is higher. This is because AOH group in 4-AP possesses a negative charge and a high charge-to-radius ratio that yields a large ionic contribution to complex formation. In contrast, the ANH2 group (present in both 4-AP and ANI) pos-
sesses a neutral formal charge, and the ionic contribution to bonding is limited to the interaction between ionic charge on the metal surface and the dipole moment of the amino group. Since ion-dipole interactions are known to be weak, the ANH2 group has a lower contribution to ionic bonding (Vasudevan
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
10 and Stone, 1996). Also, compared to ANI that possess only one ANH2 group, the presence of both AOH and ANH2 group in 4-AP is likely to exert stronger cumulative effect in complex formation with SIP surface, hence the greater degree of adsorption. The sorption of aromatic amines observed in the experiment is seen to increase with time, possibly because of the simultaneous increase in oxidation of SIP surface, as revealed by SEM-EDS and FTIR analysis. Previous researches performed using cast iron have reported the partitioning of contaminants onto graphitic inclusions (Sinha and Bose, 2009, 2011, 2014; Lama et al., 2015). Hence, adsorption of aromatic amines onto the graphitic inclusions present in the SIP matrix also cannot be ruled out. However, such speculations required experimental validation which is not the scope of current research. Whatever be the adsorbent, it is likely to get influenced by solution pH that affects the surface charge of adsorbents as well as the degree of ionization of the adsorbate in the solution. In this study, highest adsorption was observed at pH 7 followed by pH 9 and pH 3. Based on the literature, iron (hydr)oxides have pKa1 and pKa2 values of 5 and 10 (Cornell and Schwertmann, 2006), respectively, resulting a pHzpc on the order of 8.1. At pH < pHzpc, the SIP surface has a net positive charge; at pH > pHzpc, the surface has a net negative charge. This could facilitate adsorption of neutral and positively charged aromatic amines at pH 7 and pH 9 systems. This adsorption behaviour is in agreement with those reported in the literature for polar aromatic amines where maximum adsorption occurs around pH 7 (Vasudevan and Stone, 1996; Bandara et al., 2001; Shanker et al., 2013). At pH 3, initially the adsorption was slow possibly because of increased dissolution of iron surface. However, adsorption increased with time as the solution pH approached towards neutral values (refer temporal pH variation profiles shown in Fig. S8, Supplementary information). Hence, the observed adsorption trend of aromatic amines here in this study follows the descending order: BZN > 4-AP > ANI. This order also exhibits the descending basic nature of these amines, conferred upon them by the number of electron-donating substituents attached to the aromatic structure, which reflects the lesser availability of electrons for the interaction.
R. Kumar et al. reactor, it was subjected to a biological treatment step. Fig. 6(a), (b) and (c) shows the respective concentration profile of BZN, 4-APl and ANI in the column reactor effluent and the final effluent from aerobic bioreactor. Complete degradation of DG1 was observed in the SIP column reactor. However, a large amount of degradation by-products were found to remain adsorb within the column reactor. The concentration of BZN in the column effluent was between 0.5 and 0.9 mg/ L as against its expected stoichiometric concentration of 14.40 mg/L. ANI concentration in the column effluent was between 2 and 3 mg/L as against its expected stoichiometric concentration of 7.26 mg/L. The concentration of 4-AP in the effluent was 3.0–4.5 mg/L while its expected stoichiometric concentration was 8.52 mg/L. This suggests that ANI and 4AP were partially adsorbed through the column while BZN was contained largely within the column by adsorption. Careful analysis of DOC data (Fig. S10, Supporting Information) and accounting for the dye purity, suggest substantial adsorption of TAHNDS within the column reactor. DOC of the
3.6. Biodegradability assessment The preliminary biodegradability studies performed for batch reactor effluent on the basis of BOD5 were encouraging and formed the basis for the subsequent hybrid treatment. The BOD5 of untreated DG1 solution was 0.93 mg/L which indicates that raw dye containing effluent inhibits the respiratory activity of bacterial seed. Therefore, besides being difficult to biodegrade, the dye containing effluent can also decrease the efficiency of biological treatment process. After 120 min of SIP treatment in batch system, the BOD5 increased to 12 mg/L (refer Fig. S9, Supplementary information). Evidently, the reductive splitting of azo linkages of DG1 by SIP forms simpler fragments that easily accept electrons and are more amenable to subsequent aerobic biological treatment process. In the hybrid treatment process, after the DG1 dyecontaining effluent was pre-treated with SIP in the column
Figure 6 Concentration of (a) benzidine, (b) 4-aminophenol, and (c) aniline in the effluent from SIP column and the aerobic bioreactor.
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Complete mineralization of benzidine based azo dye effluent influent DG1 wastewater passed through SIP column was 30.8 mg/L while the column effluent had DOC of 5– 9 mg/L. Calculated DOC of TAHNDS (i.e. 9.4 mg/L for 60% pure dye) is higher than the effluent DOC, thus pointing towards its adsorption. In the bioreactor, all of the observed aromatic amines generated from SIP column were completely mineralized; however, DOC removal was only around 62%. Assuming complete mineralization of glucose and acetate (present in feed), the residual DOC could be attributed to TAHNDS, since the microbes were not pre-acclimatized to this compound. No trace of other amines was found in the samples analysed even after 6 h of aerobic treatment, indicating their rapid biodegradation. Morphological study of activated sludge biomass in the bioreactor reveals the predominance of rodshaped bacilli (Fig. S11, Supplementary Information). Many researchers have utilized rod-shaped bacterial species such as Pseudomonas sp. (Kumar et al., 2006; Barsing et al., 2011); Shewanella sp. (Cai et al., 2012) and Burkholderia sp. (Takenaka et al., 2003) for the treatment of 4-aminophenol, benzidine and aniline. The above observations clearly demonstrate that the pre-treatment with SIP reduced the toxicity of DG1 and was capable of stimulating rapid mineralization of the residual pollutant. 4. Conclusions Scrap cast iron particles (SIP) showed enormous potential towards rapid enhancing of biodegradability of textile wastewater. SIP effectively cleaved azo bonds of DG1 to form bio-available aromatic amines. Dye degradation was highest at acidic pH whereas the concurrent adsorption of degradation by-products was maximal at neutral pH. Biodegradability assessment revealed significant increase in the BOD5 of the SIP-treated dye solution within 120 min. The daughter aromatic amines were rapidly mineralized within 6 h in preacclimatized bioreactor. The proposed hybrid treatment system can be a favourable alternative to energy demanding physical and chemical treatment options for toxic wastewaters.
Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.arabjc. 2017.03.001. References Alves de Lima, R.O., Bazo, A.P., Salvadori, D.M.F., Rech, C.M., Oliveira, D.P., Umbuzeiro, G.A., 2007. Mutagenic and carcinogenic potential of a textile azo dye processing plant effluent that impacts a drinking water source. Mutat. Res. 626, 53–60. Azetsu, A., Koga, H., Isogai, A., Kitaoka, T., 2011. Synthesis and catalytic features of hybrid metal nanoparticles supported on cellulose nanofibers. Catalysts 1, 83–96. Bafana, A., Krishnamurthi, K., Devi, S.S., Chakrabarti, T., 2008. Biological decolourization of C.I. Direct Black 38 by E. gallinarum. J. Hazard. Mater. 157, 187–193. Balasubramaniam, R., Kumar, A.V.R., 2000. Characterization of Delhi iron pillar rust by X-ray diffraction, Fourier transform infrared spectroscopy and Mossbauer spectroscopy. Corros. Sci. 42, 2085–2101. Bandara, J., Mielczarski, J.A., Kiwi, J., 1999. Molecular mechanism of surface recognition. Azo Dyes degradation on Fe, Ti, and Al oxides through metal sulfonate complexes. Langmuir 15, 7670–7679.
11 Bandara, J., Tennakone, K., Kiwi, J., 2001. Surface mechanism of molecular recognition between aminophenols and iron oxide surfaces. Langmuir 17, 3964–3969. Barsing, P., Tiwari, A., Joshi, T., Garg, S., 2011. Application of a novel bacterial consortium for mineralization of sulphonated aromatic amines. Bioresour. Technol. 102, 765–771. Cai, P.J., Xiao, X., He, Y.R., Li, W.W., Chu, J., Wu, C., He, M.X., Zhang, Z., Sheng, G.P., Lam, M.H.W., Xu, F., Yu, H.Q., 2012. Anaerobic biodecolorization mechanism of methyl orange by Shewanella oneidensis MR-1. Appl. Microbiol. Biotechnol. 93, 1769–1776. Cornell, R.M., Schwertmann, U., 2006. The Iron Oxides: Structure, Properties, Reactions, Occurrences and Uses. John Wiley & Sons Inc.. Crini, G., 2006. Non-conventional low-cost adsorbents for dye removal: a review. Bioresour. Technol. 97, 1061–1085. Elovitz, M.S., Weber, E.J., 1999. Sediment-mediated reduction of 2,4,6-trinitrotoluene and fate of the resulting aromatic (poly) amines. Environ. Sci. Technol. 33, 2617–2625. Fernando, E., Keshavarz, T., Kyazze, G., 2013. Simultaneous cometabolic decolourisation of azo dye mixtures and bio-electricity generation under thermophilic (50 °C) and saline conditions by an adapted anaerobic mixed culture in microbial fuel cells. Bioresour. Technol. 127, 1–8. Fernando, E., Keshavarz, T., Kyazze, G., 2014. Complete degradation of the azo dye Acid Orange-7 and bioelectricity generation in an integrated microbial fuel cell, aerobic two-stage bioreactor system in continuous flow mode at ambient temperature. Bioresour. Technol. 156, 155–162. Geetha, K.S., Surender, G.D., 1994. Solid–liquid mass transfer in the presence of micro-particles during dissolution of iron in a mechanically agitated contactor. Hydrometallurgy 36, 231–246. Hou, M., Li, F., Liu, X., Wang, X., Wan, H., 2007. The effect of substituent groups on the reductive degradation of azo dyes by zerovalent iron. J. Hazard. Mater. 145, 305–314. Huang, Y.H., Zhang, T.C., 2006. Reduction of nitrobenzene and formation of corrosion coatings in zerovalent iron systems. Water Res. 40, 3075–3082. Johnson, T.L., Tratnyek, P.G., 1995. Dechlorination of carbon tetrachloride by iron metal: The role of competing corrosion reactions. Preprint extended abstract, Am. Chem. Soc., Anaheim, CA, 35, 699-701. Jurado-Sanchez, B., Ballesteros, E., Gallego, M., 2012. Occurrence of aromatic amines and N-nitrosamines in the different steps of a drinking water treatment plant. Water Res. 46, 4543–4555. Kumar, K., Devi, S.S., Krishnamurthi, K., Gampawar, S., Mishra, N., Pandya, G.H., Chakrabarti, T., 2006. Decolorisation, biodegradation and detoxification of benzidine based azo dye. Biores. Technol. 97, 407–413. Kumar, R., Sinha, A., 2013. Degradation of mono-azo dye in aqueous solution using cast iron filings. IJRET 2, 227–231. Kumar, R., Sinha, A., 2016. Reductive transformation and enhancement in biodegradability of mono-azo dye by high carbon iron filings (HCIF). Desalin. Water Treat. 57, 3205–3217. Lama, Y., Sinha, A., Singh, G., Masto, R.E., 2016. Reductive dehalogenation of endosulfan by cast iron: kinetics, pathways and modeling. Chemosphere 150, 772–780. Mielczarski, J.A., Atenas, G.M., Mielczarski, E., 2005. Role of iron surface oxidation layers in decomposition of azo-dye water pollutants in weak acidic solutions. Appl. Catal. B 56, 289–303. Min, S., Risheng, Y., Yahua, Y., Shengsong, D., Wenxia, G., 2007. Degradation of 4-aminophenol by hydrogen peroxide oxidation using enzyme from Serratia marcescensas catalyst. Front. Environ. Sci. Engin. China 1, 95–98. Noubactep, C., 2009. Metallic iron for environmental remediation: learning from the Becher process. J. Hazard. Mater. 168, 1609– 1612.
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12 Quan, X., Zhang, X., Xu, H., 2015. In-situ formation and immobilization of biogenic nanopalladium into anaerobic granular sludge enhances azo dyes degradation. Water Res. 78, 74–83. Robinson, T., McMullan, G., Marchant, R., Nigam, P., 2001. Remediation of dyes in textile effluent: a critical review on current treatment technologies with a proposed alternative. Bioresour. Technol. 77, 247–255. Shanker, U., Singh, G., Kamaluddin, K., 2013. Interaction of aromatic amines with iron oxides: implications for prebiotic chemistry. Orig. Life Evol. Biosph. 43, 207–220. Sharan, R., 2011. Development of a cost effective integrated package for the treatment of coke plant effluents, Ph.D. Thesis, submitted to Indian School of Mines, Dhanbad. Singh, K., Arora, S., 2011. Removal of synthetic textile dyes from wastewaters: a critical review on present treatment technologies. Crit. Rev. Env. Sci. Tec. 41, 807–878. Sinha, A., Bose, P., 2009. Interaction of 2,4,6-trichlorophenol with high carbon iron filings: reaction and sorption mechanisms. J. Hazard. Mater. 164, 301–309.
R. Kumar et al. Sinha, A., Bose, P., 2011. 2-Chloronaphthalene dehalogenation by High Carbon Iron Filings (HCIF): impact of formation of corrosion products on HCIF surface. Environ. Engg. Sci. 28, 701–710. Sinha, A., Bose, P., 2014. Modeling of 2-chloronaphthalene interaction with high carbon iron filings (HCIF) in semi-batch and continuous systems. Environ. Sci. Pollut. Res. 21, 10442–10452. Skoog, D.A., Leary, J.J., 1992. Principles of Instrumental Analysis. Saunders College Publishing HBJ, U.S., pp. 150–155. Takenaka, S., Okugawa, S., Kadowaki, M., Murakami, S., Aoki, K., 2003. The metabolic pathway of 4-aminophenol in Burkholderia sp. Strain AK-5 differs from that of ANILINE and aniline with C-4 substituents. App. & Environ. Microb. 69, 5410–5413. Vasudevan, D., Stone, A.T., 1996. Adsorption of catechols, 2aminophenols, and 1,2-phenylenediamines at the metal (hydr) oxide/water interface: effect of ring substituents on the adsorption onto TiO2. Environ. Sci. Technol. 30, 1604–1613.
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King Saud University
Arabian Journal of Chemistry www.ksu.edu.sa www.sciencedirect.com
ORIGINAL ARTICLE
Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent Raja Kumar a, Alok Sinha a,*, Gautam C. Mondal b, Reginald E. Masto c a
Department of Environmental Science & Engineering, Indian Institute of Technology (ISM), Dhanbad 826004, Jharkhand, India Central Institute of Mining and Fuel Research (CSIR-CIMFR), Barwa Road, Dhanbad 826001, Jharkhand, India c Environmental Management Division, Central Institute of Mining and Fuel Research (Digwadih Campus), PO: FRI, Dhanbad 828108, India b
Received 11 November 2016; accepted 4 March 2017
KEYWORDS Adsorption; Biodegradability; Direct green 1 (DG1); Reduction; Scrap cast iron particles (SIP)
Abstract This study proposed that hybrid scrap cast iron particles (SIP)-aerobic biodegradation technology could enhance the biodegradability of toxic wastewater. SIP cleaved the azo linkages of Direct Green1 dye to form benzidine, 4-aminophenol, aniline and 1,2,7-triamino-8-hydroxynap thalene-3,6-disulfonic acid. SIP-mediated dye reduction was effective at wide pH range; however, kinetic analysis revealed fastest pseudo-first order dye reduction rate at acidic pH 3 (kd = 0.549 min1) followed by pH 9 (kd = 0.383 min1) and pH 7 (kd = 0.318 min1). The daughter aromatic amines produced were partially adsorbed onto the SIP surface and maximally at neutral pH. The adsorption process followed pseudo-second order adsorption kinetics and Langmuir isotherm. Benzidine was adsorbed more than 4-aminophenol and aniline. BOD5 of the SIP-treated effluent increased from 0.93 to 12 mg/L showing improved biodegradability. The daughter amines were rapidly mineralized in the aerobic bioreactor within 6 h. Cost-effective SIP pre-treatment could accelerate mineralization and detoxification of recalcitrant wastewater. Ó 2017 The Authors. Production and hosting by Elsevier B.V. on behalf of King Saud University. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
1. Introduction * Corresponding author. E-mail addresses: [email protected] (R. Kumar), [email protected] (A. Sinha), [email protected] (G.C. Mondal), [email protected] (R.E. Masto). Peer review under responsibility of King Saud University.
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Synthetic dyes have been extensively used in textile, power loom and dying industries. Accumulation of these dyestuff and dye wastewater creates not only environmental pollution, but also medical and aesthetic problems. More than 500 azo dyes that are based on carcinogenic amines are also recognized as hazardous contaminants and must be removed from waste effluents prior to discharge (Bafana et al., 2008). Besides, dyes wastes can also generate dangerous aromatic amine by-products through chemical transformation taking place in the aquatic-phase which are often more harmful to the environment than the parent dye molecule (Alves de Lima et al., 2007). These
http://dx.doi.org/10.1016/j.arabjc.2017.03.001 1878-5352 Ó 2017 The Authors. Production and hosting by Elsevier B.V. on behalf of King Saud University. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/). Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
2 amines have potential to persist and their physical properties make them highly mobile contaminants in reducing aquifers and sediments (Elovitz and Weber, 1999). Several aromatic amines, such as aniline (AN), 4-aminophenol (4-AP) and benzidine (BZN), are highly toxic chemicals which can lead to carcinogenesis, teratogenesis, and mutagenesis (Jurado-Sanchez et al., 2012). Many of them have been listed as priority pollutants by the US EPA, EU, and China EPA. Hence, it is necessary to contain azo dyes and their daughter amines, either by physico-chemical or by microbial routes. Available literature reports scores of physical, chemical and biological technologies evolved so far for azo dye effluent treatment. Unfortunately, the practical application of most of the reported technologies is limited by low efficiency, longer retention time, difficulty of operation, unfavourable economics, production of large volume of sludge, generation of secondary waste products and environmental incompatibility (Robinson et al., 2001; Crini, 2006). Biological treatment is considered as cheapest alternative for remediation of most of the wastewaters. However, aerobic biodegradation systems are not successful for decolorizing the azo dye-containing effluents because of their strong electron-withdrawing azo linkages (AN‚NA), which protects them from attack by the oxygenases (Singh and Arora, 2011). Anaerobic biological treatment of azo dye has been investigated by several researchers (Fernando et al., 2013; Quan et al., 2015) but this process can only decolorize azo dye (by splitting the azo bond) while leaving behind potentially toxic aromatic amines generated from the reaction. The fact that these aromatic amines can be degraded by aerobic microflora (Fernando et al., 2014) has drawn attention towards applicability of combined anaerobic–aerobic biological methods for the treatment of azo dye-containing effluent. However, a major limitation to any biological technique is longer treatment time. To overcome this problem a rapid dye degradation technology is sought after which should be cheap and environmentally sustainable. Zero valent iron (ZVI) seems to be the most appropriate replacement for the anaerobic biological pre-treatment. ZVI is an electron donor, a strong reducing agent, and the spent agent, Fe2+ is environmentally innocuous. Azo dye decolourization studies performed using ZVI have shown very rapid dye reduction at varied experimental conditions (Mielczarski et al., 2005; Kumar and Sinha, 2013). Combination of ZVI based dye reduction technology with aerobic biological treatment can achieve complete mineralization of azo dye solutions, that too in much quicker time. Therefore, in the present study, a hybrid method combining the reducing effect of ZVI to discolour the azo dye effluent, making it amenable for subsequent aerobic biodegradation step, was investigated. The form of ZVI used was scrap cast iron particles (SIPs) which are available in plenty at much lower cost compared to pure electrolytic iron. This study explores the efficiency of SIPs in decolourization of azo dye. Available literature is deficient of information on the interaction of azo dyes and their reductive transformation products with metallic iron. Both qualitative and quantitative analyses of formation and adsorption of daughter aromatic amines have been done. The iron-treated synthetic solution has been screened for amenability to biological treatment. To this end, this study may provide new insight into the interaction of metallic iron with aqueous contaminants and may help in the development of an integrated detoxification process for complete remediation of textile dye-bath effluent.
R. Kumar et al. without any purification. Analytical grade benzidine (BZN), aniline (ANI) and 4-aminophenol (4-AP) were procured from Sigma–Aldrich (Germany), Merck (USA), and Acros Organics (Belgium), respectively. HPLC grade methanol used as an HPLC eluent was purchased from Merck. Deionised water was used throughout the whole experimental process. 2.2. Preparation and characterization of SIPs The SIPs used in this research were prepared using cast iron chips obtained from a local foundry. Cast iron chips are generated while cutting the edges of casted billets in order to make its surface smooth. The collected cast iron chips were ground into smaller iron particles in a dough-size ball mill. The particles were sieved and those passing through sieve #40 and remaining on sieve #80 were retained for use. These particles were washed 5 times with N2-sparged 1 N HCl with periodic shaking. The SIPs, thus obtained, were rinsed 8–10 times with N2-sparged deionised water; instantly rinsed twice with methanol and dried for an hour under anaerobic condition. The SIPs were then stored in a dessicator filled with nitrogen gas. X-ray diffraction (XRD) analyses (data not shown) of the SIP surface carried out using X-ray Powder Diffractometer (Bruker Advanced X-ray Solutions, GMBH, Germany) indicated that the SIP surface was devoid of any oxide coating. The carbon content (per cent weight basis) of the SIPs as determined by LECO Induction Furnace Instrument (LECO Corporation, St. Joseph, USA) was found to be 3.08%. Specific surface area of the SIPs was determined by BET (Brunauer Emmett Teller) surface area analyzer (NOVA 4000e, Quantachrome Instruments, USA) to be 2.16 m2/g. 2.3. Batch reduction experiments The reduction in Direct Green 1 (DG1) azo dye by SIPs in aqueous solution was investigated thoroughly in batch and column experiments. All the experiments were carried out in 35 mL capacity borosilicate glass vials bearing screw caps (Borosil, India). The concentration of DG1 dye and SIPs was pre-optimized and was kept uniform at 100 mg/L and 57.12 g/L, respectively, for all the experiments. The desired initial pH value in aqueous solution was adjusted with 0.1 M HCl or 0.1 M NaOH. The vials were properly shaken on an end-toend test tube rotator (Rotospin, Tarsons, India) at 30 rpm in room temperature (25 ± 2 C). With every test, a SIP-free dye solution was run as a control. The vials were periodically withdrawn from the rotator in duplicate along with a control for sampling and analysis, and were put on a magnet to allow iron particles to settle down quickly. The supernatants from each vial were filtered through GF/C filter paper (1.2 mm nominal pore size, Whatman, Springfield Mill, England) and the filtrate was quantified for colour and reduction by-products.
2. Materials and methods 2.4. Desorption of reduction by-products adsorbed onto the SIPs 2.1. Chemicals and reagents Commercially available tris-azo dye C.I. Direct Green 1 (2,7Naphthalenedisulfonic acid, 4-amino-5-hydroxy-3-((40 -((4-hyd roxyphenyl) azo) (1,10 -biphenyl)-4-yl)azo)-6-(phenylazo)-, disodium salt; M.F.: C34H23N7Na2O8S2; M.W.: 767.70) was obtained from Mac Dye Chem, Ahmedabad (India), and used
To extract the DG1 reduction by-products adsorbed to the SIP surface, the treated dye solution of the 35 mL reaction vial was transferred to another vial. To the vial containing the used SIPs, 5 mL methanol was added and the contents were vortex mixed for 10 min. The solvent along with desorbed compounds was transferred to another vial. Again 5 mL of methanol was
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
Complete mineralization of benzidine based azo dye effluent added to the solid phase and the contents were mixed for another 10 min to ensure desorption of any residual compounds. This solvent was also transferred to the same vial. The 10 mL of methanol, thus collected, was filtered through 0.22 mm filter disc. The filtrate was used for qualitative and quantitative analysis of adsorbed daughter products. 2.5. Hybrid scrap iron-aerobic biodegradation treatment system To evaluate the feasibility of treatment of DG1 dye-containing wastewater in a continuous flow through system and to assess the biodegradability of daughter products after SIP pretreatment, a hybrid textile effluent treatment system (as shown in Fig. S1, Supplementary Information) was constructed. The pre-treatment column was fabricated using 4 mm thick transparent acrylic plastic tubing (60 cm L 2.5 cm i.d.) provided with an inlet and an outlet. The column was loaded with SIPs and well rinsed sterilized sand (20–30 mesh size) mixture in 1:1 ratio. Column end caps, valves, and tubing were made of polypropylene and a variable speed peristaltic pump (Rivotek, India) was used to pump the solution through the columns at a flow rate of 10 mL/min in an up-flow mode to prevent gas entrainment. Aerobic biodegradation studies were carried out in 4 L batch reactors, with a working volume of 2 L, where constant air flow was maintained using aquarium air pumps and porous stone diffusers. The bioreactor was seeded with activated sludge collected from Durgapur Coke oven effluent Treatment Plant, West Bengal, India. The reactors were initially fed with synthetic wastewater containing COD concentration of 1000 mg/L along with 1 mL/L trace metal solution (Sharan, 2011) (The detailed characteristics of the feed and composition of trace metal solution are given in Tables A1 and A2, respectively, in Supplementary Information). Both the reactors were operated for 3–4 weeks under room temperature (25 ± 2 °C) until the steady state conditions were established. Before exposing the microbial flora to the SIP-treated effluent, the bioreactors were acclimatized to the aromatic amines detected after reductive degradation of DG1. During operation, the treated effluent was collected from the exit port of the SIP/ sand column and was analysed by HPLC for the quantification of degradation by-products formed and then was transferred to the aerobic bioreactor. After 24 h of HRT in the bioreactor, the hybrid system effluent was quantified. 2.6. Analytical procedure Preliminary experiments were carried out at solution pH 3 using 57.12 g/L SIP dose to understand DG1 degradation mechanism and to identify the reduction by-products. The surface morphology and composition of SIP surface during the course of DG1 degradation reaction were studied using FESEM (Field Emission Scanning Electron Microscopy) and Energy dispersive spectra (EDS) and FTIR, respectively. UV–Vis spectral analysis and LC-MS/MS were used for qualitative analysis of dye degradation and reduction by-products formed. For kinetic studies, colour removal and concentration of by-products were analysed using UV–Vis spectrophotometer and HPLC, respectively. The details of analytical procedure and conditions using each instrument are given in the Supplementary Information.
3 3. Results and discussion 3.1. Interaction of azo dye with SIP surface: characterization of fresh and used SIP surface Fig. 1(a) –(c) shows the SEM images of fresh SIP surface, SIP surface after 10 min of reaction and SIP surface after 120 min of reaction, respectively. Fresh SIP surface was free from rust; hence, no peak for oxygen was observed in its EDS spectra (Fig. S2-a, Supplementary information). Also, it showed highest iron content of 93.73% (see Table A3, Supplementary information) and was free from any surface deposition. SEM image of SIP taken after 10 min of reaction (Fig. 1b) shows large amount of deposition on its surface which could be adsorbed dye molecules. This statement is supported by the corresponding EDS spectra (Fig. S2-b, Supplementary information) which shows increase in carbon (17.63%) and sulphur content (1.41%) indicating towards the adsorption of sulphonated organic compound, that could be DG1 dye molecule and/or its reduction by-product. After 120 min of reaction, when decolourization was almost complete, the SIP surface exhibited rusted morphology with some amount of surface deposition (Fig. 1c). The carbon (9.74%) and sulphur (1%) content decreased marginally (Fig. S2-c, Supplementary information), but not completely, suggesting adsorption of degradation by-products onto the rusted surface. 3.1.1. FTIR spectra analysis of SIP surface Fig. 2 illustrates the FTIR spectra of fresh SIP surface and used SIP surface (sampled after 10 min and 120 min of reaction time). FTIR spectroscopy of SIPs performed after 10 min of reaction shows an intense band at 1514 cm1, attributable to the azo-chromophore (AN‚NA) bond vibrations of adsorbed unreduced dye molecule. This band was absent in CIP sampled after 120 min of reaction suggesting breaking of azo bonds. The broad peak at 3000–3500 cm1 is due to hydration of rust and is formed due to OAH stretching (Balasubramaniam and Kumar, 2000). The medium-weak, multiple bands between 1400 and 1600 cm1 were assigned to C‚C aromatic skeleton vibrations, while the band at 1403 cm1 was ascribed to the OAH bending vibrations and deformation (Bandara et al., 1999). Strong bands at 1104, 1156 and 1207 cm1 were assigned to the symmetric and asymmetric stretching of sulphonate anions (ASO 3 ) of naphthalene ring (Hou et al., 2007). The weak frequencies at 1360 cm1 were assigned to the CAN stretching due to the resonance effect of the aromatic ring and those between 1617 and 1680 cm1 were attributable to the in-plane bending vibrations for NAH of aromatic amines. These bands were found in all the used SIP samples. This suggests fast reduction in DG1 and adsorption of its reduction by-products bearing ASO3H and ANH2 functional groups onto the SIP surface. 3.2. Identification of reduced and adsorbed daughter products 3.2.1. UV–Vis spectral analysis The molecular structure of DG1 has three azo groups. The chromophore containing azo linkage in the DG1 dye molecule has absorption in the visible region (k = 603.5 nm), while the characteristic wavelength of benzene-ring (k = 234.5 nm) and
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Figure 1 Scanning electron micrographs of (a) fresh SIP surface, (b) SIP surface after 10 min of reaction, and (c) SIP surface after 120 min of reaction.
Figure 2 FTIR spectra of SIP surface employed for Direct Green 1 degradation.
naphthalene ring (k = 376 nm) in the UV region (Fig. 3a). During the decolourization process, the absorbance band at 603.5 nm disappeared completely within 20 min of the reduction reaction using SIPs. Also, the absorbance bands at 234.5 nm and 376 nm disappeared; instead, a new peak appeared at 280 nm which became more prominent with reaction progress. This peak is attributable to the aromatic ring structures of BZN and 4-AP (Min et al., 2007; Azetsu et al., 2011) and has been matched with their reference spectrum
(Fig. S3, Supplementary information). ANI was also expected to form due to breaking of one of the azo bonds but no separate peak was observed at 232 nm wavelength. This could be either due to adsorption of ANI onto SIPs and/or due to expectedly low stoichiometric formation of ANI. However, the secondary absorbance peak of ANI is also reported to form at 280 nm, which is observed in the spectra. On the basis of the sole UV–VIS spectra, no other species were detected. Fig. 3b shows the UV–VIS absorbance spectra of DG1 dye desorbed from SIPs using methanol at definite time intervals of 10 min, 30 min and 120 min. In this case, two peaks, i.e. at 603.5 nm and 376 nm, of the major peaks found in the parent dye molecule were found in the solution desorbed after first 10 min. The initial peak found earlier at 234.5 nm shifted towards shorter wavelength (at 227 nm) when the p bonds of characteristic conjugated chromophores in an azo dye molecule were broken (Skoog and Leary, 1992). The new peak at 280 nm was also observed to form in solution desorbed after 10 min of reductive treatment. All the peaks observed in parent compound vanished after 120 min whereas the new peak at 280 nm, which was attributed to aromatic amine formation, increased in height indicating towards increase in their adsorbed concentration. The peak for ANI was formed at the wavelength of 232 nm. The intensity of the peak was low due to the low concentration of adsorbed ANI. 3.2.2. Liquid chromatography and mass spectrometric analysis The HPLC chromatograms of SIP treated aqueous phase and methanol extracted solid phase showed three prominent peaks for daughter products at 3.2, 3.6, and 4.2 min retention times (Fig.S4, Supplementary information). Fig. 4(a) and (b) shows
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
Complete mineralization of benzidine based azo dye effluent
5
(b) Figure 3 UV–VIS spectra of (a) treated aqueous phase DG1 solution as a function of reaction time (b) methanol extracted solid phase at definite reaction time (Initial conditions: pH 3; dye conc. 100 mg/L; SIP dose 57.12 g/L; mixing speed 30 rpm).
the mass spectra of the SIP-treated DG1 aqueous solution recorded after 10 min and 120 min of reaction intervals, respectively. The spectra exhibit three major species with m/z ratio of 185, 109 and 94, of which former was the most abundant. These positive ions were attributable to BZN, 4-AP and ANI, respectively, which were formed by the reductive cleavage of the azo linkages of DG1 molecule. Their identification was further validated by the MS/MS fragmentation pattern of BZN, 4-AP and ANI (Fig.S5, Supplementary information) that matched their respective NIST library record. It is confirmed that only these three daughter products were formed after DG1 reduction. The other expected daughter product,
namely 1,2,7-triamino-8-hydroxynapthalene-3,6-disulfonic acid (TAHNDS), could not be identified. The mass spectrum of dye solution desorbed from SIP surface using methanol after 120 min also exhibits same three by-products (Fig. 4c) of which species with m/z 185 was the most abundant, indicating its higher degree of adsorption. LC-MS analysis could confirm only these three daughter amines formed after DG1 reduction. However, SEM-EDS and FTIR analysis reveals adsorption of a species bearing –SO3H functional group onto the SIP surface. This could be the other expected daughter product, namely TAHNDS. The proposed mechanism involved in the transformation of DG1 by SIP is represented in Fig. 4(d).
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Figure 4 ESI mass spectrum of solution during reductive degradation of DG1, (a) after 10 min, (b) after 120 min of reaction, (c) desorbed from SIP surface after 120 min of reaction (initial DG1 concentration; 100 mg/L; SIP dosage: 57.12 g/L; pH 3.0), (d) proposed pathway of Direct Green 1 reductive transformation (SIP:scrap iron particle; 4-AP: 4-aminophenol; ANI: aniline, BZN: benzidine, TAHNDS:1,2,7-triamino-8-hydroxynapthalene-3,6-disulfonic acid).
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
Complete mineralization of benzidine based azo dye effluent 3.3. Kinetics of DG1 reduction and formation of aromatic amines In water or wastewater treatment, pH of the aqueous solution is a complex parameter since it is related to the ionization state of the metal surface and that of the reactants and products. Hence, kinetics of degradation of DG1 and formation of its daughter by-products were investigated at three levels of pH 3, 7, and 9 (representing acidic, neutral, and alkaline conditions, respectively). Dye degradation by metallic iron is assumed to be pseudo-first order reaction; hence, the rate of degradation of DG1 and the rate of formation of daughter aromatic amines were evaluated by nonlinear regression of the experimental data using Eqs. (1) and (2), respectively. Ct ¼ C0 ekd t
ð1Þ
Ct ¼ C0 ð1 ekf t Þ
ð2Þ
where Ct (mg/L) is the concentration of analyte at time t (min), C0 (mg/L) is the initial or final concentration, kd (min1) is the pseudo-first order DG1 degradation rate coefficient and kf (min1) is the rate coefficient of formation of daughter aromatic amines. The temporal variation in the concentration of DG1 and simultaneous increase in the total concentration of its reduction by-products at pH 3,7 and 9 are exhibited in Fig. 5(a) (b) and (c), respectively, and the result of kinetic analysis is shown in Table 1. The final DG1 degradation efficiencies recorded after 120 min at pH 3, 7 and 9 were 99.82%, 99.35% and 99.50%, respectively. The respective final concentration of BZN, 4-APl and ANI was 13.82 mg/L and 8.57 mg/ L and 7.2 mg/L at pH 3; 11.87 mg/L, 7.25 mg/L and 7.12 mg/L at pH 7; 13.40 mg/L and 8.25 mg/L and 7.17 mg/L at pH 9. The rate of degradation of DG1 followed the order (from the highest to lowest) pH 3 > pH 9 > pH 7. At acidic pH 3, H+ would obviously promote the corrosion of SIPs providing more amount of Fe2+ to the reaction. This could account for a higher rate of DG1 degradation and appearance of daughter amines. At alkaline pH 9, OH would significantly enhance the formation of the iron hydroxides and oxy-hydroxides. It is reported that the iron oxide(s) can act as semi-conductor to relay electron transfer from iron to the contaminants (Huang and Zhang, 2006). Additionally, the oxides in Fe0/ H2O systems are good adsorbents for FeII originating from iron oxidation yielding adsorbed FeII (structural FeII). Structural FeII is a very strong reducing agent for inorganics and organics (Geetha and Surender, 1994). This could be the reason for the dye getting reduced at pH 9 and stoichiometrically higher concentration of daughter amines formed compared to that at pH 7. Reduction in DG1 dye at pH 9 and formation of high amount of by-products contradict the literature reports that claim removal of dye molecules from aqueous phase by adsorption and/or co-precipitation at alkaline pH, regardless of redox activity (Noubactep, 2009). Low dye reduction at pH 7 could be because of slower iron corrosion (Johnson and Tratnyek, 1995); hence, neither much of the rejuvenated iron surface and H+ were present (as in case of acidic conditions) nor was the oxide and hydroxide coated adsorptive surface area (as in case of alkaline conditions) (Kumar and Sinha, 2016). The rate of appearance of ANI and 4-AP was faster than that of BZN. This could be because these two amines were
7 formed due to splitting of single azo bonds present at the either end of DG1 molecule, which were supposed to lie in closer proximity to the iron surface thus facilitating faster sharing of electrons. On the other hand, the formation of BZN required breaking of two azo bonds one of which lied deep in the DG1 structure such that it’s splitting could get delayed. Mass balance (%) evaluated for these three by-products (considering their stoichiometric, aqueous and adsorbed concentration) at different solution pH was between 55% and 60% (see Table 1). The purity of DG1 used in this study was approximately 60%; therefore, the obtained mass balance was sustainable 3.4. Simultaneous adsorption of daughter aromatic amines: Isotherm and kinetics Considerable amount of daughter aromatic amines was found to adsorb onto the SIP surface. Hence, the interactive behaviour of aromatic amines with SIP was studied at different pH. The feasibility of the adsorption process is predicted using adsorption isotherm, which also predicts the distribution of adsorbate molecules between the liquid and solid phase at equilibrium. The experimental isotherm data were fitted using two most common adsorption models, i.e. the Langmuir model (corresponding to monolayer homogeneous adsorbent surface) and the Freundlich model (corresponding to heterogeneous adsorbent surface). The details of the adsorption isotherms are given in the Supplementary information. Fig. S6 (in Supplementary information) shows the applicability of both models at different pH and the Freundlich and Langmuir parameters as calculated using the nonlinear regression analysis are reported in Table 2. Based on the higher correlation coefficient (r2) and lower root mean square error (RMSE) value (Table 2), it is evident that Langmuir isotherm model shows better mathematical fit with experimental data. This suggests that the adsorption is monolayer at different pH. The value of Langmuir constant ‘KL’ followed the descending order: pH 7 > pH 9 > pH 3. This specifies the stronger affinity between the adsorption site and aromatic amines at neutral pH. In order to analyse the adsorption kinetics of aromatic amines, correlations between their adsorbed amounts and time were looked for, through the nonlinear regression of different mathematical expressions corresponding to pseudo-first-order equation and pseudo-second-order reaction. The details of the kinetic models are given in the Supporting Information. The kinetics data of aromatic amine adsorption studied at different pH (Fig. S7, Supporting Information) fitted best into pseudo-second-order model (based on better r2 and lower RMSE, see Table 3), which is based on the assumption that the rate limiting step may be chemisorption involving valency forces through sharing or exchange of electrons between the functional groups of the amines and the SIP surface. 3.5. Discussion on the adsorption of aromatic amines Comparative analysis of the data for sorption of aromatic amines onto the SIP surface reveals that sorption of BZN per gram SIP was the highest followed by 4-AP and ANI. This observation can be rationalized on the basis of aqueous solvation properties, structural and electrochemical properties of the
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R. Kumar et al.
Figure 5 Temporal variation in the concentration of DG1, daughter aromatic amines and stoichiometric mass balance for the daughter aromatic amines at (a) pH 3, (b) pH 7 and (c) pH 9.
three amines. The aqueous solubility of the three amines at 25 °C follows the descending order: BZN (520 mg/L) > 4-AP (15 g/L) > ANI (36 gm/L). Lesser the solubility of a compound, greater is the exclusion from bulk solution and higher is the extent of adsorption. However, the reported values of
octanol-water partition coefficients (log Kow) for the three amines follow a different trend, i.e. BZN (1.34) > ANI (0.90) > 4-AP (0.04). The value of log Kow increases as the compounds become more hydrophobic; therefore, highest BZN adsorption is primarily due to hydrophobicity. Although
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
Complete mineralization of benzidine based azo dye effluent Table 1
9
Kinetic parameters for the degradation of DG1 and formation of daughter aromatic amines.
pH
Dye/Amines
kd (1/min)
r2
3
DG1 BZN 4-AP ANI
0.549
0.988
DG1 BZN 4-AP ANI
0.318
DG1 BZN 4-AP ANI
0.383
7
9
kf (1/min)
r2
Mass balance (%)
0.284 0.347 0.342
0.990 0.984 0.966
58.10 59.70 59.88
0.187 0.230 0.228
0.989 0.944 0.935
57.98 57.68 59.70
0.211 0.286 0.265
0.967 0.948 0.941
56.50 59.18 59.22
0.985
0.993
Table 2
Parameters of Langmuir and Freundlich isotherms for adsorption of daughter aromatic amines.
pH
Amines
Langmuir isotherm
Freundlich isotherm 2
qm (mg/g)
KL (L/mg)
r
RMSE
KF (mg/g)
n
r2
RMSE
3
BZN 4-AP ANI
0.1080 0.0606 0.0535
0.111 0.243 0.298
0.993 0.996 0.990
0.775 0.851 2.186
0.015 0.012 0.010
0.580 0.557 0.560
0.991 0.989 0.981
1.225 1.406 2.235
7
BZN 4-AP ANI
0.0866 0.0550 0.0437
0.336 0.400 0.435
0.987 0.992 0.995
0.890 0.518 0.640
0.026 0.019 0.016
0.401 0.404 0.421
0.975 0.991 0.984
1.519 0.696 0.881
9
BZN 4-AP ANI
0.0831 0.0520 0.0415
0.204 0.334 0.236
0.992 0.988 0.975
1.020 0.823 1.680
0.024 0.017 0.011
0.419 0.471 0.489
0.982 0.985 0.969
1.257 0.956 1.994
Table 3
Kinetic parameters for the adsorption of daughter aromatic amines.
pH
Amines
Pseudo-first-order
Pseudo-second-order
K1 (1/min)
qe (mg/g)
r2
RMSE
K2 (g/mg /min)
qe (mg/g)
r2
RMSE
3
BZN 4-AP ANI
0.443 0.412 0.388
0.062 0.035 0.027
0.953 0.969 0.977
1.43 1.25 1.91
10.712 17.430 18.819
0.065 0.036 0.028
0.987 0.980 0.994
1.34 0.61 0.85
7
BZN 4-AP ANI
0.522 0.485 0.422
0.064 0.039 0.030
0.957 0.979 0.960
1.82 1.59 2.10
13.272 28.254 35.354
0.067 0.041 0.032
0.976 0.993 0.984
0.68 0.79 1.23
9
BZN 4-AP ANI
0.490 0.426 0.410
0.061 0.036 0.028
0.959 0.973 0.966
2.00 1.82 4.05
12.457 24.290 24.297
0.064 0.038 0.029
0.979 0.992 0.988
0.88 0.81 3.21
log Kow of ANI is greater than 4-AP, the observed adsorption of 4-AP is higher. This is because AOH group in 4-AP possesses a negative charge and a high charge-to-radius ratio that yields a large ionic contribution to complex formation. In contrast, the ANH2 group (present in both 4-AP and ANI) pos-
sesses a neutral formal charge, and the ionic contribution to bonding is limited to the interaction between ionic charge on the metal surface and the dipole moment of the amino group. Since ion-dipole interactions are known to be weak, the ANH2 group has a lower contribution to ionic bonding (Vasudevan
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10 and Stone, 1996). Also, compared to ANI that possess only one ANH2 group, the presence of both AOH and ANH2 group in 4-AP is likely to exert stronger cumulative effect in complex formation with SIP surface, hence the greater degree of adsorption. The sorption of aromatic amines observed in the experiment is seen to increase with time, possibly because of the simultaneous increase in oxidation of SIP surface, as revealed by SEM-EDS and FTIR analysis. Previous researches performed using cast iron have reported the partitioning of contaminants onto graphitic inclusions (Sinha and Bose, 2009, 2011, 2014; Lama et al., 2015). Hence, adsorption of aromatic amines onto the graphitic inclusions present in the SIP matrix also cannot be ruled out. However, such speculations required experimental validation which is not the scope of current research. Whatever be the adsorbent, it is likely to get influenced by solution pH that affects the surface charge of adsorbents as well as the degree of ionization of the adsorbate in the solution. In this study, highest adsorption was observed at pH 7 followed by pH 9 and pH 3. Based on the literature, iron (hydr)oxides have pKa1 and pKa2 values of 5 and 10 (Cornell and Schwertmann, 2006), respectively, resulting a pHzpc on the order of 8.1. At pH < pHzpc, the SIP surface has a net positive charge; at pH > pHzpc, the surface has a net negative charge. This could facilitate adsorption of neutral and positively charged aromatic amines at pH 7 and pH 9 systems. This adsorption behaviour is in agreement with those reported in the literature for polar aromatic amines where maximum adsorption occurs around pH 7 (Vasudevan and Stone, 1996; Bandara et al., 2001; Shanker et al., 2013). At pH 3, initially the adsorption was slow possibly because of increased dissolution of iron surface. However, adsorption increased with time as the solution pH approached towards neutral values (refer temporal pH variation profiles shown in Fig. S8, Supplementary information). Hence, the observed adsorption trend of aromatic amines here in this study follows the descending order: BZN > 4-AP > ANI. This order also exhibits the descending basic nature of these amines, conferred upon them by the number of electron-donating substituents attached to the aromatic structure, which reflects the lesser availability of electrons for the interaction.
R. Kumar et al. reactor, it was subjected to a biological treatment step. Fig. 6(a), (b) and (c) shows the respective concentration profile of BZN, 4-APl and ANI in the column reactor effluent and the final effluent from aerobic bioreactor. Complete degradation of DG1 was observed in the SIP column reactor. However, a large amount of degradation by-products were found to remain adsorb within the column reactor. The concentration of BZN in the column effluent was between 0.5 and 0.9 mg/ L as against its expected stoichiometric concentration of 14.40 mg/L. ANI concentration in the column effluent was between 2 and 3 mg/L as against its expected stoichiometric concentration of 7.26 mg/L. The concentration of 4-AP in the effluent was 3.0–4.5 mg/L while its expected stoichiometric concentration was 8.52 mg/L. This suggests that ANI and 4AP were partially adsorbed through the column while BZN was contained largely within the column by adsorption. Careful analysis of DOC data (Fig. S10, Supporting Information) and accounting for the dye purity, suggest substantial adsorption of TAHNDS within the column reactor. DOC of the
3.6. Biodegradability assessment The preliminary biodegradability studies performed for batch reactor effluent on the basis of BOD5 were encouraging and formed the basis for the subsequent hybrid treatment. The BOD5 of untreated DG1 solution was 0.93 mg/L which indicates that raw dye containing effluent inhibits the respiratory activity of bacterial seed. Therefore, besides being difficult to biodegrade, the dye containing effluent can also decrease the efficiency of biological treatment process. After 120 min of SIP treatment in batch system, the BOD5 increased to 12 mg/L (refer Fig. S9, Supplementary information). Evidently, the reductive splitting of azo linkages of DG1 by SIP forms simpler fragments that easily accept electrons and are more amenable to subsequent aerobic biological treatment process. In the hybrid treatment process, after the DG1 dyecontaining effluent was pre-treated with SIP in the column
Figure 6 Concentration of (a) benzidine, (b) 4-aminophenol, and (c) aniline in the effluent from SIP column and the aerobic bioreactor.
Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001
Complete mineralization of benzidine based azo dye effluent influent DG1 wastewater passed through SIP column was 30.8 mg/L while the column effluent had DOC of 5– 9 mg/L. Calculated DOC of TAHNDS (i.e. 9.4 mg/L for 60% pure dye) is higher than the effluent DOC, thus pointing towards its adsorption. In the bioreactor, all of the observed aromatic amines generated from SIP column were completely mineralized; however, DOC removal was only around 62%. Assuming complete mineralization of glucose and acetate (present in feed), the residual DOC could be attributed to TAHNDS, since the microbes were not pre-acclimatized to this compound. No trace of other amines was found in the samples analysed even after 6 h of aerobic treatment, indicating their rapid biodegradation. Morphological study of activated sludge biomass in the bioreactor reveals the predominance of rodshaped bacilli (Fig. S11, Supplementary Information). Many researchers have utilized rod-shaped bacterial species such as Pseudomonas sp. (Kumar et al., 2006; Barsing et al., 2011); Shewanella sp. (Cai et al., 2012) and Burkholderia sp. (Takenaka et al., 2003) for the treatment of 4-aminophenol, benzidine and aniline. The above observations clearly demonstrate that the pre-treatment with SIP reduced the toxicity of DG1 and was capable of stimulating rapid mineralization of the residual pollutant. 4. Conclusions Scrap cast iron particles (SIP) showed enormous potential towards rapid enhancing of biodegradability of textile wastewater. SIP effectively cleaved azo bonds of DG1 to form bio-available aromatic amines. Dye degradation was highest at acidic pH whereas the concurrent adsorption of degradation by-products was maximal at neutral pH. Biodegradability assessment revealed significant increase in the BOD5 of the SIP-treated dye solution within 120 min. The daughter aromatic amines were rapidly mineralized within 6 h in preacclimatized bioreactor. The proposed hybrid treatment system can be a favourable alternative to energy demanding physical and chemical treatment options for toxic wastewaters.
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Please cite this article in press as: Kumar, R. et al., Effective scrap iron particles (SIP) pre-treatment for complete mineralization of benzidine based azo dye effluent. Arabian Journal of Chemistry (2017), http://dx.doi.org/10.1016/j.arabjc.2017.03.001