- Email: [email protected]
heterogonous Fenton-type processes – A review
Chemosphere 174 (2017) 665e688
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Review
Removal of pharmaceuticals from water by homo/heterogonous Fenton-type processes e A review Amir Mirzaei, Zhi Chen*, Fariborz Haghighat, Laleh Yerushalmi Department of Building, Civil and Environmental Engineering, Concordia University, Montreal, H3G 1M8, Canada
h i g h l i g h t s Safe, environmentally-benign and relatively cheap reagents of Fenton-type reactions. Comprehensive review of controlling parameters on the Fenton-type reactions. Narrow pH range and iron-containing sludge as the homogenous Fenton limitations. The slow reaction kinetics is the major drawback of heterogeneous Fenton reaction. Recent strategies to address these limitations are presented.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 27 September 2016 Received in revised form 1 February 2017 Accepted 3 February 2017 Available online 4 February 2017
The presence of emerging contaminants such as pharmaceuticals in natural waters has raised increasing concern due to their frequent appearance and persistence in the aquatic ecosystem and the threat to health and safety of aquatic life, even at trace concentrations. Conventional water treatment processes are known to be generally inadequate for the elimination of these persistent contaminants. Therefore, the use of advanced oxidation processes (AOPs) which are able to efficiently oxidize organic pollutants has attracted a great amount of attention. The main limitation of AOPs lies in their high operating costs associated with the consumption of energy and chemicals. Fenton-based processes, which utilize nontoxic and common reagents and potentially can exploit solar energy, will considerably reduce the removal cost of recalcitrant contaminants. The disadvantages of homogeneous Fenton processes, such as the generation of high amounts of iron-containing sludge and limited operational range of pH, have prompted much attention to the use of heterogeneous Fenton processes. In this review, the impacts of some controlling parameters including the H2O2 and catalyst dosage, solution pH, initial contaminants concentrations, temperature, type of catalyst, intensity of irradiation, reaction time and feeding mode on the removal efficiencies of hetero/homogeneous Fenton processes are discussed. In addition, the combination of Fenton-type processes with biological systems as the pre/post treatment stages in pilot-scale operations is considered. The reported experimental results obtained by using Fenton and photo-Fenton processes for the elimination of pharmaceutical contaminants are also compiled and evaluated. © 2017 Elsevier Ltd. All rights reserved.
Handling Editor: Xiangru Zhang Keywords: Water treatment Emerging contaminants Pharmaceuticals Toxicity Photo-Fenton Heterogeneous Fenton process
Contents 1. 2.
3.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 666 Mechanism of Fenton-type reactions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 667 2.1. Homogeneous Fenton processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 667 2.2. Heterogeneous Fenton processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 669 2.2.1. Iron oxide minerals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 669 2.2.2. Support material in heterogeneous Fenton processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 669 Major controlling parameters in the Fenton reaction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 674
* Corresponding author. E-mail address: [email protected] (Z. Chen). http://dx.doi.org/10.1016/j.chemosphere.2017.02.019 0045-6535/© 2017 Elsevier Ltd. All rights reserved.
666
4. 5. 6. 7.
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
3.1. Effect of reagents concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 674 3.2. Effect of initial contaminants concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 678 3.3. Effect of catalyst type . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 678 3.4. Effect of pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 679 3.5. Effect of intensity of radiation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 680 3.6. Effect of water matrix . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 681 3.7. Effect of the salinity of substrate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 681 3.8. Effect of feeding mode . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 682 3.9. Effect of temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 682 3.10. Effect of reaction time . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 682 3.11. Deactivation of catalyst . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 683 3.12. By-products formation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 683 Toxicity and biodegradability tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 683 Large-scale application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 684 Summary of findings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 684 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 685 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 685
Nomenclature AAP ALC AMX AOP ATL BLB BZA BZF CAP CBZ CPC CWPO DCF DOC DW EC EDDP EDTA FeOx GAC HA HP IBP
acetaminophen amphetamine-like compound amoxicillin advanced oxidation process b-blocker atenolol black blue lamp benzotriazole bezafibrate chloramphenicol carbamazepine compound parabolic collector catalytic wet peroxide oxidation diclofenac dissolved organic carbon distilled water emerging contaminant 2-ethylene 1,5-dimethyl3,3-diphenylpyrrolidine ethylenediaminetetraacetic acid ferrioxalate granular activated carbon humic acid high pressure ibuprofen
1. Introduction Emerging contaminants (ECs) such as pharmaceuticals (PhACs) appear in the water and wastewater treatment plants mainly through excretion and/or improper disposal of outdated or unused medication (Boix et al., 2016). The presence of these contaminants has raised concerns since they are potentially toxic to aquatic organisms even at trace concentrations (ng/L or mg/L) (Yuan et al., 2013; Mirzaei et al., 2016). Besides, some emerging contaminants exhibit non-target effects as well as mixture toxicity in the environment. This means that PhACs have the potential to show specific effects, which are irrelevant to their therapeutic purposes. In addition, their incomplete mineralization may lead to the forma et al., tion of additional toxic chemicals (Escher et al., 2006; Catala 2015). These facts, as well as the low biodegradability of PhACs, are
L-H LP MET MP MR MWTP NF NXA OFX PAC PCT PCPs PhACs ph-F PILCs PPG RES S.E. SMX STP TC TOC TSS WWTPs
LangmuireHinshelwood low pressure metoprolol medium pressure molar ratio municipal wastewater treatment plant nanofiltration nalidixic acid ofloxacin powder activated carbon paracetamol personal care products pharmaceuticals photo-Fenton pillared interlayered clays procaine penicillin G resorcinol synthetic MWTP effluent sulfamethoxazole sewage treatment plant tetracycline total organic carbon total suspended solids wastewater treatment plants
significant concerns for public health and require urgent attention rez-Estrada et al., 2005b). Therefore, wastewaters containing (Pe PhACs should be treated prior to their discharge into the surface waters. In most cases, due to the poor removal of emerging contaminants in conventional treatment systems, these chemical compounds end up in soil, surface waters or even in drinking water (Mompelat et al., 2009; Hu et al., 2011; Matamoros et al., 2012; Veloutsou et al., 2014). Studies have demonstrated that about 64% of ECs are removed by less than 50% while 9% are not removed at all by conventional biological treatment processes (Chi et al., 2013). Several review papers have addressed the occurrence, fate and transport of ECs in different environmental compartments (Mompelat et al., 2009; Pal et al., 2010; Huerta-Fontela et al., 2011; Li, 2014; Tijani et al., 2016). Although some physical processes such as adsorption are effective for the removal of organic pollutants,
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
they are non-destructive and merely cause a phase transfer of contaminants, producing highly contaminant secondary wastes which require further treatment (Dantas et al., 2006; Mirzaei et al., 2013; Prieto-Rodríguez et al., 2013). In addition, the adsorption capacity of ECs on adsorbents (mostly GAC) strongly depends on environmental parameters such as pH, adsorbent concentration, presence of humic acid and salinity (Liu et al., 2009). A severe decrease in the adsorption of ECs on activated carbon was reported for the effluent of WWTPs as compared to surface waters (Snyder et al., 2007). Moreover, the widely used thermal regeneration of spent activated carbon is costly and indirectly causes an additional environmental problem (Georgi and Kopinke, 2005). Membrane processes have the ability to produce high quality effluents and reduce or remove trace amounts of micropollutants including ECs (Snyder et al., 2007; Liu et al., 2009). However, membrane filtration processes are often costly at full-scale operation because of frequent fouling and the requirement for backwashing that may limit their application, especially in the treatment of wastewaters at high flow rates (Chi et al., 2013; Ganiyu et al., 2015). The use of chemical techniques such as advanced oxidation processes (AOPs) may be considered as an alternative to this problem (Yuan et al., 2013). AOPs do not generate waste and they can be applied as a pre-treatment process prior to the biological treatment which can enhance the biodegradability of refractory compounds (Soon and Hameed, 2011). As an example of chemical techniques, the high removal efficiencies of a broad range of PhACs by ozonation have been reported (Nakada et al., 2007). However, concerns related to the potential toxicity of the partially oxidized PhACs during ozonation and the large energy consumption during ozone generation are considered as major obstacles to the commercial application of ozonation process (Nakada et al., 2007; Chi et al., 2013). Among AOPs, Fenton-type reactions have been identified as effective methods which produce hydroxyl radicals by the reaction between iron salts and hydrogen peroxide (Hartmann et al., 2010; Maezono et al., 2011). The advantages of these processes are the safe and environmentally-benign nature of reagents and relatively simple operating principles as well as short reaction time and the rez-Estrada et al., 2005b; absence of mass transfer limitations (Pe squez et al., 2014). Kajitvichyanukul and Suntronvipart, 2006; Vela Mackulak et al. (2015) successfully removed 13 drugs to concentrations below detection limit from a wastewater treatment plant by Fenton and Fenton-like processes. In addition, the Fenton process can be used as an effective pre-treatment to improve the biodegradability of wastewater contaminants, especially during the treatment of concentrated wastewaters containing recalcitrant compounds such as those emerging from drug factories and/or hospitals (Tekin et al., 2006; Sanchis et al., 2013). The photo-Fenton process is similar to the Fenton process but it offers higher removal efficiencies by employing irradiation. In this process, the generation of hydroxyl radicals accelerates in comparison with the “simple” Fenton process (Herney-Ramirez et al., 2010). However, the homogeneous Fenton and photo-Fenton reactions produce a secondary pollutant (ferric ion) following the treatment of organic contaminants (Sum et al., 2005). Thus, numerous attempts have recently been made to develop heterogeneous catalysts with high photocatalytic activity, long-term stability, and low cost for the Fenton and photo-Fenton oxidation in wastewater treatment. The heterogeneous Fenton process is a promising method that can allow operation at neutral pH and ambient temperature without the need to neutralize the effluent after the treatment (Navalon et al., 2010). This is because the heterogeneous catalysts are easier to separate from the effluent and they are noncorrosive and environmentally-benign (Soon and Hameed, 2011). The efficiency and performance of Fenton-type
667
processes, especially the heterogeneous Fenton processes for the destruction of pharmaceutical compounds, as an important category of emerging contaminants, and their removal from water have not been extensively reviewed. Therefore, this review addresses recent developments in the application of Fenton-type processes and the controlling parameters for PhACs degradation, as well as changes in the toxicity and biodegradability of treated water. Finally, the application of Fenton type processes as a pre-treatment step before conventional biological processes at pilot-scale operations for the removal of PhACs is considered.
2. Mechanism of Fenton-type reactions 2.1. Homogeneous Fenton processes The Fenton reaction is based on the generation of highly reactive hydroxyl radicals by the combination of peroxides (mainly hydrogen peroxide) and ferrous ions in acidic condition according to Eq. (1) (Kavitha and Palanivelu, 2004; Escher et al., 2006). Hydroxyl radicals are strong species that can remove one electron from any substance present in the solution to form hydroxide anion. Hydroxyl radicals can also gain a hydrogen atom from hydrocarbons to compensate for their missing atom. Considering the bond energy equal to 109 kcal/mol of OeH bond, any other bond such as CeH in organic compounds with a bond energy lower than this value is thermodynamically susceptible to oxidation (Navalon et al., 2010). At the first stage of reaction, the degradation of contaminants is very fast. The rapid decomposition of pollutants is attributed to the rapid production of hydroxyl radicals, owing to the presence of high Fe2þ species in the solution based on Eq. (1). At the second stage, due to the consumption of Fe2þ ions and generation of Fe3þ, the rate of reaction will decrease. This is because the reaction of Fe3þ and H2O2 leads to the production of HO2 radicals (E ¼ 1.65 V) which are weaker oxidants compared to the OH radicals (E ¼ 2.80 V) and have a lower rate of production (Eq. (2)) (de Luna et al., 2013). The generation of hydroxyl radicals and degradation efficiencies of contaminants are enhanced by the irradiation of UV light in the photo-Fenton process. The photo-Fenton process is based on the production of hydroxyl radicals by catalytic decomposition of hydrogen peroxide in acidic media and under irradiation. In order to prevent the precipitation of ferrous ion, this process is conducted at a pH of around 3. In the photo-Fenton reaction, the produced Fe3þ ions act as light-absorbing species to generate another hydroxyl radical based on Eqs. (3) and (4). In Fenton and photo-Fenton processes, the iron sludge formed after the reaction should be removed. However, in the photo-Fenton process, the required amount of ferrous catalyst and the volume of generated sludge are considerably lower than those in the Fenton process (Hermosilla et al., 2009). In addition, exploiting UV or solar light in the photoFenton process has important effects on the disinfection of treated water, through the inactivation of microorganisms (Rahim Pouran et al., 2015). The following reactions (Eqs. (1)e(8)) may take place in Fenton and Photo-Fenton processes. A list of complete reactions which may contribute to the Fenton and Photo-Fenton processes can be found elsewhere (Kusi c et al., 2006a).
Fe2þ þ H2 O2 /Fe3þ þ OH þ OH
(1)
Fe3þ þ H2 O2 /Fe2þ þ HO2 þ Hþ
(2)
Fe3þ þ H2 O þ hy/Fe2þ þ OH þ Hþ
(3)
668
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
Table 1 Homogeneous Fenton process for the removal of PhACs. [PhAC]/Initial concentration
Type of catalyst
Experimental conditions
Remarks and findings
[Sulfathiazole]0 ¼ 47 mmol/L
FeSO4$7H2O (Sigma eAldrich)
[Amoxicillin]0 ¼ 1000 mg/L
FeSO4$7H2O (Sinopharm Chemical Reagent)
squez et al., 84% PhAC removal was observed after 8 min reaction. (Vela 2014) 30% TOC removal was achieved after 60 min reaction. 40% COD removal was reported after 15 min reaction. Oxamic acid as a byproduct remained stable after long periods of reaction. (Guo et al., 2015) 80% removal of AMX was achieved. The Fenton reaction successfully applied as a pretreatment process for the conventional activated sludge process.
[Acetaminophen]0 ¼ 5 mM
FeSO4.H2O (Merck)
[Amoxicillin]0 ¼ 104 mg/L And [Ampicillin]0 ¼ 105 mg/L And [Cloxacillin]0 ¼ 103 mg/L
FeSO4$7H2O (Essex)
Distilled water pH 3 [Fe2þ] ¼ 192 mmol/L [H2O2] ¼ 1856 mmol/L Reaction time ¼ 60 min Distilled water pH 4 [Fe2þ ]/[H2O2] ¼ 20 Temperature ¼ 40 C Reaction time ¼ 70 min Distilled water pH 3 [Fe2þ] ¼ 0.5 mM [H2O2] ¼ 25 mM [Fe2þ ]/[H2O2] ¼ 0.02 Reaction time ¼ 20 min Distilled water pH 3 [H2O2]:[COD] ¼ 3:1(MR) [H2O2]:[Fe2þ] ¼ 10:1 (MR) [COD]:[H2O2]:[Fe2þ] ¼ 1:3:0.3 (MR) Reaction time ¼ 50 min
Drug manufacturing sewage [COD]0 ¼ 900e7000 mg/L
FeSO4$7H2O (Merck)
[Dipyrone]0 ¼ 50 mg/L
Modelled effluent pH 3.5 Room temperature [H2O2]:[Fe2þ] ¼ 155 (MR) [H2O2] ¼ 0.3 M [Fe2þ] ¼ 2 mM Reaction time ¼ 30 min (Fe(NH4)2(SO4)2$6H2O) Distilled water (Merck) Temperature ¼ 25 C pH 3.5 [H2O2]0 ¼ 22.5 mM [Fe2þ ]0 ¼ 2.25 mM Reaction time ¼ 50 min
Drug manufacturing sewage [COD]0 ¼ 4100 mg/L
FeSO4$7H2O (Merck)
Drug manufacturing effluent, pH 3 [COD ]0/[H2O2]0 ¼ 1:2.2 [Fe2þ ]0/[H2O2]0 ¼ 1:50 Reaction time ¼ 90 min
[Etodolac]0 ¼ 511 mg/L [COD]0 ¼ 18,000 mg/L
FeSO4$7H2O (Merck)
[Metronidazole]0 ¼ 1 mg/L
(FeSO4$7H2O) (Fisher Scientific)
Pharmaceutical wastewater Temperature ¼ 25 C pH 3 [H2O2]0 ¼ 1 M [Fe2þ ]0 ¼ 0.05 M Reaction time ¼ 30 min Deionized water pH 3.5 [H2O2]0 ¼ 1 mg/L [Fe2þ ]0 ¼ 11.76 mM Reaction time ¼ 5 min Distilled water pH 3.52 Temperature ¼ 25 C [H2O2]0 ¼ 1.39 104 mol/L [Fe2þ ]0 ¼ 1.25 105 mol/L [Fe3þ ]0 ¼ 1.68 105 mol/L Reaction time ¼ 120 min
[Carbamazepine]0 ¼ 4.98 mg/L (FeSO4$7H2O) and (Fe2(SO4)3.7H2O) (Merck)
Fe3þ þ H2 O2 þ hy/Fe2þ þ HO2 þ Hþ
(4)
OH þ H2 O2 /H2 O þ HO2
(5)
The photo-Fenton process has been shown to be a cost-effective advanced oxidation process. However, the associated costs may be reduced substantially when solar energy is used instead of artificial
Ref.
99.6% removal of AAP was achieved. (de Luna et al., Second order rate constant was 1.68 mM min1. 2013) Increasing initial concentration of reagent was shown a positive effect on AAP removal.
Complete antibiotics degradation was achieved after (Elmolla and Chaudhuri, 2010) 2 min of reaction. 80% COD removal was achieved after 50 min of reaction. 53% DOC removal was achieved after 50 min of reaction. Improvement in biodegradability was shown after treatment. First order rate constant of PhACs degradation was 0.0144 min1 and t1/2 was 69.3 min. No significant differences were shown at room (Tekin et al., 2006) temperature and 50 C. 52% COD removal was reported after treatment. About 85% BOD5 removal was achieved. Improvement in biodegradability and reduction in toxicity were shown after the treatment. 94.1% PhAC removal was achieved after 45 min treatment. 73.5% PhAC removal was shown after 2.5 min of oxidation. 42.78% TOC removal was obtained in 5 min. Biodegradability (BOD5/COD) was improved from ~0.1 to 0.62 in 10 min. 87% COD removal was achieved after 1.5 h. BOD5/COD was improved from 0.26 to 0.5. Fenton was very effective for pre-treatment of this type of wastewater. 97.5 and 99.6% removal of paracetamol and diclofenac were observed, respectively. 99.9% PhAC removal was achieved after treatment. 82% COD removal was obtained after reaction. Fenton process was an effective pre-treatment method to NF process.
No degradation of PhAC was observed without ferrous ions. 76% metronidazole removal was obtained after 5 min. First order rate constant of PhAC degradation was 0.066 min1 and t1/2 was 1 min. H2O2 initial concentration was the most influential of the studied parameters, followed by Fe2þ concentration, pH, and, Fe3þ ion concentration. Total degradation of CPZ was achieved at optimum conditions. In this integrated process, Fenton’s reaction was more important than Fenton-like reaction.
(Giri and Golder, 2014)
(Badawy et al., 2009)
(Vergili and Gencdal, 2015)
(Shemer et al., 2006)
(Domínguez et al., 2012)
light (Elmolla and Chaudhuri, 2010). The homogeneous Fenton reactions are widely used in water treatment (Hermosilla et al., 2009). However, the limitations of homogeneous reactions are mainly related to the high concentration (50e80 ppm) of Fenþ ions needed for effective removal of contaminants, which is significantly higher than the acceptable concentration in the effluent of water treatment (about 2 ppm) to be discharged into the environment
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
(Guo et al., 2015). Therefore, additional treatment is necessary to reduce the concentration of Fenþ ions in the effluent. Moreover, the treatment of iron-containing sludge which has been accumulated at the end of process is expensive and usually requires a large amount of chemicals (Bobu et al., 2008; Rache et al., 2014). Tables 1 and 2 provide an overview of the PhACs degraded by homogeneous Fenton and photo-Fenton processes, respectively, and summarize the mineralization efficiency of these contaminants under optimum conditions. 2.2. Heterogeneous Fenton processes As stated before, the homogeneous Fenton process has been widely used because of its simplicity, possibility of using conventional equipment and operation at ambient pressures and temperatures. Nevertheless, due to certain limitations such as high iron loss to the environment, need for iron recovery before the discharge of effluents to the receiving waters to comply with the standards (about 2 ppm), high H2O2 consumption and high operation cost, the development of heterogeneous Fenton process as an alternative to the homogeneous Fenton process has attracted much attention (Hermosilla et al., 2009). In addition, homogeneous Fenton processes have limitations such as the narrow range of operating pH and the formation of different Fe3þ complexes by changing the pH of solution (Garrido-Ramírez et al., 2010). As mentioned before, homogeneous Fenton-based processes proceed optimally at acidic conditions. Therefore, subsequent neutralization and the resulting precipitation of Fe(OH)3 as sludge will require additional handling and treatment (Kuan et al., 2015). A further limitation of the use of homogeneous Fenton process is the complexation of iron ions by the substances present in real effluents such as EDTA, or the generation of possible by-products such as oxalic acid (Hartmann et al., 2010). Finally, in the homogeneous Fenton process, catalyst deactivation may occur by iron complexing agents such as phosphate anions (Kuznetsova et al., 2004) (see in Tables 3 and 4). Using a heterogeneous Fenton-like process offers advantages compared to the homogeneous Fenton reactions. The heterogeneous Fenton reactions aim to expand the feasible operating pH range and reduce the problems associated with the separation of high dosage of iron ions that remain after the treatment (ArzateSalgado et al., 2016). An example of a simple Fenton-like solid catalyst is magnetite (Fe3O4), which can be easily removed from the solution through its magnetism (Kuan et al., 2015). Therefore, recent researches have focused on the development of stable heterogeneous catalysts in an effort to minimize leaching, while increasing catalytic activities and long-term stability (Hermosilla et al., 2009; Soon and Hameed, 2011). Improvements in the heterogeneous catalytic activity not only enhance the feasibility of Fenton process for wastewater treatment, but they also eliminate the need to pre-adjust the solution’s pH or to neutralize it at the end of the process (Bobu et al., 2008). However, the presence of a small fraction of iron on the surface of catalysts in the heterogeneous Fenton processes results in slow reaction kinetics compared to the homogeneous Fenton processes (Hermosilla et al., 2009; Bae et al., 2013). While the homogeneous Fenton processes depend only on chemical reactions between the reagents and contaminants, in the heterogeneous Fenton reactions physical processes on the surface of catalysts in addition to chemical reactions control the overall outcome of the process. The physical processes include surface reactions and desorption while the chemical reactions refer to those between contaminants and the produced radicals (Soon and Hameed, 2011). In the heterogeneous Fenton reaction three possible mechanisms can occur: (i) chemisorption of
669
pharmaceuticals on the surface of catalysts; (ii) homogeneous transformation of hydrogen peroxide to hydroxyl radicals by iron leaching to the reaction solution and/or (iii) reaction of hydrogen peroxide with the iron species on the surface of catalysts and its decomposition to hydroxyl radicals (Rahim Pouran et al., 2015). Therefore, potential mass transfer limitations affect the resulting reaction rates (Kuan et al., 2015). In the heterogeneous Fenton process, the catalyst characteristics such as surface area, pore volume, density, porosity, pore size and distribution need to be taken into consideration. Hence, the catalyst surface can be designed or modified to achieve higher catalytic performance such as activity, selectivity and stability (Soon and Hameed, 2011). In heterogeneous Fenton processes, the iron ions can leach from the catalyst during the reaction. This phenomenon not only causes a loss in activity with time, it also produces secondary metal ion contaminants (Hartmann et al., 2010). Currently, most studies on heterogeneous Fenton processes focus on the stability and reaction kinetics rather than the process mechanism such as the leaching of iron ions into the aqueous phase and its effect on the process. The difficulties and complications in identifying pertinent chemical pathways have hindered the identification of relevant mechanisms. Dissolved homogeneous ions can be deactivated and they can precipitate as a solid phase, especially due to pH variations as a result of the generation of intermediate compounds during the reaction (Kuan et al., 2015). Extensive efforts have been made to find an effective solid catalyst for the iron ions to overcome major disadvantages of the homogeneous Fenton process. 2.2.1. Iron oxide minerals Iron oxide minerals such as magnetite (Fe3O4), hematite (Fe2O3), goethite (a-FeOOH), pyrite (FeS2) and lepidocrocite (g-FeOOH), which are common constituents of soil, can serve as potential environmentally-benign material for the elimination of emerging contaminants. Moreover, these minerals are widespread in the environment and can be utilized in in situ removal processes (Sun and Lemley, 2011; Bae et al., 2013; Velichkova et al., 2013). Due to their specific properties, such as stability, reusability and ease of separation compared to other iron minerals and ferrous salts, and their environmental compatibility, magnetite (Fe3O4) has attracted much attention for use in Fenton processes (Sun et al., 2013). Magnetite can be formed naturally via different pathways such as Fe2þ oxidation, iron metal corrosion and chemical and biological reduction of Fe3þ ions. Magnetite can oxidize organic compounds into less hazardous or more biodegradable forms. However, total mineralization can be achieved only by Fenton-like processes in the presence of hydrogen peroxide (Sun and Lemley, 2011). In addition, Fe3O4 nanoparticles have a tendency to agglomerate due to the intrinsic magnetic interactions and van der Waals forces. Agglomeration can reduce the surface/volume ratio and catalytic activities (Jaafarzadeh et al., 2015). It seems that pyrite (FeS2) is a very effective heterogeneous catalyst in the Fenton process. Using pyrite has two positive effects since it provides aqueous Fe2þ ions, while decreasing the pH of solution to the proper level (normally pH 3e4) (Bae et al., 2013). The use of immobilized iron on support material can also be considered as an effective method to overcome the abovementioned limitations of homo/heterogonous Fenton type processes (Sun and Lemley, 2011; Sun et al., 2013). 2.2.2. Support material in heterogeneous Fenton processes Supported metal (mostly iron) compounds and iron-coated particles show potential to overcome the problems associated with homogenous Fenton reactions such as the limited operational range of pH, release of iron into the effluent, removal of the iron
670
Table 2 Homogeneous photo-Fenton process for the removal of PhACs. Type of catalyst
Experimental conditions
Source of energy
[Metoprolol]0 ¼ 50 mg/L
FeSO4$7H2O (Panreac PA)
BLB (3 8 W, lmax ¼ 365 nm) or solar Deionized water pH 3 irradiation or artificial solar irradiation Temperature ¼ 25 C (Xe lamp 1 kW) or UVC lamp [Fe2þ] ¼ 10 mg/L [H2O2] ¼ 150 mg/L Reaction time ¼ 150 min
[ATL]0 ¼ 20 mg/L and [Metoprolol]0 ¼ 20 mg/L
Fe(ClO4)3 (Sigma eAldrich)
[Diclofenac]0 ¼ 50 mg/L
Fe2 (SO4)3 H2O
[Sulfathiazole]0 ¼ 47 mmol/L
FeSO4$7H2O (Sigma eAldrich)
[Amoxicillin]0 ¼ 104 mg/L, [Cloxacillin]0 ¼ 105 mg/L, [Ampicillin]0 ¼ 103 mg/L
FeSO4$7H2O (Essex)
[Sulfonamides]0 ¼ 200 mg/L
FeSO4$7H2O (Panreac PA)
[Sulfonamides]0 ¼ 200 mg/L
FeSO4$7H2O (Panreac PA)
Distilled water, pH about 2.9 Temperature ¼ 30 e35 C [iron ions] ¼ 2.8 e5 mg/L [H2O2] ¼ 95 e100 mg/L Reaction time ¼ 180 min Distilled water pH 7 (with no buffer capacity) Temperature ¼ 30 e40 C [H2O2] ¼ 15 mM [Catalyst] ¼ 0.05 mM Reaction time ¼ 110 min Distilled water, pH 3 [Fe2þ] ¼ 157 mmol/L [H2O2] ¼ 1219 mmol/ L Reaction time ¼ 60 min Distilled water, pH 3 [H2O2]/[Fe2þ ] ¼ 20 Temperature ¼ 22 C [COD]0 ¼ 520 mg/L [H2O2]/[COD] ¼ 1.5 Reaction time ¼ 50 min Distilled water, pH 2.8 Temperature ¼ 23 e31 C (not controlled) [Fe2þ] ¼ 10 mg/L [H2O2] ¼ 400 mg/L Reaction time ¼ 100 min Distilled water, pH 2.8
UV 125 W HP, Hg lamp (lmax ¼ 290 nm)
Remarks and findings
Ref.
Photolysis alone, hydrogen peroxide alone and photolysis with iron were not able to eliminate MET. 100% MET removal after 7 min and 81.2% TOC removal after 90 min were achieved by BLB lamp. By using artificial solar 97.3% of MET elimination after 7 min and 78.8% of TOC conversion after 120 min were observed. With UVC, 100% of MET elimination after 20 min and 17.6% of TOC conversion after 60 min were reported. 100% removal of PhACs was achieved after 1 min. Total mineralization was taken about 150 min. Conduction experiment in river and lake water was caused a delay in degradation and mineralization and increased H2O2 consumption. First order rate constant of ATL degradation was 0.024 min1, and t1/2 was 28.4 min. First order rate constant of metoprolol degradation was 0.048 min1, and t1/2 was 14.3 min.
(Romero et al., 2016b)
(Veloutsou et al., 2014)
Solar irradiation
rez-Estrada After 50 min of reaction, 100% removal of PhAC was achieved. (Pe About 98% DOC removal was also achieved after 110 min of reaction. et al., 2005b) The H2O2 amount needed for total oxidation was about 8 times more than theoretical stoichiometric demand.
6 20 W solarium lamps (365 nm) 3.5 mW cm2
UV Lamp 6W (365 nm)
Complete degradation of amoxicillin, cloxacillin and ampicillin were (Elmolla and achieved in 2 min. Chaudhuri, 2009) 58.4% DOC removal was achieved in 50 min. 80.8% COD removal was achieved in 50 min. Biodegradability (BOD5/COD) was improved from 0 to 0.4 after 50 min which is considered adequate for biological treatment.
LP UV lamp Hg vapor 88W (l ¼ 350e400 nm)
95% removal was achieved after 60 min reaction. 75% TOC removal was reported after 60 min reaction. 90% COD removal was achieved after 15 min reaction. Photo-Fenton reaction was twice faster than Fenton reaction.
Complete degradation of PhAC was achieved in 50 min. 56.4% TOC removal was shown after 94 min. 82.2% COD removal was achieved after 94 min. The energy consumption was estimated 2.5 kJ/L in this research.
87% mineralization was achieved. 96% COD removal was achieved after 94 min.
(Vel asquez et al., 2014)
(Gonzalez et al., 2009)
(Gonzalez et al., 2009)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
[PhAC]/Initial concentration
[Amoxicillin]0 ¼ 138 mg/L [Cloxacillin]0 ¼ 84 mg/L [COD]0 ¼ 670 mg/L [DOC]0 ¼ 145 mg/L [BOD5]0 ¼ 70 mg/L
FeSO4$7H2O (Essex)
[Diclofenac]0 ¼ 0.12 mM
FeCl3.6H2O (Fluka)
Effluent from antibiotic industry pH 3 [H2O2]/[COD] ¼ 2.5 [H2O2]/[Fe2þ] ¼ 20 [H2O2] ¼ 44.9 mM Reaction time ¼ 30 min Distilled water pH 2.8 Temperature ¼ 50 C [Fe ions] ¼ 14 mg/L [H2O2] ¼ 340 mg/L Reaction time ¼ 60 min Distilled water or effluent from STP
Solar radiation (300 < l < 500 nm) z121.6 W/m2
The energy consumption was estimated 19 kJ/L.
Solar radiation
UV lamps 10 W (lmax ¼ 254 nm)
Total degradation of paracetamol, amoxicillin, ampicillin and diclofenac were achieved after 60, 90, 120, and 120 min of irradiation, respectively. First order rate constant of paracetamol, amoxicillin, ampicillin and diclofenac degradation was 0.070 min1, 0.056 min1, 0.042 min1 and 0.51 min1, respectively. The biodegradability (BOD5/COD) was enhanced from 0.3 to 0.52. The removal of BOD5, TOC and COD were reported at 61%, 52% and 77%, respectively. Photo-Fenton was an efficient pre-treatment process in order to improve biodegradability. TOC, COD and BOD5 removal were decreased dramatically with increasing initial COD concentration.
15 W Blacklight lamp UVA (320e400 nm) 19 W m-2
The degradation was observed only when three key components of photo- (Batista and Nogueira, 2012) Fenton were present (Light, Fe and H2O2). 92 and 90% mineralization was achieved after 42 min of reaction in the case of sulfadiazine and sulfathiazole, respectively.
UV lamp 9 W (362 nm) 12 W m-2
LP UV (254 nm) 1.95 mW cm-2 & MP UV (200e400 nm) 1.9 mW cm-2
6% removal was obtained by direct photolysis with LP while 12% removal was achieved by MP. For both lamps about 60% removal was observed by adding 25 mg/L of H2O2 after 2.5 min. 94% removal was achieved by the photo-Fenton process under optimum conditions. First order rate constant of PhAC degradation was 0.23 min1, and t1/2 was 0.5 min. Complete removal of the antibiotics occurred in 1 min. 59 and 49% soluble COD and DOC removal were achieved after 30 min. Biodegradability (BOD5/COD) was improved to 0.44 after treatment.
UV lamp 6 W (365 nm)
UV 400 W (254 nm)
(Alalm et al., 2015)
(Kajitvichyanukul and Suntronvipart, 2006)
(Giri and Golder, 83.2% drug removal was shown at 2.5 min of oxidation. 2014) 96.4% PhAC removal was achieved after 45 min reaction. In the absences of ferrous ion, the PhAC removal was 74.4% in 45 min. 56.0% TOC removal was obtained in 5min. Biodegradability (BOD5/COD) was improved from ~0.1 to 1.51 after 10 min of reaction. (Shemer et al., 2006)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
Temperature ¼ 26 C [Fe2þ] ¼ 10 mg/L [H2O2] ¼ 550 mg/L Reaction time ¼ 100 min Distilled water, pH 3 [Paracetamol]0 ¼ [Amoxicillin]0 ¼ FeSO4$7H2O (Sigma [Ampicillin]0 ¼ [Diclofenac]0 ¼ eAldrich) [Fe2þ] ¼ 500 mg/L 100 mg/L [H2O2] ¼ 1500 mg/L Reaction time ¼ 120 min FeSO4$7H2O Hospital effluent Hospital effluent [TSS] ¼ 115 (mg/L); [COD] ¼ 1350 (Merck) pH 3 (mg/L); [Fe2þ] ¼ 135 mg/L [BOD5] ¼ 410 (mg/L); [COD]:[ Fe2þ] ¼ 1:0.1 [TOC] ¼ 1050 (mg/L) [COD]:[H2O2] ¼ 1:4 Temperature ¼ 25 e31 C Reaction time ¼ 120 min [Sulfadiazine]0 ¼ 25 mg/L, Prepared Distilled water, pH [Sulfathiazole]0 ¼ 25 mg/L (K3Fe(C2O4)3$3H2O) 2.5 [Fe3þoxalalate] ¼ 0.2 mM [H2O2] ¼ 5 mmol/L Reaction time ¼ 45 min [Dipyrone]0 ¼ 50 mg/L (Fe(NH4)2(SO4)2$6H2O) Distilled water, (Merck) Temperature ¼ 25 C pH 3.5 [H2O2]0 ¼ 22.5 mM [Fe2þ ]0 ¼ 2.25 mM Reaction time ¼ 45 min [Metronidazole]0 ¼ 1 mg/L FeSO4$7H2O (Fisher Deionized water Scientific) pH 3.5 [H2O2]0 ¼ 1 mg/L [Fe2þ ]0 ¼ 11.76 mM Reaction time ¼ 5 min
(Elmolla and Chaudhuri, 2011)
A direct proportional between the DCF and TOC removal and UV light (Ravina et al., 2002) intensity was reported. Total DCF removal was achieved after only 0.5 min while about 50 min was required for total TOC removal.
et al., (Trovo 2008) 671
(continued on next page)
672
Table 2 (continued ) [PhAC]/Initial concentration
Type of catalyst
[Amoxicillin]0 ¼ 42 mg/L, or [Bezafibrate]0 ¼ 13 mg/L, or [Paracetamol]0 ¼ 20 mg/L
(K3Fe(C2O4)3$3H2O) or pH 2.5 [catalyst] ¼ 0.2 mM Fe(NO3)3 [H2O2] ¼ 5.0 mM (Mallinckrodt) Reaction time ¼ 30 min
[Tetracycline]0 ¼ 24 mg/L
(K3Fe(C2O4)3$3H2O) Mallinckrodt or Fe(NO3)3
[Diclofenac]0 ¼ 50 mg/L
[Paracetamol]0 ¼ 50 mg/L
[Ofloxacin]0 ¼ 10 mg/L
Drug formulation effluent [COD]0 ¼ 600 mg/L
Distilled water or Effluent from STP pH 2.5 [catalyst] ¼ 0.2 mM [H2O2] ¼ 10 mM Reaction time ¼ 60 min FeSO4$7H2O Distilled water pH 7 Temperature ¼ 30 e40 C [H2O2] ¼ 15 mM [Catalyst] ¼ 0.05 mM Reaction time ¼ 110 min FeSO4$7H2O (Panreac Distilled water PA) pH 2.8 Temperature ¼ 25 C [H2O2] ¼ 300 mg/L [Catalyst] ¼ 10 mg/L FeSO4$7H2O (POCH SA) Distilled water or (K3Fe(C2O4)3$3H2O) pH 2.5e2.8 Temperature ¼ 25 C (Mallinkrodt) [H2O2] ¼ 120 mg/L [Catalyst] ¼ 0.05 mM Reaction time ¼ 300 min FeSO4$7H2O WWTP effluent (Riedel-de Haen) (secondary treated) pH 3 Temperature ¼ 25 C [H2O2] ¼ 2.714 mM [Catalyst] ¼ 5 mg/L Reaction time ¼ 120 min Fe(NO3)3$9H2O
Drug formulation effluent pH 3 [H2O2] ¼ 25 mM [Catalyst] ¼ 1.5 mM Reaction time ¼ 30 min
Source of energy
Remarks and findings
Ref.
Black-light lamp (15 W fluorescent (for 98% removal of BZF and PCT was achieved after 5 min of reaction using FeOx; AMX and PCT), lmax ¼ 365, 19 W m2) whereas, by using Fe(NO3)3, 89 and 53% removal was achieved respectively & solar light (20 W m2 for BZF) at the same time. Total degradation of AMX was achieved after 0.5 min of reaction using both catalysts. TOC removal rate was higher when using FeOx for all considered compounds. First order rate constant of BZF and PCT, was 0.94 and 0.70 min1, respectively at optimum conditions. Black-light lamp (15 W fluorescent, Total removal of TC was achieved under black-light by using Fe(NO3)3 after (Bautitz and 1 min. Also, 72% TOC removal was obtained after 60 min reaction, while by Nogueira, 2007) lmax ¼ 365, 19 W m2) or solar light using FeOx, more reaction time (8 min) was needed for 100% removal of TC. (15e20 W. m2) Under solar irradiation, 100% removal was achieved by FeOx in 0.5 min while, Fe(NO3)3 needed 3 min for total removal in the same condition.
rez-Estrada (Pe et al., 2005a)
Solar irradiation
Total DCF removal was achieved after 60 min exposure to sunlight. 100% DOC removal (as total mineralization) was shown after 100 min. About 18 intermediates were identified during the mineralization.
UV lamp lmax ¼ 365 3 (8 W)
lez et al., (Gonza About total removal was reported at the end of the reaction. (BOD5/COD) was improved with increasing H2O2 dose and reached to the 2007) 0.25 by using 400 mg/L H2O2. The effluent was less toxic after the treatment.
Solar radiation simulator (1100 W xenon arc lamp)
et al., No photolysis of PCT was observedat natural pH (4.2) after 300 min (Trovo irradiation. 2012) 100% PCT removal was achieved after 120 min by FeSO4 and 180 min by FeOx. 79% TOC removal was reported by FeSO4 and 58% by FeOx after 300 min.
Solar radiation simulator (1 kW Xenon lamp)
Less than 6% OFX removal was observed after 120 min by photolysis. Total removal of OFX was achieved after 90 min by Fenton reaction while by photo-Fenton reaction 30 min was required. Photo-Fenton was more efficient than TiO2 not only for the OFX degradation but also for the DOC removal. At optimum conditions, 50% and 10% DOC removal were achieved by PhotoFenton and TiO2, respectively. First order rate constant of OFX degradation by the photo-Fenton process was 0.1128 min1, and t1/2 was 6.14 min. 56 and 44% COD removals were achieved by photo-Fenton and Fenton process, respectively. The BOD5/COD ratio was increased from 0.10 to 0.45 and 0.24 after application of the photo-Fenton and Fenton, respectively. 35 and 42% TOC removal were reported by Fenton and phot-Fenton processes respectively.
UV-A 125 W (365 nm)
(Michael et al., 2010)
(Arslan-Alaton and Gurses, 2004)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
[Sulfamethoxazole]0 ¼ 200 mg/L
Experimental conditions
[Metoprolol]0 ¼ 50 mg/L and [Resorcinol]0 ¼ 50 mg/L
[Sulfamethoxazole]0 ¼ 50 mg/L
[Ibuprofen]0 ¼ 0.87 mM
FeSO4.7H2O (Panreac)
Deionized water pH 6.2 (natural) Temperature ¼ 25 C [H2O2] ¼ 150 mg/L [Catalyst] ¼ 10 mg/L Reaction time ¼ 180 min FeSO4$7H2O (Sigma Distilled water eAldrich) Or Sea water pH 2.5e2.8 Temperature ¼ 25 C [H2O2] ¼ 210 mg/L [Catalyst] ¼ 5.2 mg/L Reaction time ¼ 7 h FeSO4$7H2O (Panreac) Distilled water pH 3 Temperature ¼ 30 C [H2O2] ¼ 0.32 mM [Catalyst] ¼ 1.2 mM Reaction time ¼ 120 min FeSO4$7H2O (POCH SA) or (K3Fe(C2O4)3$3H2O) (Mallinkrodt)
Combination of six emerging contaminants
Fe(ClO4)3$H2O (Sigma eAldrich)
Pharmaceutical laboratory effluent FeSO4$7H2O Merck
Distilled water pH 2.5e2.8 Temperature ¼ 25 C [H2O2] ¼ 120 mg/L [Catalyst] ¼ 0.05 mM Reaction time ¼ 240 min Secondary effluent from a MWTP pH 3 [H2O2]/[Fe3þ] ¼ 6.1 (mass ratio) [Catalyst] ¼ 2.8 mg/L Reaction time ¼ 5 h Diluted waste water Ambient temperature [TOC]0 ¼ 400 ppm pH 2.7 [H2O2] ¼ 2500 ppm [Catalyst] ¼ 20 ppm Reaction time ¼ 120 min
Solar radiation simulator (1100 W xenon arc lamp)
Xe lamp 1 kW 6.9 mE s1 (290e400 nm)
Solar radiation simulator (1100 W xenon arc lamp)
Solar radiation
Solar radiation
Adding resorcinol (ligand iron complex) to the solution let to perform the reaction in neutral pH condition. By Fenton reaction, 100 and 92% removal were achieved for MET and RES after 20 min respectively. By photo-Fenton reaction, 100 and 94.4% removal were achieved for MET and RES after 3 min respectively. Biodegradability increased at the end of the processes. No hydrolysis was reported at the pH studied (2.5, 4.8 and 9.0) after 48 h. After the process, the toxicity was increased from 16 to 86% in seawater. Total degradation of SMX was achieved after 16 min in distilled water and 105 min in seawater.
UV/H2O2 process led to the 30% IBP removal after 2 h, however, TOC did not decrease. Maximum 60% and 10% IBP and TOC removal were achieved by Fenton process, About 100% IBP (after 60 min) and 40% TOC removal were reported by the photo-Fenton reaction. Biodegradability was improved at the end of the process (BOD5 was reached to less than 1 mg/L from 25 mg/L). No AMX removal by photolysis was observed at natural pH (6.2) after 6 h irradiation. 100% removal of AMX was achieved after 5 min by FeOx and 15 min by FeSO4. After 240 min, between 73 and 81% TOC removal was reported. 73 and 81% TOC removal was obtained in the presence of FeSO4 and FeOx, respectively. Total ECs removal were achieved in less than 90 min. At the optimum point, 2.5-fold increase of the BOD5/COD ratio was shown after treatment. About 35% TOC removal was obtained after 5 h treatment.
(Romero et al., 2016a)
et al., (Trovo 2009)
ndez-Arriaga (Me et al., 2010)
et al., (Trovo 2011)
~ ones et al., (Quin 2015)
sito et al., (Expo About 80% TOC removal was achieved after treatment. Some acids like acetate, formate and oxalate were formed during the 2016) oxidation.
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
[Amoxicillin]0 ¼ 50 mg/L
BLB 8W (365 nm)
673
674
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
sludge after the treatment, and slow reaction kinetics of the heterogeneous Fenton processes (Hu et al., 2011). Therefore, increasing attention has been paid to the development of support medium for the heterogeneous Fenton catalysts such as pillared clays (Feng et al., 2006; Li et al., 2006, 2015; Bobu et al., 2008), activated carbon (Georgi and Kopinke, 2005; Dantas et al., 2006; Zazo et al., 2006; Jaafarzadeh et al., 2015; Lan et al., 2015), alumina (Bautista et al., 2010a; Kunde and Yadav, 2015) and zeolite (Kusi c et al., 2006b; Kasiri et al., 2008; Tekbas¸ et al., 2008; Aleksi c et al., 2010; Perisic et al., 2016). Using mesoporous material as the catalyst support in Fenton or Fenton-like processes is a promising approach due to the synergistic effect of oxidation and adsorption (Song et al., 2015). In addition, the incorporation of iron oxide nanoparticles into porous support materials such as activated carbon reduces the strong tendency of agglomeration of nanoparticles due to intraparticle interactions. Moreover, immobilization of metal catalysts into porous adsorbents reduces the risk of iron release in the treated effluent (Jonidi Jafari et al., 2016). Zhang et al. (2016b) used mesoporous manganese oxide with a high surface area as a support material for copper in Fenton-like processes for the removal of BZA at neutral pH. A simultaneous adsorption on activated carbon along with oxidation by Fenton reaction of TC is also reported (Jonidi Jafari et al., 2016). Activated carbon can be considered as a promising support for the heterogeneous Fenton reaction owing to its high surface area, wide availability, high stability in acidic or alkaline environments and low cost (Jaafarzadeh et al., 2015; Lan et al., 2015). Exploiting carbon materials as the support in heterogeneous Fenton processes is an innovative approach because it provides simultaneous adsorption and catalytic oxidation of soluble contaminants (Dantas et al., 2006). Whether the adsorption process is beneficial or not is not fully clear. Most investigators state that the adsorption offers advantages and will improve the efficiency of oxidation (Ramirez et al., 2007). Nevertheless, Georgi and Kopinke (2005) considered adsorption as a disadvantage because they claimed that the contaminants adsorbed on the activated carbon were nearly nonreactive and could not be degraded by hydroxyl radicals. It should be noted that the catalytic activity of activated carbon even without using iron for the decomposition of H2O2 to hydroxyl radicals has been demonstrated (Ramirez et al., 2007). The catalytic activity of activated carbon is related to its surface chemistry. However, the oxidation of contaminants by the hydroxyl radicals that are produced by bare activated carbon in the presence of H2O2 needs a long reaction time to achieve a sufficient extent of removal. Therefore, despite several advantages of bare activated carbon as a Fenton-like catalyst, such as the lack of iron leachate or lower price, using iron metal in carbon structure is crucial when shorter operation times are needed (Ramirez et al., 2007). In another work, porous sea buckthorn was utilized as the support material for (bFeOOH) nanoparticles for the removal of doxycycline antibiotic at neutral pH condition (Zhang et al., 2016a). Porous sea buckthorn is a cost-free agricultural waste residue and can serve as porous support in Fenton processes to enhance the removal efficiency, owing to the synergistic effect of adsorption and Fenton-like reactions. Besides, Fenton processes can serve as an in situ regeneration step for bioadsorbents, since the generated hydroxyl radicals can effectively oxidize accumulated organic contaminants on the adsorbents (Song et al., 2015; Zhang et al., 2016a). Zeolites are hydrated aluminosilicate materials with external and internal surface areas up to several hundred square meters per gram, high cation exchange capacity and cage-like structures. Zeolites in natural or synthesized forms are used in the industry as the catalyst support, adsorbent and ion exchange resins (Tekbas¸ et al., 2008). Due to their high surface area and porosity, zeolites can act as an adsorbent in order to remove contaminants from aquatic
media (Kasiri et al., 2008). Therefore, by using zeolites as a support medium in the heterogeneous photo-Fenton processes, the adsorption process should also be taken into account besides homo/heterogeneous Fenton reaction and direct photolysis of H2O2 and contaminants (Kusic et al., 2006b). In addition, the presence of a strong electrostatic field inside the pores of the zeolite slows down the leaching of iron from Fe-zeolite catalyst during the reaction (Kasiri et al., 2008). Moreover, the preparation of Fe-zeolite catalyst is relatively simple due to the zeolite’s cation exchange capability (Kusi c et al., 2006b). Pillared clays, also known as pillared interlayered clays (PILCs) or cross-linked clays, are among the most-widely studied microporous materials that can be used as support media due to their unique characteristics such as inertness, high surface area, abundance, low cost as well as the simplicity of the pillaring process (Li et al., 2006, 2015). These materials are prepared by exchanging the interlayer cations of layered clays with inorganic polyoxocations (Bobu et al., 2008). Mixed metal oxide pillared clays such as FeeAl or FeeCr may improve the catalyst performance by increasing the surface area or by modifying surface properties such as the acidity (Li et al., 2006). Good catalyst activity and stability of Fe-clay catalyst in removing ciprofloxacin by the heterogeneous photoFenton process was reported (Bobu et al., 2008). Similar to pillared clays and zeolite, alumina can also act as a support and adsorbent. Due to its characteristics, including chemical inertness, mesoporous nature, high surface area and thermal stability, alumina is widely used in industrial operations (Kunde and Yadav, 2015). 3. Major controlling parameters in the Fenton reaction 3.1. Effect of reagents concentration The maintenance of proper concentrations of hydrogen peroxide and iron is important in the homogeneous Fenton process. An excess or deficiency of these reagents can considerably decrease the efficiency of process (Bautitz and Nogueira, 2007). In addition, the use of appropriate concentrations of reagents will prevent extra costs of operation that may result from the use of excess reagents, and will reduce the encountered difficulties in removing the excess iron, according to the effluent standards. In Fenton-based processes, the presence of H2O2 is essential because it is the main source of OH radicals during the process (Bobu et al., 2008). However, due to the radical scavenging effect of H2O2 at high concentrations (Eq. (5) and Eq. (6)), and the auto-decomposition of H2O2 to O2 and H2O (Eq. (7)), the optimum concentration of H2O2 should be determined in Fenton-based processes (Badawy et al., 2006; Li et al., 2006; Domínguez et al., 2012). Moreover, high concentrations of H2O2 may reduce the COD removal efficiency (Elmolla and Chaudhuri, 2009). The consumption of H2O2 is also a critical issue due to its major contribution to the operating cost of Fenton processes (Zazo et al., 2011; Vergili and Gencdal, 2015). It should be noted that the unconverted H2O2 cannot be recovered and has to be eliminated before discharging the final effluent into the water bodies, due to potential toxicity effects (Bautista et al., 2010a). The optimum concentration of hydrogen peroxide depends on the type and concentration of contaminants and Ferrous ion (Elmolla and Chaudhuri, 2010).
HO2 þ OH /H2 O þ O2
(6)
2H2 O2 /2H2 O þ O2
(7)
The use of an adequate concentration of reagents in the Fenton process is important. However, the H2O2/iron molar ratio is the key
Table 3 Heterogeneous Fenton-like process for the removal of PhACs. [PhAC]/Initial concentration
Type of catalyst
Experimental conditions
Source of energy
Remarks and findings
Ref.
[Paracetamol]0 ¼ 100 mg/L
MGN1 (Fe3O4 powder < 50 nm); MGN2 (Fe3O4 powder < 5 nm); MGM (g-Fe2O3 powder < 50 nm) (All from Sigma-Aldrich)
Distilled water, pH 2.6 Temperature ¼ 60 C [Catalyst] ¼ 6 g/L [H2O2] ¼ 153 mM Reaction time ¼ 5 h
Without external energy
(Velichkova et al., 2013)
Combination of 11 detected drugs of abuse [TOC]0 ¼ 60 mg/L
Powder silica-supported iron oxide (Fe2O3/SBA-15)
Raw river water pH 3 Temperature ¼ 22 C [Catalyst] ¼ 0.1e0.6 g/L [H2O2] ¼ 15e60 mg/L Reaction Time ¼ 6 h
UVeVisible 150 W MP mercury lamp (l > 313 nm)
[Paracetamol]0 ¼ 20 mg/L
Synthesized CoFe2O4
UV lamps 4 15 W (lmax ¼ 365 nm)
[Paracetamol]0 ¼ 104 mol/L
Synthesized Goethite (aFeOOH)
Distilled water, pH 3 Temperature ¼ 45 C [Catalyst] ¼ 200 mg/L [H2O2] ¼ 50 mM Reaction time ¼ 60 min Ultra-pure water, pH 3 [H2O2] ¼ 5 103 mol/L [catalyst]0 ¼ 1 g/L Temperature ¼ 20 C Reaction time ¼ 180 min
[Diclofenac]0 ¼ 30 mg/L
Metallurgical slag (60.90% Fe2O3)
Distilled water, pH 7 [H2O2] ¼ 180 mg/L [catalyst]0 ¼ 26.6 mg/L Temperature ¼ 35 C Reaction time ¼ 300 min
Simulated sunlight (300e500 nm) 500 W/m2
[Ofloxacin]0 ¼ 30 mg/L
Synthesized alginate-iron catalyst (4% Iron)
Distilled water, pH 3 [H2O2] ¼ 4.067 mM [catalyst]0 ¼ 400 mg/L Temperature ¼ 25 C Reaction time ¼ 180 min
Without external energy
[Carbamazepine]0 ¼ 15 mg/L or [Ibuprofen]0 ¼ 15 mg/L
Nano-Fe3O4 (Nanostructured & Amorphous MaterialsInc)
Without external energy
[Diclofenac]0 ¼ 20 mg/L
Synthesized Fe-doped CeO2
Distilled water, pH 7.0 [H2O2] ¼ 600 mM [catalyst]0 ¼ 1.84 g/L Temperature ¼ 23 C Reaction time ¼ 12 h Distilled water, pH 5 [H2O2] ¼ 10 mM
Without catalyst, paracetamol was not oxidized by H2O2. For three catalysts, total PhAC degradation was obtained after 5 h oxidation. 43%, 34% and 39% mineralization were achieved with MGN1, MGN2 and MGM, respectively. More than 70% removal of all drugs were achieved using only UVevis light. Using 0.1 g/L of catalyst without UVevis and H2O2 led to strong reduction of the opioid (75e83%), moderate reduction of benzodiazepines (32e56%) however a strong increase of cocaine (225e159%) and a moderate increase of ALCs (51e11%) were observed. The best result (>75% removal of all drugs and 86% of TOC removal) was achieved by a combination of UV and 0.6 g/L catalyst. More than 99% removal was achieved. First order rate constant of EC degradation was 0.0967 min1 The presence of phenol in solution was reduced the efficiency of removal. No detectable adsorption onto Goethite surface was observed. With the 1 mM sensibility detection, Fe2þ leachate was not detected. No degradation was shown after 3 h of direct photolysis. At optimum conditions, the first order reaction rate was 2.33 101 h Complete degradation of PhAC was achieved in 90 min. First order rate constant of PhAC degradation was 0.046 min1. 87% mineralization was achieved after 300 min treatment. 98% EC removal was achieved after treatment. About 10% decrease of support activity was shown after three successive runs. A low (1.2%) iron leaching has been reported. Increasing temperature has a substantial effect on PhAC elimination. The first order rate constant was 0.0356 (min1). 90.0% and 81.1% removal were achieved for CBZ and IBP, respectively. The first-order degradation rate of CBZ and IBP were 0.182 h1 and 0.127, respectively.
Without external energy
Removal efficiency of PhAC with H2O2 alone was only 2.2% in 40 min.
(Rad et al., 2015)
(Mameri et al., 2016)
(Arzate-Salgado et al., 2016)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
LP mercury UV lamps (365 nm) 1840 mW cm2
(Catal a et al., 2015)
(Titouhi and Belgaied, 2016)
(Sun et al., 2013)
(Chong et al., 2016) 675
(continued on next page)
[PhAC]/Initial concentration
676
Table 3 (continued ) Type of catalyst
Experimental conditions
Source of energy
[catalyst]0 ¼ 0.5 g/L Reaction time ¼ 40 min
Synthesized Fe-ZSM5
Distilled water, pH 4 [H2O2] ¼ 50 mM [catalyst]0 ¼ 2 mM Temperature ¼ 25 C Reaction time ¼ 120 min
UVA lamps (lmax ¼ 365 nm) 3.04 107 E s1
[Norfloxacin]0 ¼ 50 mg/L
Synthesized Fe3O4@ALG/Fe
Without external energy
[Ciprofloxacin]0 ¼ 0.15 mM
Synthesized Feclaynanocomposites
[Tetracycline]0 ¼ 20 mg/L
Synthesized nano carbon/Fe3O4
Deionized water, pH 3.5 Temperature ¼ 40 C [Catalyst] ¼ 0.4 g/L [H2O2] ¼ 0.98 mM Reaction time ¼ 6 h Distilled water, pH 3 Temperature ¼ 25 C [catalyst] ¼ 1 g/L [H2O2] ¼ 60 mM Reaction time ¼ 30 min Deionized water, pH 3 Temperature ¼ 25 C [catalyst] ¼ 0.3 g/L [H2O2] ¼ 60 mM Reaction Time ¼ 5 h
[Sulfathiazole]0 ¼ 50 mg/L
Synthesized nano Fe3O4/HA
Deionized water, pH 3.5 Temperature ¼ 40 C [Catalyst] ¼ 3 g/L [H2O2] ¼ 0.39 M Reaction time ¼ 6 h
Without external energy
[Diclofenac]0 ¼ 5 mg/L
Pyrite (Ward’s natural science) or FeSO4$7H2O (Sigma eAldrich)
Deionized water pH 4 Temperature ¼ 25 C [Pyrite] ¼ 0.5 mM [H2O2] ¼ 80 mM Reaction time ¼ 200 s
Without external energy
HP Hg lamp 125 W 9 104 mE m2 s1
Without external energy
85.25% PhAC removal was achieved after 40 min. An excellent chemical stability with negligible leaching ions was reported for this catalyst. Both adsorption (early stage) and oxidation (final stage) were contributed in the DCF and TOC removal. More than 80% and 98.9% of DCF was removed after 2 and 120 min reaction, respectively. 66.4% TOC removal was achieved after 15 min. (BOD5/COD) was improved significantly from 0.043 to 0.63 in 120 min. Catalytic efficiency was increased to 2e4 folds by modifying Fe3O4. 100% of NOF and 90% of TOC is removed within 60 min. Total CFX degradation was achieved after 30 min. About 57% TOC removal also was reported. About 90% COD removal was achieved after 30 min. TC could not be degraded by H2O2 alone. 44.8% adsorption on PAC was achieved after 180 min. By Fenton-like process, 94% TC removal was obtained after 240 min. The pseudo-first order rate constant was 0.019 (min1) for TC degradation. Leachate iron from catalyst was less than allowed iron concentration in the effluent. The catalyst was shown high reusability and activity during the four sequential runs. Process could be conducted in a wide range of pH value in Fe3O4/HA-H2O2 system. Modification of catalyst with HA enhances the removal efficiency about 3.4 times. About total PhAC degradation was achieved after 1 h and 90% TOC removal was reported after 6 h. The pseudo-first order rate constant for PhAC and TOC removal were 0.034 and 0.0048 min1, respectively. 100% removal of DCF was observed in 120s with pyrite Fenton system. 65% of DCF was removed by classic Fenton system in 180s. The pseudo-first order rate constant was 0.164 (s1) for pyrite Fenton system.
Ref.
(Perisic et al., 2016)
(Niu et al., 2012)
(Bobu et al., 2008)
(Jaafarzadeh et al., 2015)
(Niu et al., 2011)
(Bae et al., 2013)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
[Diclofenac]0 ¼ 0.1 mM
Remarks and findings
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
677
Table 4 Selected articles on application of Fenton-type process in pilot and industrial scale. Considered contaminants
Experimental conditions
Remarks
Ref.
UV254, Fenton and ph-Fenton processes were used as the posttreatment stage for the removal of recalcitrant ECs from MWTP effluent. By using a conventional biological process, the degradation efficiency of CBZ, DCF, SMX, benzotriazole and mecoprop were 0, 34, 86, 0 and 0% respectively. During 33 s irradiation by UV254 (alone) an overall mineralization of 26% and 9.6% was achieved for all examined ECs. The natural dissolved iron content (1.6 mg/L) was good enough to remove ECs higher than 80%. The Fenton process limitations such as low pH range and high reagent concentration were overcome by using powerful UV lamps. Mineralization efficiencies of 0%e22% was achieved. [Chloramphenicol]0 ¼ 229 mg/L Reactor Type: CPC Primary laboratory-scale tests were performed to find the optimum Illuminated volume: 12 L [DOC]0 ¼ 111 mg/L condition for pilot-scale operations. Light source: Sunlight (41.5 W/m2) During the first 40 min, due to the formation of carboxylic acids, the [COD]0 ¼ 289 mg/L toxicity was increased, then it was decreased. [Fe2þ] ¼ 10 mg/L 100% CAP removal was achieved by Fenton process after 40 min. [H2O2] ¼ 400 mg/L After 60 min of reaction, 92% and 98% DOC and COD removal were pH 2.5e2.8 achieved, respectively. Temperature ¼ 45e55 C Solar ph-Fenton process was used as the post-treatment after the [NXA]0 ¼ 38 mg/L Pharmaceutical wastewater bioreactor operation. [Fe2þ] ¼ 20 mg/L [DOC]0 ¼ 725 mg/L No removal of NXA was shown by conventional biological treatment. [COD]0 ¼ 3400 mg/L [H2O2] ¼ 300 mg/L After 25 min illumination in solar ph-Fenton, NXA was totally Reactor Type: CPC degraded and DOC concentration reached 20 mg/L. Light source: Sunlight, The overall efficiency of the combined system for the removal of DOC Illuminated volume: 22 L was 97%. Total volume: 40 L Using biological processes before the ph-Fenton, significantly reduced Light source: Sunlight, the operation costs which is a major drawback of using AOPs alone. pH 2.6e2.8 No toxicity was shown based on the D. magna bioassay after treatment by ph-Fenton process. [AAP]0 ¼ 10 mg/L Primary experiments confirmed that no hydrolysis and photolysis of Distilled water & S.E. 2þ PhACs occurred without the catalyst. [ATL]0 ¼ 10 mg/L [Fe ] ¼ 5 mg/L Total degradation of AAP and ATL were reported after 12 and 3.8 min [H2O2] ¼ 10e20 mg/L of illumination by solar ph-Fenton in DW, respectively. Reactor Type: CPC Total degradation of AAP and ATL were reported after 21.8 and 30 min Light source: Sunlight of illumination by solar ph-Fenton in S.E., respectively. Illuminated volume: 36 L The ph-Fenton treatment was more efficient for the degradation of Total volume: 150 L ATL and ACTP than the photocatalytic treatment with TiO2. pH 2.6e2.8 Solar ph-Fenton, ozonation and solar/TiO2 were used as the postCombination of 66 ECs include MWTP 2þ 16 PhACs treatment after the secondary biological treatment process and the [Fe ] ¼ 5 mg/L [DOC]0 ¼ 13e23 mg/L removal efficiency showed the following trend: Solar ph[H2O2] ¼ 60 mg/L Fenton > ozonation > solar/TiO2. [COD]0 ¼ 32e63 mg/L Reactor Type: CPC Light source: Sunlight Over 98% ECs removal was achieved in only a few minutes of reaction Illuminated volume: 44.6 L by the ph-Fenton process. Total volume: 75 L ph-Fenton treatment produced an effluent with extremely low (non Temperature ¼ 35 C toxic) inhibition (<23% V. fischeri). Solar ph-Fenton is a potential advanced treatment in MWTPs with pH 2.8 treatment costs of <0.4 V/m3. Total removal of 4-MAA and 10% TOC removal were observed during [dipyrone]0 ¼ 50 mg/L Demineralised water the dark Fenton reaction within 15 min. [Fe2þ] ¼ 2 mg/L By applying the ph-Fenton process, 75% mineralization was rapidly [H2O2] ¼ 200e500 mg/L achieved in the first 30 min. Reactor Type: CPC By using solar/TiO2, 4-MAA was degraded at a considerably lower rate Light source: Sunlight Illuminated volume: 22 L than that observed during the ph-Fenton process. Total volume: 35 L V. fischeri assays demonstrated that in ph-Fenton treatment solutions, Temperature ¼ 30e40 C toxicity did not increase during the process. pH 2.8 A heterogeneous Fenton process was used to remove ECs from ECs include PhACs and MWTP biologically pre-treated municipal wastewater. hormones [H2O2] ¼ 200 mg/L [COD]0 ¼ 93.9 mg/L Reactor Type: continuous stir tank Performing the oxidation process at the natural pH of wastewater is a major advantage compared to the homogeneous Fenton process, reactor which is limited to pH in the range of 2e4. Total volume: 31.34 L >90% of the hormones and >40% of PhACs removal were reported. Temperature ¼ 12e20 C BOD was reduced to less than 1 mg/L after treatment. In addition, 30% Natural pH e40% TOC, 50% TSS, 68e91% turbidity and 90% phosphate reduction residence time ¼ 3 h were achieved. There was less than 4% iron leachate, suggesting a good stability of the catalyst. 22 ECs include 15 PhACs and 5 MWTP effluent pesticides [Fe2þ] ¼ 1.6 mg/L (natural) [H2O2] ¼ 20e50 mg/L Light source: 5 150 W UV254 lamps, Reactor volume (pilot): 37 L pH 6e7 (natural)
factor to achieve high removal efficiencies. In other words, at the optimum H2O2/iron ratio, the required iron ion and the H2O2 doses
(De la Cruz et al., 2013)
et al., 2013) (Trovo
(Sirtori et al., 2009b)
(Radjenovi c et al., 2009)
(Prieto-Rodríguez et al., 2013)
rez-Estrada et al., 2007) (Pe
(Chi et al., 2013)
can be reduced considerably (Tekin et al., 2006). The removal efficiencies of contaminants and the TOC, as well as the reaction
678
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
pathway, are affected by the ratio between various reactants ndez-Arriaga et al., 2010). In general, the optimal H2O2/iron (Me ratio depends upon the water matrix to be oxidized (Tekin et al., 2006). It is reported that the H2O2/Fe2þ ratio higher than 10 may lower the removal efficiency of contaminants due to the scavenging of hydroxyl radicals, according to Eq. (5) (Bautitz and Nogueira, 2007). Based on Eq. (1), it is clear that the formation of OH radicals in Fenton reaction also depends on the concentration of Fe2þ ions. The higher solution concentration of Fe2þ can increase the production of hydroxyl radicals, which produces a rapid removal of contaminants in the solution (Li et al., 2006; de Luna et al., 2013). However, further increases in iron ion dosage above the optimum concentration will not only lead to limited improvement in the degradation of contaminants, but will also result in the turbidity of solution due to the presence of excess iron (Badawy et al., 2006; Alalm et al., 2015). This turbidity prevents the penetration of UV light into the solution and consequently the absorption of light required for photolysis. In addition, a high concentration of Fe2þ ions may react with the produced OH radicals as a radical scavenger according to Eq. (8) (Tekin et al., 2006).
high concentrations of contaminants require higher concentrations of hydroxyl radicals for their effective removal. Therefore, the increase in the input concentration of contaminants, while maintaining a constant level of other operating parameters, will reduce the efficiency of degradation (Tekbas¸ et al., 2008; Li et al., 2015). In addition, the number of active sites, where the radical reactions take place, will become limited when the contaminant concentration is too high in the heterogeneous Fenton reaction. Moreover, the high concentration of contaminants will reduce the absorption of light by hydrogen peroxide in photocatalytic Fenton reaction, thereby hindering the OH radical generation (Michael et al., 2010; Li et al., 2015). At higher contaminant concentrations, the oxidation reaction requires higher reaction times and/or higher dosage of reactants for higher removal efficiency of contaminants. These phenomena indicate that in the Fenton processes, the optimum concentration of reagents should be determined based on the initial concentration of organic contaminants to ensure their effective removal.
Fe2þ þ OH /Fe3þ þ OH
Although, FeSO4.7H2O is commercially available for use in homogeneous Fenton and photo-Fenton processes, other types of ferrous catalysts can be used, but they will substantially affect the removal efficiencies, especially under solar light irradiation (Bautitz et al., 2011, 2012). For instance, during and Nogueira, 2007; Trovo the removal of bezafibrate and paracetamol, the use of FeOx in the photo-Fenton process as an alternative reagent to produce Fe2þ ions is highly effective, especially under solar radiation, because it strongly absorbs the solar light spectrum between 250 and 500 nm and has a high quantum efficiency of Fe2þ production (4Fe2þ ¼ 1:24 at 300 nm). Under irradiation, the Fe3þ-polycarboxylate complex will react according to the following reactions to produce Fe2þ ions et al., 2008): (Trovo
(8)
It has been shown that the photo-Fenton reaction requires a lower concentration of reagents to achieve the same degradation squez et al., 2014). This is rate compared to the Fenton reaction (Vela due to the generation of additional hydroxyl radicals and the reduction of Fe3þ ions according to Eq. (3) and Eq. (4), respectively. In heterogeneous Fenton processes, the reactions between iron ions and H2O2 at the catalyst surface similar to that in the homogenous type is expected to occur for the generation of OH radicals, according to Eqs. (1) and (2) (Aleksi c et al., 2010). Considering the similarities in heterogeneous and homogeneous Fenton-type processes, a similar impact of H2O2 concentration on the removal efficiency of contaminants is expected. Therefore, according to Eq. (5), the excess H2O2 concentration leads to the production of less reactive species such as HO2 due to the scavenging effect, causing a reduction in the removal efficiency. In addition, a higher concentration of solid catalyst will not only provide higher surface area and active sites, but it will also increase the concentration of dissolved iron ions in the solution (Bae et al., 2013). The excess amount of solid catalyst prevents efficient light absorption in the heterogeneous photocatalytic reaction by blocking the penetration of photons into the solution (Li et al., 2006; Kasiri et al., 2008). In the heterogeneous photocatalytic Fenton process, four main reactants, namely solid catalyst, H2O2, UV irradiation and oxygen are present. The catalyst and H2O2 are the most controlling parameters, since the absence of catalyst or H2O2 will result in negligible rates of reaction. The presence of UV irradiation and air (oxygen) are important since they will increase the removal rates (Tekbas¸ et al., 2008). 3.2. Effect of initial contaminants concentration The concentration of contaminants in the effluent of wastewater treatment plants varies on a daily basis. Therefore, it is essential to assess the dependence of removal efficiency on the input concentration of pollutants (Titouhi and Belgaied, 2016). Considering very short lifetime of hydroxyl radicals, the increase in the concentration of contaminants will raise the possibility of collision between OH species and organic matter, leading to an improvement in the rate of removal (Kasiri et al., 2008). On the other hand, increasing the input concentration of contaminants has been shown to reduce their removal efficiency (Michael et al., 2010). This occurs because
3.3. Effect of catalyst type
2þ ½FeðC2 O4 Þ3 þ C2 O þ 3C2 O2 4 /Fe 4 þ 2CO2
(9)
3 þ hv/Fe2þ þ 2C2 O2 FeðC2 O4Þ3 4 þ C 2 O4
(10)
2 C2 O 4 þ O2 /O2 þ 3C2 O4 þ 2CO2
(11)
Using FeOx instead of Fe(NO3)3 in the photo-Fenton process under solar irradiation was very effective for the removal of organic contaminants (Nogueira et al., 2005). The high absorption of ferrioxalate makes a large portion of the solar spectrum available for sito et al., 2016). However, the generation of hydroxyl radicals (Expo in spite of the existing beneficial effects, using FeOx as the source of iron in photo-Fenton process increases the carbon load in the solution. Oxalate can be completely converted to CO2 under optimal conditions. While using Fe(NO3)3 as the source of iron was favored for tetracycline degradation under black-light irradiation (Bautitz and Nogueira, 2007), the use of FeOx was preferred for the degradation of 4-chlorophenol under solar irradiation (Nogueira et al., 2005). Therefore, the iron source should be chosen based on the chemical characteristic of the target contaminants and the type of irradiation source (Nogueira et al., 2005; Bautitz and Nogueira, 2007). Similarly, the type and structure of the catalyst can considerably affect the efficiency of heterogeneous Fenton process. A higher TOC removal efficiency of pharmaceuticals by nano-magnetite compared to submicron-structured magnetite was reported (Velichkova et al., 2013). The nanostructure catalysts provide more surface area and a higher number of active sites which decompose hydrogen peroxide, while producing a higher concentration of
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
leached iron (Sun and Lemley, 2011). The generation of a higher concentration of leached iron ions may promote homogeneous Fenton reaction, which leads to a higher removal efficiency of contaminants. On the other hand, it can increase the generation of sludge due to the precipitation of iron. The type of catalyst may also affect the efficiency of removal. For instance, maghemite powder (g-Fe2O3) has been shown to produce higher reaction kinetics for the decomposition of H2O2 to hydroxyl radicals compared to magnetite (Fe3O4) (Velichkova et al., 2013). In addition, due to the higher electron mobility in the spinel Fe3O4, the efficiency of hydroxyl radicals generation is higher than a-Fe2O3 (Rodríguez et al., 2011). It should be kept in mind that the higher activity could be mainly attributed to the higher iron leaching. In contrast to magnetite, in hematite and goethite, iron is initially present only as Fe3þ. Therefore, in the presence of hematite and goethite, the Fe2þ should be produced by irradiation based on Eq. (3) and/or Eq. (4) (Rodríguez et al., 2011). Pyrite not only introduces Fe2þ into the solution, it also reduces the pH of solution based on Eqs. (12) and (13). These two functions are favorable for the homogeneous Fenton process (Bae et al., 2013). þ 2FeS2 þ 7O2 þ 2H2 O/2Fe2þ þ 4SO2 4 þ 4H
(12)
þ 2FeS2 þ 15H2 O2 /2Fe3þ þ 4SO2 4 þ 2H þ 14H2 O
(13)
The differences between zero point charges (zpc) of different iron catalysts can change the mechanism of reaction at various pH values. Considering that the iron catalyst can act as a semiconductor and can produce pairs of electron/holes (e/hþ) under sufficient light energy, the different band gap in various types of iron catalysts can affect the efficiency of reaction. The superior photodegradation of paracetamol was reported in a combination of goethite-H2O2-UV system due to the synergy between the two processes (Mameri et al., 2016). Under UV irradiation, hematite and goethite with 2.2 eV band gap can act as a photocatalyst. The generated electron/hole pairs on the goethite surface can react with contaminants or dissolved oxygen on the goethite active sites and produce strong oxidative radical species. On the other hand, goethite can react with H2O2 according to the photo-Fenton reaction to produce more OH. Researchers have demonstrated the synergism between iron oxides and TiO2 catalyst in the photo-Fenton system (Rodríguez et al., 2009, 2011). This is because the presence of iron oxide may reduce the energy gap of TiO2 photocatalyst, facilitating the electron transfer from the valence to the conduction band of TiO2, which leads to the generation of more hydroxyl radicals. Moreover, a combination of hematite and TiO2 not only reduces the recombination rate of generated electron/hole pairs on TiO2 surface, but also increases the formation of Fe2þ from Fe3þ by capturing the produced electrons on the conduction band of titanium oxide. Besides, the reduction of Fe3þ by capturing the produced electrons on the conduction band of titania may increase iron leaching into the solution. The leached iron should later be removed from the effluent in order to meet the standard for iron, which will increase the operating costs of the process. However, this synergy was not reported in the case of magnetite by Rodríguez et al. (2009). The iron content of catalyst has a significant influence on the degradation of pharmaceuticals. The increase of Fe loading rate beyond the optimum value may have a negative effect on the efficiency of process because the Fe dispersion will become less homogeneous (Bautista et al., 2010b; Titouhi and Belgaied, 2016). In general, the higher iron content of the catalyst offers a better distribution, while increasing the iron concentration in solution (leachate). Considering these two effects, a trade-off should be made in order to find the optimum iron content of the catalyst (Lan et al., 2015).
679
3.4. Effect of pH The pH of solution has a decisive influence on the performance of homogeneous Fenton reaction (Kwon et al., 1999; Shemer et al., 2006; Batista and Nogueira, 2012). The pH of around 3 was mentioned as the optimal value for the Fenton reaction in many studies, while values over 5 have been shown to cause the formation of ferrous hydroxide according to Eq. (14) (Kwon et al., 1999; Gogate and Pandit, 2004; Escher et al., 2006; Hartmann et al., 2010; Vel asquez et al., 2014). Due to the poor solubility of ferrous hydroxide in water, its precipitation will not only consume the ferrous ions and inhibit the Fenton reaction, but it will also reduce the transmission of radiation into the water by increasing the turbidity of solution (Alalm et al., 2015).
Fe2þ þ 2OH /FeðOHÞ2 Y
(14)
The oxidation potential considerably changes at different pH values (E0 ¼ 2.8 V at pH 0 and E0 ¼ 1.59 V at pH 14). Another reason for the decline in the efficiency of photo-degradation at pH > 4 is attributed to the low rate of H2O2 decomposition, which leads to a decrease in the production of OH (Badawy et al., 2006). At pH > 4 hydrogen peroxide is preferentially decomposed into H2O and O2 (Hartmann et al., 2010; Titouhi and Belgaied, 2016). According to Eq. (1), at high pH values, due to the presence of OH species, the reaction shifts to the left and reduces the formation of free radicals. In basic media, hydrogen peroxide is decomposed to oxygen without generating hydroxyl radicals, according to Eq. (15) and Eq. (16) (Zhang et al., 2005).
H2 O2 þ OH 4HO 2 þ H2 O
(15)
H2 O2 þ HO 2 4O2 þ H2 O þ OH
(16)
It has been concluded that at the pH range of 2e4, more Fe(OH)þ complex is formed, which has a higher activity in comparison with Feþ2 in Fenton and photo-Fenton reactions. Alalm et al. (2015) reported that while the complete removal of amoxicillin was obtained at pH 3 after 90 min of irradiation, 80% removal was achieved at neutral condition. The same trend was reported in the case of ampicillin, diclofenac and paracetamol. On the other hand, at a strong acidic condition, the rate of formation of Fe2þ from Fe3þ reduces according to Eq. (4). This happens because at strongly acidic media, where the Hþ species are abundant, the reaction will be shifted towards the left and will reduce the formation of OH radicals (Li et al., 2009). pH values lower than 2.5 increase the formation of [FeOH]2þ, [Fe2þ (H2O)]2þ and [H3O2]þ complex, which react slowly with hydrogen peroxide and generate lower amounts of hydroxyl radicals, thus decreasing the removal efficiencies (Gogate and Pandit, 2004; Shemer et al., 2006; Tekin et al., 2006; Titouhi and Belgaied, 2016). Considering Eq. (17), at very acidic condition (pH < 3) the abundant hydrogen ions (Hþ) can act as a radical scavenger and will reduce the efficiency of reaction. In this reaction, the electron can be gained from the ferrous ion based on Eq. (17) (Tang and Huang, 1996; Shemer et al., 2006; Michael et al., 2010).
OH þ Hþ þ e /H2 O
(17)
At strongly acidic media (pH about 1), H2O2 molecules cannot be decomposed to produce OH radicals. This probably occurs because under this condition, H2O2 will solvate a Hþ to form an oxonium ion ðH3 Oþ 2 Þ (Domínguez et al., 2012). Due to the electrophilic behavior of oxonium ion, the reactivity with ferrous ions is reduced (Kwon et al., 1999). It should be noted that ferric iron precipitates at the pH > 4 and can act as a photocatalyst to generate electronehole
680
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
pairs via excitation by UV irradiation. Therefore, further possible reactions are the oxidation of adsorbed contaminants on the surface of the precipitated ferrous hydroxide and production of addirez-Estrada et al., 2005b). The reactions tional oxidative radicals (Pe in Eqs. (18e26) are possible in the latter case:
hþ þ organic matterðOMÞ/OM /oxidized products vb
(18)
e cb þ O2 ðadsÞ/O2
(19)
O 2 þ H2 O2 /OH þ OH þ O2
(20)
e CB þ H2 O2 /OH þ OH
(21)
þ H2 OðadsÞ/Hþ þ OH hþ vb
(22)
hþ þ HO ðadsÞ/OH vb
(23)
OH þ H2 O2 /H2 O þ HO2
(24)
HO2 þ OH /H2 O þ O2
(25)
H2 O2 þ hþ /HO2 þ Hþ
(26)
Additionally, the pH value can affect the reaction pathways by changing the structure and characteristics of contaminants such as their solubility. The pollutant is in molecular form when the pH is less than its pKa, whereas at pH > pKa the compound loses a proton and becomes negatively charged. For example, diclofenac is insoluble at low pH values (around 3), which is a favorable condition for Fenton reaction. Therefore, due to the incompatibility between diclofenac insolubility and the photo-Fenton process performance rez-Estrada at low pH, the optimization of pH value is essential (Pe et al., 2005b). It is possible that the buffer capacity of water will alter the reaction pathway and will partly overcome the iron and contaminant precipitation. For instance, diclofenac degradation in water without buffering capacity was favorable because starting the reaction at neutral pH, where diclofenac is very soluble, led to its high degradation. During the treatment, pH changes from 7 to about 4 due to the production of intermediate carboxylic acids. As a result, iron precipitation will slow down, yielding an overall rez-Estrada et al., 2005b). accelerated degradation (Pe Recently, Romero et al. (2016a) proposed a method to resolve the problems related to the formation of ferrous hydroxide at pH value higher than 4. They used resorcinol as a chemical additive in order to form Fe-RES (Ligand-Fe) which can surround the iron and hinder the precipitation reaction at neutral pH values. At the end of reaction, the target contaminant and the chemical additive were both degraded. By using this method, the degraded concentration of the target contaminant and the TOC removal efficiency were improved considerably compared to the conventional photoFenton process (pH 2.8) (Romero et al., 2016a). However, this method requires more investigation to determine the involved reactions and the implicated chemical pathway. Although the homogeneous Fenton process is efficient at the acidic pH of around 3, heterogeneous type processes can be effectively operated at milder pH conditions. A substantial dependence of the heterogeneous Fenton reaction on the solution’s pH was reported (Feng et al., 2006). Similar to the results obtained in the homogeneous Fenton process, the removal efficiency increases in the heterogeneous Fenton process by lowering the pH value. This is because of the enhanced leaching of iron ions from the solid catalyst at lower pH values which imposes additional limitations on the
process (Tekbas¸ et al., 2008). Under strongly acidic conditions (pH lower than 2), the removal efficiency declines due to the formation of H3 Oþ 2 ions, thus decreasing its reactivity with ferrous ion (Aleksic et al., 2010). It should be stressed that in the heterogeneous Fenton processes, the leaching of iron ions into the solution can be associated with the removal of contaminants, which will increase the removal efficiency of the process. Thus, increasing the pH of solution can restrain the homogeneous Fenton reaction through the formation of stable ferrous hydroxide, according to Eq. (14). Using the ZSM5 zeolite as the support medium in the heterogeneous Fenton processes can extend the pH range of the homogeneous Fenton reaction. Such effect could be related to the specific characteristics of ZSM5 zeolite, where iron cations are under the influence of a strong electrostatic field in the zeolite framework. The distribution of Fe cations over the zeolite structure with appropriate distribution of the negative charge can prevent or postpone the formation of iron hydroxides as pH approaches neutral conditions (Kasiri et al., 2008; Aleksi c et al., 2010). Moreover, Aleksi c et al. (2010) reported that the volume of leached iron from the zeolite support strongly decreases with the increase of pH value. This happens because at the strongly acidic conditions, the electrostatic force is weaker, thus allowing the leaching of iron ions. The same trend was reported for pillared clays used as the support medium in heterogeneous Fenton reactions (Feng et al., 2006). However, incorporating aluminum besides iron in the clay structure could markedly reduce the iron leachate even at low pH values due to the presence of a strong interaction between aluminum and iron (Li et al., 2006). 3.5. Effect of intensity of radiation The direct photolysis of contaminants occurs when the chemicals absorb light with enough energy (Elmolla and Chaudhuri, 2009). The direct photo-oxidation reaction is based on electronic excitation of the organic compounds, denoted in most cases an electron transfer from the excited-state C* (Eq. (27)) to molecular oxygen (Eq. (28)) or hemolysis (Eq. (29)) to produce radicals which will then react with oxygen (Eq. (30)) (Legrini et al., 1993). hy
C!C *
(27)
C * þ O2 /C $þ þ O$ 2
(28)
hy
R X!R$ þ X $
(29)
R$ þ O2 /RO2
(30)
Therefore, the efficiency of direct photolysis depends on the photon rate at the wavelength of excitation, the absorption cross sectional area of medium, the quantum yield of the process, and the concentration of dissolved molecular oxygen (Legrini et al., 1993). In most cases, due to the low absorption of chemical contaminants, the direct photolysis is insignificant or may lead to a high reaction time. For instance, no degradation of paracetamol was shown after 3 h of irradiation by 365 nm (Mameri et al., 2016). This is because paracetamol has a weak absorption in the range from 300 to 800 nm. The degradation efficiencies of amoxicillin, ampicillin and cloxacillin were reported to be 2.9%, 3.8% and 4.9%, respectively, after 5 h of irradiation under 365 nm wavelength lamp (Elmolla and Chaudhuri, 2009). The maximum absorption of metronidazole occurs at 310 nm. Therefore, applying medium-pressure UV lamp (200e400 nm) which has a substantial overlap with metronidazole absorption band, increased the removal rate constant by five-fold compared to using a low-pressure lamp (254 nm) (Shemer et al.,
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
2006). Ravina et al. (2002) used various types of UV lamps with different intensities and reported direct proportionality between the mineralization of diclofenac and the intensity of radiation. It is proven that the maximum absorbance of H2O2 occurs at 210e230 nm and the proteolysis is low at a wavelength higher than 300 nm (Pignatello et al., 1999; Elmolla and Chaudhuri, 2009). For instance, at UV-A region (365 nm), the production of hydrogen radicals by direct photolysis was very low due to the low molar absorptivity (~0.01 M-1 cm-1) (Rodríguez et al., 2011). At lower wavelengths, the direct proteolysis of chemical contaminants and H2O2 to hydroxyl radicals takes place simultaneously. In the Fenton reactions, free radicals can be produced in the absence of light energy by the reaction between ferrous salt and H2O2, according to Eq. (1). In this case, the production of hydroxyl radicals is limited since one molecule of Fe2þ produces only one OH radical. The illumination of Fenton reaction (photo-Fenton) is a proven method to overcome the limitation in the production of sufficient hydroxyl radicals (Eq. (3) and Eq. (4)). Considering Eq. (3), the regenerated Fe2þ is subsequently re-oxidized by H2O2 and produces new OH radicals. Consequently, the oxidation of organic contaminants is accelerated (Badawy et al., 2006). Therefore, the contaminants can be degraded more efficiently using fewer reagents or in less reacsquez et al., 2014). H2O2 can be decomposed to OH tion time (Vela radicals in the presence of UV light according to Eq. (4), while UV irradiation may excite contaminates to act as intermediate initiators to produce OH radicals in a chain reaction. At the pH between 2 and 3, [Fe(OH)(H2O)5]2þ and [Fe(OH)]2þ are predominant. These species are photosensitive to radiation in the range of 280e405 nm, making available the use of sunlight based on Eq. (31) (Vel asquez et al., 2014).
FeðOHÞðH2 OÞ5
2þ
2þ þ H2 O þ hy/ FeðH2 OÞ6 þ OH$
(31)
While the presence of solid catalyst increases the turbidity of solution and reduces the solution transparency and radiation flux, the heterogeneous photo-Fenton process may be more effective compared to the homogeneous photo-Fenton processes at quasi-neutral pH values. Higher removal efficiencies, higher mineralization and milder operating pH values have been reported as the advantages of photo- over dark-heterogeneous Fenton processes (Aleksi c et al., 2010). However, a thorough cost analysis should be performed in order to evaluate the viability of the overall process for the treatment of emerging contaminants in a water matrix.
3.6. Effect of water matrix The degradation of pharmaceuticals is dependent on the type of water matrix such as pure water, surface water and sewage treatment plant effluent. Therefore, in the case of pharmaceuticals which appear in the effluent of WWTPs, it is important to assess the et al., 2008). effect of this matrix on the removal efficiency (Trovo The WWTPs effluent usually contains chlorides, organic and inorganic carbon (carbonate and bicarbonate). The presence of these species can strongly reduce the removal efficiencies of pharmaceuticals by the scavenging hydroxyl radicals. Humic substances may act as an inner filter and severely decrease removal efficiencies (Legrini et al., 1993). The organic matters present in the real STPs can interfere in different manners and compete with the target components for degradation or by radiation attenuation in photoFenton reactions. The high carbonate content of solution can reduce the efficiency of removal due to its scavenging effect, according to Eqs. (32) and (33) (Bautitz and Nogueira, 2007).
681
$ $ HCO 3 þ OH /H2 O þ CO3
(32)
$ CO2 3 þ OH /OH þ CO3
(33)
3.7. Effect of the salinity of substrate A high concentration of inorganic salts, especially NaCl, has been detected in industrial wastewaters, such as those producing pharmaceuticals, pesticides, dyes and herbicides (Bacardit et al., 2007). 2 The presence of inorganic ions such as Cl ; SO2 4 ; HPO4 ; HPO4 may reduce the efficiency of Fenton processes. This may occur since anions react with hydroxyl radicals and generate anion radicals that are less reactive compared to hydroxyl radicals (Eqs. (34) and (35)) (Sirtori et al., 2009b). The interaction of chloride with Fe2þ/Fe3þ ions may produce complex chemical substances according to Eqs. (36)e(39) (Bacardit et al., 2007).
Cl þ OH$ 4ClOH$ 4Cl$ þ OH
(34)
ClOH$ þ Hþ 4ClOH2$ 4Cl$ þ H2 O
(35)
Fe2þ þ Cl /FeClþ
(36)
FeClþ þ Cl /FeCl02
(37)
Fe3þ þ Cl 4FeCl2þ
(38)
Fe3þ þ 2Cl /FeClþ 2
(39)
The chloride ions can react with Fe3þ under irradiation which may further produce organic radicals (R) or chloride radicals Cl, which are less reactive in comparison with the hydroxyl radical (Eqs. (40) and (41)). * e3þ
F
*
hy
ðaqÞCl !Fe2þ ðaqÞ þ Cl$
FeCl3 þ RH/FeCl2 þ HCl þ R$
Cl$ þ Cl /Cl$ 2
(40) (41) (42)
Where *Fe indicates the photoexcited species. These photoexcited species can also participate in the reaction but at a lower rate. In the distilled water, 76% removal of sulfamethoxazole by photolysis was reported after 7 h, while under the same condition, et al., 2009). only 14% removal was shown by using seawater (Trovo The inhabitation of Fenton and photo-Fenton reactions by a high concentration of sodium chloride strongly depends on the pH of solution. At the pH of equal or less than 2, the photolysis of Fe3þ complexes in the presence of Cl ions leads to the formation of the Cl$ 2 radical anion which is less reactive than OH radicals (Eq. (38), (40) and (42)). In addition, under these conditions, the scavenging of hydroxyl radical by the chloride ion (Eq. (34) and (35)) becomes important (Machulek et al., 2007). In the photo-Fenton process, at the initial stage of reaction where the pH of solution is adjusted to 3e3.5 (optimum pH for Fe(OH)2þ formation), the removal efficiency is only slightly affected by the presence of chloride ions, even at a high concentration (0.5 M). However, since the pH of solution is usually not controlled during the oxidation process, the formation of organic acids will cause a progressive acidification and drop in the pH value. In this condition, chloride ion begins to
682
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
scavenge any OH radicals formed in the system and will form Cl$ 2 species as a product. Therefore, there is an obvious influence of chloride ion on the photocatalytic oxidation process (Machulek et al., 2006). The inhibition of reaction can be circumvented by maintaining the solution pH at approximately 3 throughout the course of reaction, even in the presence of high chloride ion concentrations (0.5 M) (Machulek et al., 2006, 2007). Bacardit et al. (2007) demonstrated that the presence of high concentrations of NaCl reduced the rate of removal reaction in the photo-Fenton process. This is due to the fact that sodium chloride prevents or reduces the generation of hydroxyl radicals due to the chain reactions of chloride and iron ions. These investigators also reported that the presence of phosphate ðPO3 4 Þ completely inhibits the Fenton process and severely affects TOC removal. This should be expected since the addition of iron ions leads to the formation of iron-phosphate complexes, commonly used for phosphate precipitation and its removal in wastewater treatment plants. Finally, the introduction of acid or base for the adjustment of solution pH can affect the removal efficiency of contaminants. For instance, the acidification of solution by HCl introduces a high amount of conjugated anions (Cl) into the solution. These anions can reduce the efficiency of removal due to the drastic competition between the anions and contaminants for the hydroxyl radicals (Eq. (34)) (Kasiri et al., 2008). 3.8. Effect of feeding mode Several researchers have investigated the influence of the feeding mode of Fenton reagent to the reactor on the efficiency of process (Bowers et al., 1989; Turan-Ertas and Gurol, 2002; Zhang et al., 2005). Due to the competition between the species that participate in the reaction such as iron ions, radicals, hydrogen peroxide, target contaminants and produced intermediates, changing of the feeding mode may alter the reaction mechanism and the removal efficiency of pharmaceuticals. When H2O2 was added in two steps, the efficiency of COD removal was improved, as reported by Zhang et al. (2005). This effect might be due to the radical scavenging effect of H2O2 when introducing a large initial amount of hydrogen peroxide according to Eq. (5). Therefore, successive addition of the H2O2 will keep the hydrogen peroxide concentration at relatively low levels, decreasing the detrimental effect of hydroxyl radical scavenging (Turan-Ertas and Gurol, 2002; Zhang et al., 2005). In the solar reactor, the recirculation rate may affect the process efficiency. A faster recirculation of solution in the photo-reactor leads to the improvement of mixing inside the reactor. However, a very high recirculation rate can reduce the reactor’s residence time. The reduction of contact time between the contaminants and Fenton reagents will lead to the generation of fewer radicals and a reduced reaction rate (Ravina et al., 2002). 3.9. Effect of temperature Temperature is another important parameter in the Fenton processes. Theoretically, the increase of temperature should enhance the reaction kinetics (Velichkova et al., 2013). In the Fenton reaction, the increase of temperature accelerates the reaction rate between ferrous ions and H2O2, which causes the generation of additional hydroxyl radicals. The increase of reaction temperature can enhance the decomposition rate of H2O2 by a factor of 2 at each 10 C temperature increase (Hartmann et al., 2010). However, conducting the reaction at a temperature higher than 40 C may decrease the removal rate. It is generally accepted that at high temperatures, the decomposition of H2O2 to water and oxygen is possible (Guo et al., 2015). Zazo et al. (2011) reported that a
significant improvement in the mineralization of phenol was observed up to 90 C. They concluded that the increase of temperature could enhance hydrogen peroxide transformation into hydroxyl radicals. Therefore, a considerable improvement of the oxidation rate and the mineralization efficiency can result, leading to a reduction in the reagent dosage at high temperatures. The decrease of Fe2þ and hydrogen peroxide dose reduces sludge generation and improves the efficiency of process by minimizing scavenging reactions. Besides, it is expected that COD removal can be enhanced by the increase of temperature (Zhang et al., 2005). Nevertheless, Tekin et al. (2006) reported that no significant differences were observed in the outcome of reactions conducted at room temperature or at 40 C. It is worth mentioning that, although the increase of temperature imposes a large economical burden to the process in real STPs, it does not represent a major drawback for the treatment of many industrial effluents, which usually have a high temperature. Heat recovery from the treated stream allows saving energy (Zazo et al., 2011). Therefore, various effects of operation temperature must be taken into consideration when defining the optimal operating temperature for organic matter removal (Bautista et al., 2010b). Similar trends are expected in the heterogeneous Fenton process. The increase of operating temperature improves the efficiency of organic matter removal, as the higher temperatures allow the reagents to overcome the activation energy barrier based on the Arrhenius law, thus facilitating the oxidation reactions. However, high temperatures may introduce certain inconveniences such as increasing the possible presence of iron leachate in the solution (Rodríguez et al., 2009). For instance, increasing the temperature from 50 to 70 C reduced the reaction time by half during the removal of phenol, while at the same time increasing the loss of iron from the catalyst by three times compared to that at 50 C (Zazo et al., 2006). In addition, high temperature can cause the deterioration of support medium which will cause the loss of catalyst smoothness and its decomposition into fine particles (Titouhi and Belgaied, 2016). The reaction pathway can also change with temperature. For instance, the production of oxalic acid as a by-product of phenol oxidation was increased by increasing the temperature (Zazo et al., 2006). As mentioned before, high temperature may also enhance the thermal degradation rate of hydrogen peroxide. Therefore, the oxidation rate of organic contaminants should be a function of competition between thermal degradation of H2O2 and hydroxyl radical formation (Liou et al., 2005). 3.10. Effect of reaction time Because of the toxicity of intermediate by-products, the moment of reaction termination should be carefully evaluated if complete mineralization is not the goal. This implies that in certain cases, while the complete removal of target contaminant has been achieved after a short period of reaction, a major fraction of the contaminant has actually been transformed into the intermediates. Consequently, more reaction time is needed to ensure the mineralization of contaminants and their reliable removal (Li et al., 2015). For instance, the degradation of sulfathiazole led to the formation of maleic acid whose concentration reached a peak concentration squez et al., after 2 min and totally disappeared after 8 min (Vela 2014). Therefore, the evaluation of reaction time for the total mineralization of target contaminants and by-products is recommended. As stated before, the degradation rate of pharmaceutical depends on the presence and abundance of hydroxyl radicals. At the early stages of reaction, usually high degradation rates of contaminants are achieved due to the abundance of produced hydroxyl radicals. With the progress of reaction, the degradation rate
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
decreases due to the consumption of hydroxyl radicals and reduction in the concentration of contaminants (Alalm et al., 2015).
3.11. Deactivation of catalyst In the heterogeneous Fenton reaction, the performance of catalyst may decrease after consecutive runs. The main reasons for the deactivation of catalyst are the leaching of iron from the catalyst structure and/or the decay of active catalytic sites (Tekbas¸ et al., 2008). The leaching of iron which mainly occur at the low pH values will cause the reaction to shift to the homogeneous Fenton reaction. In this case, the removal rate usually increases. However, the capacity of catalyst in the subsequent stages of the process will substantially decrease. Moreover, the leaching of iron from a catalyst results in the formation of a secondary pollution (Nidheesh et al., 2013). The leachate of iron species from iron oxides depends on the structure of the catalyst. The corundum structure (aFe2O3) is more stable than Fe3O4 with spinel structure. Thus, the leachate from (a-Fe2O3) is considerably lower than Fe3O4 (Rodríguez et al., 2009; Nidheesh et al., 2013). The decrease of pH value increases the iron leaching, while the leachate from iron oxides is negligible at neutral pH (Rodríguez et al., 2009; Perisic et al., 2016). As mentioned before, the increase of temperature causes higher leaching of iron and consequently produces the deactivation of catalyst in successive cycles (Ramirez et al., 2007; Rodríguez et al., 2009). In the latter case, consecutive runs may cause surface blockage, which changes the porous structure and limits the accessibility of target compounds to the pores after reuse. It is possible that the adsorption of produced contaminants and residual organic matter at the surface of the catalyst, will lead to the formation of deposits on the surface (Nidheesh et al., 2013; Titouhi and Belgaied, 2016). Some modifications to the catalyst structure can improve the stability of catalyst by diminishing iron leachate during the oxidation. For instance, an excellent stability of heterogeneous Fenton catalyst using aluminum as mixing metal with iron in pillared clay was demonstrated (Li et al., 2006, 2015).
3.12. By-products formation Considering the non-selective nature of hydroxyl radicals, various by-products may be formed at low concentrations in radical-induced reactions (Radjenovi c et al., 2009). Therefore, during the treatment of effluents of wastewater treatment plants, complete mineralization is more desirable than the elimination of specific target contaminants. This is important since a complete removal of TOC will ensure complete mineralization of contaminants and potential intermediate compounds into CO2 and H2O (Feng et al., 2006). It should be noted that the produced intermediates can play incompatible roles as they can form complexes with iron ions and change the efficiency of degradation. For instance, organic intermediates accelerated the degradation of pchlorophenol by the reduction of ferric ion to ferrous ion (Kwon et al., 1999). Conversely, some intermediates can act as ferrous scavenger and prevent or delay the Fe2þ/Fe3þ regeneration cycle by forming stable complexes. The formation of a stable complex of organic intermediates with ferrous and ferric ions has been reported (Kwon et al., 1999). In some cases, the production of iron leachate and the deactivation of catalyst are related to the production of intermediate compounds. Some intermediate compounds such as oxalic acid can capture iron ions from the catalyst and form iron complexes. When these intermediates are mineralized into H2O and CO2, the iron ions return to the surface of catalyst.
683
4. Toxicity and biodegradability tests The degradation products and/or metabolites of emerging contaminants should be taken into consideration since the byproducts produced during the degradation processes are sometimes more toxic than the parent compounds (Ballesteros Martín et al., 2008; Oller et al., 2011; Prieto-Rodríguez et al., 2013; Veloutsou et al., 2014). A very high toxicity was reported in the case of isoproturon degradation by the photo-Fenton process, especially during the early stages of reaction (Parra et al., 2000). However, toxicity assessments demonstrated that photo-Fenton processes showed better performance in comparison with TiO2 rez-Estrada photocatalysis for the removal of pharmaceuticals (Pe et al., 2007). The generated intermediates may be more toxic to the microorganisms than the original contaminants (Oller et al., 2011). For that reason, a toxicity assessment is crucial when AOPs are used as a pretreatment step in hybrid systems (PrietoRodríguez et al., 2013). Toxicity assays may be applied as a major criterion for selecting the type of pretreatment process upstream of biotreatment operations (Lapertot et al., 2006). It is also important to mention that most AOPs reduce the pH of solution due to the formation of inorganic acids. In Fenton processes, the solution pH should be around 3. Hence, prior to conducting toxicity tests, the solution should be neutralized and chemical reagents such as hydrogen peroxide, catalysts, etc. must be removed from the media. Various tests such as D. magna, Selenastrum capricornutum, V. fischeri, Pseudomonas, Staphylococcus aureus, Escherichia coli, Phaeodactylum tricornutum, Pseudokirchneriella subcapitata and Lepidium sativum have been used for toxicity assessment (Oller et al., 2011). By using D. magna toxicity test, Arslan-Alaton and Gurses (2004) concluded that the photo-Fenton-like process is an effective method for the complete detoxification and partial oxidation of PPG production effluent. In another research, the high efficiency of Fenton process as a pretreatment stage for the elimination of acute toxicity and increased biodegradability of an effluent was shown by conducting activated sludge inhibition test (ISO 8192) (Alaton and Teksoy, 2007). The improvement of biodegradability and decrease of ecotoxicity to the marine bacteria V. fischeri by solar photo-Fenton treatment of ECs were reported (Zapata et al., 2009). It is worth mentioning that the water matrix can substantially alter the generated intermediates and consequently the toxicity of the treated water. For instance, D. magna bioassays showed that the degradation of SMX when performed in distilled water led to the decrease of toxicity from 85% to 20%. However, the toxicity increased from 16% to 86% when the reaction et al., 2009). During the removal was performed in seawater (Trovo of paracetamol by using FeSO4, a significant reduction of the sample toxicity was reported compared to using FeOx as the iron source et al., 2012). This is because the intermediate compounds (Trovo generated in distilled water are different from those generated in seawater. However, by conducting the V. Fischeri assay, Zapata et al. (2008) concluded that although the photo-Fenton process was very effective in the degradation of pharmaceuticals, the acute toxicity of effluents persisted until the end of treatment, while biodegradability was only slightly improved. Although toxicity tests provide valuable information regarding the effect of Fenton reactions on subsequent biological processes, further biodegradability tests (BOD5/COD) should be conducted to ensure the effectiveness of biological processes. This is important since bacteriaor fungi, which are commonly used in conventional biological water treatment processes, are sensitive to changes in the environmental parameters such as pH, temperature, dissolved oxygenconcentration, salinity, redox potential, and the presence of toxic compounds (Oller et al., 2011). Biodegradability is not directly related to toxicity, but it could be inferred that if toxicity enhances
684
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
during the treatment process, the biodegradability of solution should decrease (Lapertot et al., 2006). The enhancement in biodegradability of pharmaceuticals has been reported in various studies following Fenton-based reactions (Tekin et al., 2006; Badawy et al., 2009; Elmolla and Chaudhuri, 2009; Giri and Golder, 2014). For instance, biodegradability (BOD5/COD) was improved significantly from 0.043 to 0.63 after 120 min of exposure to Fenton reactions, in the case of DCF removal (Perisic et al., 2016). At the optimum condition, about 2.5 fold enhancement of (BOD5/ COD) ratio was reported after the treatment of a secondary effluent ~ ones et al., 2015). of a municipal wastewater treatment plant (Quin García-Ripoll et al. (2009) used Pseudomonas putida culture as a convenient, reproducible and reliable test for evaluating both biodegradability and toxicity, as an alternative to other timeconsuming tests for ECs solution partially treated by photoFenton process. 5. Large-scale application Extensive studies assessing Fenton-type processes for the removal of a wide range of pharmaceuticals from drinking water were reviewed and discussed in this manuscript. Most of the reviewed papers are based on bench-scale operations as there are only few reports on pilot-scale or industrial applications of Fenton technologies. Small-scale investigations often used a model system prepared by dissolving a single EC in distilled water and evaluated parameters such as COD, BOD5 and DOC in laboratory-scale operations (Zapata et al., 2010). Despite the high efficiency of Fenton processes in the removal of toxic and recalcitrant contaminants such as ECs, the high operating costs limit their commercial or industrial use. One attractive cost-cutting approach is the exploitation of renewable energy resources such as sunlight in photoFenton processes (Sirtori et al., 2009b, 2011; Zapata et al., 2010; et al., 2013). Another promPrieto-Rodríguez et al., 2013; Trovo ising solution is the integration of Fenton processes with the conventional biological treatment processes (Sirtori et al., 2009b, 2011). From this point of view, Fenton reactions are mostly utilized as a pre-treatment process in order to improve the biodegradability of contaminants. Compared to biological processes, AOPs are very effective in the complete mineralization of contaminants, but they are expensive when applied alone due to the requirement for energy and chemical reagents (Oller et al., 2011). Therefore, exploiting a chemical oxidation process such as Fentontype processes as the pre-treatment to transform the initially recalcitrant contaminants into more biodegradable intermediates is a promising approach with a significantly lower cost (Sarria et al., 2003; Lapertot et al., 2006). In a typical design approach, a Fentontype process oxidizes toxic substances, and once the required biodegradability has been achieved, the resulting effluent is transferred to a conventional biological treatment system (Sirtori et al., 2009a). Zapata et al. (2010) optimized the combination of photo-Fenton and biological processes based on H2O2 consumption in both pilot and full-scale plants. De la Cruz et al. (2013) successfully removed 22 ECs (more than 80%) from the effluent of secondary MWTP by ph-Fenton process at neutral pH condition and by using the iron content of effluent. The problem associated with the limitations of Fenton process, i.e. working at low pH value and high reagent consumption, were addressed by the design optimization of pilot setup at very low residence times (De la Cruz et al., 2012; De la Cruz et al., 2013). In addition, some treatment plants such as pharmaceutical wastewater treatment plants produce acidic effluents. Therefore, the natural pH level of these effluents are favorable for Fenton reactions without requiring any adjustment (Sirtori et al., 2009b). In another research, the combination of high frequency ultrasound and solar photo-Fenton process was used in
order to decrease both H2O2 consumption and the treatment time for ECs removal (Papoutsakis et al., 2015). In this case, a significant synergy was shown in the removal of both ECs and TOC. Nevertheless, the high cost of ultrasound process in comparison with photo-Fenton process still remains an obvious challenge during the treatment of real industrial effluents. Fenton-type processes can also be applied as the post-treatment step. Sirtori et al. (2009b) significantly reduced chemical consumption by employing the ph-Fenton process as a polishing step for the conventional biological stage. As a post-treatment step to polish off the effluent of rez-Estrada et al. (2007) used an iron sewage treatment plant, Pe catalyst at low concentration (2 mg/L) in order to avoid additional steps to reduce the iron content in the final effluent. Chi et al. (2013) utilized the heterogeneous Fenton process at pilot-scale for the removal of PhACs and hormones from secondary MWTP effluent for the first time. By using the post-treatment Fenton process, not only were the ECs removed successfully at natural pH and temperature, but also other parameters such as BOD, TOC, TSS, phosphate and turbidity were reduced; hence rendering this technology more competitive for full-scale applications. 6. Summary of findings The commercial availability of H2O2, its possible on-site storage, high solubility in water, thermal stability, production of two hydroxyl radicals by each molecule of H2O2, lack of mass transfer limitations associated with the dissolution of gas phase into the water (compared to O3), cost-effectiveness as a source of OH radicals, and the simple operation procedure highlight the advantages of using H2O2 as a the source of hydroxyl radicals (Kumar and Bansal, 2013). Nevertheless, there are certain limitations associated with the use of H2O2. They include the small absorption cross section of H2O2 at 254 nm which limits its application, especially when dissolved organic matters act as inner filters. The use of another source of irradiation such as Xedoped Hg arcs, which exhibit a strong emission in the spectral region of 210e240 nm, can resolve this problem since hydrogen peroxide has a higher molar absorption coefficient in this wavelength range. Conducting reactions in distilled water usually produces higher removal efficiencies in comparison with using surface waters or seawater. However, these results generally lead to erroneous large-scale designs and the overestimation of real process efficiency. Therefore, the presence of ions in real effluents and their contribution to increasing the removal efficiency of contaminants can be considered as a promising approach. On the other hand, applying the Fenton process in waters with high salinity may lead to the formation of by-products which are more toxic et al., 2009). than their parent compounds (Trovo The need for performing Fenton and photo-Fenton processes under acidic condition is often considered as one of their major disadvantages because it will impose additional costs for acidification and subsequent neutralization. Recently, it is proven that using ligands as a chemical additive can resolve this problem by forming complexes with iron ions (Romero et al., 2016a). Time-based kinetic relationships are commonly used to compare and evaluate the effectiveness of various removal processes. This is because other parameters such as energy demand and cost estimation can be determined based on the kinetic data. In the photo-Fenton process, the direct photolysis of chemical contaminants along with the reactions with hydrogen peroxide contribute to the degradation of contaminants. Despite the controlling effects of wavelength and irradiation intensity on the reaction rates of photolysis processes, these parameters are often ignored when kinetic models are developed.
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
Therefore, removal efficiencies cannot be properly evaluated by using the process dynamics, especially by using the data generated in bench-scale operations. Using the Fenton and photo-Fenton processes as a pretreatment step to biological processes produces promising results and has received attention in recent years. This approach will improve the biodegradability of organic pollutants and will enhance the efficiency of biological treatment (Kajitvichyanukul and Suntronvipart, 2006). However, the combined process will require the adjustment of pH to approximately 7 upstream of biological processes, proper handling of the chemical oxidant to prevent adverse effects on the biomass, and optimization of chemical dosage and reaction time. During the operation of homogeneous Fenton and photo-Fenton processes, the selection of iron source is important for effective wastewater treatment, and it should be selected based on the chemical characteristic of the target contaminants (Bautitz and Nogueira, 2007). If the composition of contaminants is unknown, a simple evaluation test at bench scale can be carried out in order to determine the most appropriate iron source. A major advantage of the photo-Fenton processes, besides their simplicity, is the possibility of operation at room temperature and at atmospheric pressure and the use of sunlight or near UV light as the energy sources. These operational aspects result in significant economic savings, especially in sunny countries. The use of Ferrioxalate as the source of iron is very efficient when using solar light irradiation, mostly due to its high absorption in the UVevis region and its high quantum yield of Fe2þ generation sito et al., 2016). (Expo In the case of effluents with high salinity, the photo-Fenton process is very slow and has a low economic viability, particularly when using an artificial radiation source (Bacardit et al., 2007). In these cases, employing solar photo-Fenton process is strongly advised. The use of heterogeneous Fenton-type process instead of the homogenous type offers significant advantages, such as the operation at milder pH, formation of significantly less sludge and elimination of the need for pH adjustment or neutralization. However, it has been demonstrated that the transparency of solution may be reduced by using suspended catalyst particles (Kaur et al., 2016). Therefore, the application of fixed catalysts, instead of suspended catalysts, and immobilization of iron onto a support medium are recommended.
7. Conclusion The presence of emerging contaminants such as pharmaceuticals in aquatic media has received considerable attention due to their potential health effects on human and their toxicity to various microorganisms. Conventional wastewater treatment processes such as flocculation, filtration, coagulation and sedimentation, are not efficient in the removal of pharmaceuticals from water. Therefore, these contaminants have the potential to accumulate and persist in the aquatic environment. Although pharmaceutical chemicals appear in the water at relatively low concentrations (ng/L to mg/L), they may impose acute toxicity effects on the environment. Advanced oxidation processes, which are based on the oxidation of organic contaminants by hydroxyl radicals, can be considered as an effective method to eliminate emerging contaminants from the environment. Among oxidation processes, Fenton processes are easy to operate, they are safe and relatively cheap, produce fast reactions and require environmentally-benign reagents. However, homogeneous Fenton processes suffer from certain limitations such as operation at a narrow pH range and the
685
generation of iron-containing sludge. In addition, the regeneration of iron ions is not feasible and the final effluent should be treated to meet the discharge standards for iron concentration. The immobilization of iron species on a support medium, in a process known as the heterogeneous Fenton, is an effective approach to overcome these limitations. In this process, the stability of catalyst and leaching of iron ions into the solution play significant roles and control the rate of reactions. Photo-Fenton process not only enhances the removal efficiency of pharmaceuticals, but also reduces the reagents consumption and waste generation. Although, Fenton processes do not always mineralize the contaminants, they can significantly enhance their biodegradability. Therefore, by integrating Fenton reactions with biological processes, it is possible to develop an advanced treatment process with higher efficiency and lower costs. Future research should focus on the use of solar energy, which can minimize energy costs, while reducing fossil fuel consumption and greenhouse gas emission, advancing sustainable development aspects of contaminant removal processes. References Alalm, M.G., Tawfik, A., Ookawara, S., 2015. Degradation of four pharmaceuticals by solar photo-Fenton process: kinetics and costs estimation. J. Environ. Chem. Eng. 3, 46e51. Alaton, I.A., Teksoy, S., 2007. Acid dyebath effluent pretreatment using Fenton’s reagent: process optimization, reaction kinetics and effects on acute toxicity. Dyes Pigm. 73, 31e39. Aleksi c, M., Kusi c, H., Koprivanac, N., Leszczynska, D., Bo zi c, A.L., 2010. Heterogeneous Fenton type processes for the degradation of organic dye pollutant in water d the application of zeolite assisted AOPs. Desalination 257, 22e29. Arslan-Alaton, I., Gurses, F., 2004. Photo-Fenton-like and photo-fenton-like oxidation of Procaine Penicillin G formulation effluent. J. Photochem. Photobiol. A Chem. 165, 165e175. rez, A.-A., Solís-Lo pez, M., Ramírez-Zamora, R.-M., Arzate-Salgado, S.-Y., Morales-Pe 2016. Evaluation of metallurgical slag as a Fenton-type photocatalyst for the degradation of an emerging pollutant: Diclofenac. Catal. Today 266, 126e135. €tzner, J., Chamarro, E., Esplugas, S., 2007. Effect of salinity on the Bacardit, J., Sto photo-fenton process. Ind. Eng. Chem. Res. 46, 7615e7619. Badawy, M.I., Ghaly, M.Y., Gad-Allah, T.A., 2006. Advanced oxidation processes for the removal of organophosphorus pesticides from wastewater. Desalination 194, 166e175. Badawy, M.I., Wahaab, R.A., El-Kalliny, A.S., 2009. Fenton-biological treatment processes for the removal of some pharmaceuticals from industrial wastewater. J. Hazard. Mater. 167, 567e574. Bae, S., Kim, D., Lee, W., 2013. Degradation of diclofenac by pyrite catalyzed Fenton oxidation. Appl. Catal. B Environ. 134e135, 93e102. rez, J.A., Acie n Ferna ndez, F.G., Casas Lo pez, J.L., Ballesteros Martín, M.M., S anchez Pe García-Ripoll, A.M., Arques, A., Oller, I., Malato Rodríguez, S., 2008. Combined photo-Fenton and biological oxidation for pesticide degradation: effect of phototreated intermediates on biodegradation kinetics. Chemosphere 70, 1476e1483. Batista, A.P.S., Nogueira, R.F.P., 2012. Parameters affecting sulfonamide photoFenton degradation e iron complexation and substituent group. J. Photochem. Photobiol. A Chem. 232, 8e13. Bautista, P., Mohedano, A.F., Casas, J.A., Zazo, J.A., Rodriguez, J.J., 2010a. Oxidation of cosmetic wastewaters with H2O2 using a Fe/g-Al2O3 catalyst. Water Sci. Technol. 61, 1631e1636. ndez, N., Casas, J.A., Rodriguez, J.J., 2010b. CataBautista, P., Mohedano, A.F., Mene lytic wet peroxide oxidation of cosmetic wastewaters with Fe-bearing catalysts. Catal. Today 151, 148e152. Bautitz, I.R., Nogueira, R.F.P., 2007. Degradation of tetracycline by photo-Fenton processdsolar irradiation and matrix effects. J. Photochem. Photobiol. A Chem. 187, 33e39. Bobu, M., Yediler, A., Siminiceanu, I., Schulte-Hostede, S., 2008. Degradation studies of ciprofloxacin on a pillared iron catalyst. Appl. Catal. B Environ. 83, 15e23. ~ ez, M., Fabregat-Safont, D., Morales, E., Pastor, L., Sancho, J.V., S Boix, C., Ib an anchezRamírez, J.E., Hern andez, F., 2016. Behaviour of emerging contaminants in sewage sludge after anaerobic digestion. Chemosphere 163, 296e304. Bowers, A.R., Gaddipati, P., Eckenfelder, W.W., Monsen, R.M., 1989. Treatment of toxic or refractory wastewaters with hydrogen peroxide. Water Sci. Technol. 21, 477e486. , M., Domínguez-Morueco, N., Migens, A., Molina, R., Martínez, F., Valc Catala arcel, Y., pez de Alda, M., Barcelo , D., Segura, Y., 2015. Elimination of Mastroianni, N., Lo drugs of abuse and their toxicity from natural waters by photo-Fenton treatment. Sci. Total Environ. 520, 198e205. Chi, G.T., Churchley, J., Huddersman, K.D., 2013. Pilot-scale removal of trace steroid hormones and pharmaceuticals and personal care products from municipal wastewater using a heterogeneous Fenton’s catalytic process. Int. J. Chem. Eng. 2013, 10.
686
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
Chong, S., Zhang, G., Zhang, N., Liu, Y., Zhu, J., Huang, T., Fang, S., 2016. Preparation of FeCeOx by ultrasonic impregnation method for heterogeneous Fenton degradation of diclofenac. Ultrason. Sonochem. 32, 231e240. , H.J., Rodrigues, A.E., Moreira, R.F.P.M., 2006. Dantas, T.L.P., Mendonça, V.P., Jose Treatment of textile wastewater by heterogeneous Fenton process using a new composite Fe2O3/carbon. Chem. Eng. J. 118, 77e82. De la Cruz, N., Esquius, L., Grandjean, D., Magnet, A., Tungler, A., de Alencastro, L.F., Pulgarín, C., 2013. Degradation of emergent contaminants by UV, UV/H2O2 and neutral photo-Fenton at pilot scale in a domestic wastewater treatment plant. Water Res. 47, 5836e5845. nez, J., Esplugas, S., Grandjean, D., de Alencastro, L.F., De la Cruz, N., Gime Pulgarín, C., 2012. Degradation of 32 emergent contaminants by UV and neutral photo-fenton in domestic wastewater effluent previously treated by activated sludge. Water Res. 46, 1947e1957. de Luna, M.D.G., Briones, R.M., Su, C.-C., Lu, M.-C., 2013. Kinetics of acetaminophen degradation by Fenton oxidation in a fluidized-bed reactor. Chemosphere 90, 1444e1448. lez, T., Palo, P., Cuerda-Correa, E.M., 2012. Fenton þ fentonDomínguez, J.R., Gonza like integrated process for carbamazepine degradation: optimizing the system. Ind. Eng. Chem. Res. 51, 2531e2538. Elmolla, E.S., Chaudhuri, M., 2009. Degradation of the antibiotics amoxicillin, ampicillin and cloxacillin in aqueous solution by the photo-Fenton process. J. Hazard. Mater. 172, 1476e1481. Elmolla, E.S., Chaudhuri, M., 2010. Comparison of different advanced oxidation processes for treatment of antibiotic aqueous solution. Desalination 256, 43e47. Elmolla, E.S., Chaudhuri, M., 2011. Combined photo-FentoneSBR process for antibiotic wastewater treatment. J. Hazard. Mater. 192, 1418e1426. Escher, B.I., Bramaz, N., Richter, M., Lienert, J., 2006. Comparative ecotoxicological hazard assessment of beta-blockers and their human metabolites using a mode-of-action-based test battery and a QSAR approach. Environ. Sci. Technol. 40, 7402e7408. sito, A.J., Dura n, A., Monteagudo, J.M., Acevedo, A., 2016. Solar photoExpo degradation of a pharmaceutical wastewater effluent in a semi-industrial autonomous plant. Chemosphere 150, 254e257. Feng, J., Hu, X., Yue, P.L., 2006. Effect of initial solution pH on the degradation of Orange II using clay-based Fe nanocomposites as heterogeneous photo-Fenton catalyst. Water Res. 40, 641e646. Ganiyu, S.O., van Hullebusch, E.D., Cretin, M., Esposito, G., Oturan, M.A., 2015. Coupling of membrane filtration and advanced oxidation processes for removal of pharmaceutical residues: a critical review. Sep. Purif. Technol. 156 (Part 3), 891e914. García-Ripoll, A., Amat, A.M., Arques, A., Vicente, R., Ballesteros Martín, M.M., rez, J.A.S., Oller, I., Malato, S., 2009. Confirming Pseudomonas putida as a reliPe able bioassay for demonstrating biocompatibility enhancement by solar photooxidative processes of a biorecalcitrant effluent. J. Hazard. Mater. 162, 1223e1227. Garrido-Ramírez, E.G., Theng, B.K.G., Mora, M.L., 2010. Clays and oxide minerals as catalysts and nanocatalysts in Fenton-like reactions d a review. Appl. Clay Sci. 47, 182e192. Georgi, A., Kopinke, F.-D., 2005. Interaction of adsorption and catalytic reactions in water decontamination processes: Part I. Oxidation of organic contaminants with hydrogen peroxide catalyzed by activated carbon. Appl. Catal. B Environ. 58, 9e18. Giri, A.S., Golder, A.K., 2014. Fenton, photo-fenton, H2O2 photolysis, and TiO2 photocatalysis for dipyrone oxidation: drug removal, mineralization, biodegradability, and degradation mechanism. Ind. Eng. Chem. Res. 53, 1351e1358. Gogate, P.R., Pandit, A.B., 2004. A review of imperative technologies for wastewater treatment I: oxidation technologies at ambient conditions. Adv. Environ. Res. 8, 501e551. lez, O., Sans, C., Esplugas, S., 2007. Sulfamethoxazole abatement by photoGonza Fenton: toxicity, inhibition and biodegradability assessment of intermediates. J. Hazard. Mater. 146, 459e464. Gonzalez, O., Sans, C., Esplugas, S., Malato, S., 2009. Application of solar advanced oxidation processes to the degradation of the antibiotic sulfamethoxazole. Photochem. Photobiol. Sci. 8, 1032e1039. Guo, R., Xie, X., Chen, J., 2015. The degradation of antibiotic amoxicillin in the Fenton-activated sludge combined system. Environ. Technol. 36, 844e851. Hartmann, M., Kullmann, S., Keller, H., 2010. Wastewater treatment with heterogeneous Fenton-type catalysts based on porous materials. J. Mater. Chem. 20, 9002e9017. Hermosilla, D., Cortijo, M., Huang, C.P., 2009. Optimizing the treatment of landfill leachate by conventional Fenton and photo-Fenton processes. Sci. Total Environ. 407, 3473e3481. Herney-Ramirez, J., Vicente, M.A., Madeira, L.M., 2010. Heterogeneous photo-Fenton oxidation with pillared clay-based catalysts for wastewater treatment: a review. Appl. Catal. B Environ. 98, 10e26. Hu, X., Liu, B., Deng, Y., Chen, H., Luo, S., Sun, C., Yang, P., Yang, S., 2011. Adsorption and heterogeneous Fenton degradation of 17a-methyltestosterone on nano Fe3O4/MWCNTs in aqueous solution. Appl. Catal. B Environ. 107, 274e283. Huerta-Fontela, M., Galceran, M.T., Ventura, F., 2011. Occurrence and removal of pharmaceuticals and hormones through drinking water treatment. Water Res. 45, 1432e1442. Jaafarzadeh, N., Kakavandi, B., Takdastan, A., Kalantary, R.R., Azizi, M., Jorfi, S., 2015. Powder activated carbon/Fe3O4 hybrid composite as a highly efficient heterogeneous catalyst for Fenton oxidation of tetracycline: degradation mechanism and kinetic. RSC Adv. 5, 84718e84728.
Jonidi Jafari, A., Kakavandi, B., Jaafarzadeh, N., Rezaei Kalantary, R., Ahmadi, M., Akbar Babaei, A., 2016. Fenton-like catalytic oxidation of tetracycline by AC@ Fe3O4 as a heterogeneous persulfate activator: adsorption and degradation studies. J. Ind. Eng. Chem. http://dx.doi.org/10.1016/j.jiec.2016.09.044. Kajitvichyanukul, P., Suntronvipart, N., 2006. Evaluation of biodegradability and oxidation degree of hospital wastewater using photo-Fenton process as the pretreatment method. J. Hazard. Mater. 138, 384e391. Kasiri, M.B., Aleboyeh, H., Aleboyeh, A., 2008. Degradation of Acid Blue 74 using FeZSM5 zeolite as a heterogeneous photo-Fenton catalyst. Appl. Catal. B Environ. 84, 9e15. Kaur, A., Umar, A., Kansal, S.K., 2016. Heterogeneous photocatalytic studies of analgesic and non-steroidal anti-inflammatory drugs. Appl. Catal. A General 510, 134e155. Kavitha, V., Palanivelu, K., 2004. The role of ferrous ion in Fenton and photo-Fenton processes for the degradation of phenol. Chemosphere 55, 1235e1243. Kuan, C.-C., Chang, S.-Y., Schroeder, S.L.M., 2015. Fenton-like oxidation of 4chlorophenol: homogeneous or heterogeneous? Ind. Eng. Chem. Res. 54, 8122e8129. Kumar, J., Bansal, A., 2013. Photocatalysis by nanoparticles of titanium dioxide for drinking water purification: a conceptual and state-of-art review. Mater. Sci. Forum 764, 130e150. Kunde, G.B., Yadav, G.D., 2015. Synthesis, characterization and application of ironaluminate nodules in advanced Fenton’s oxidation process. J. Environ. Chem. Eng. 3, 2010e2021. Kusi c, H., Koprivanac, N., Bo zi c, A.L., Selanec, I., 2006a. Photo-assisted Fenton type processes for the degradation of phenol: a kinetic study. J. Hazard. Mater. 136, 632e644. Kusi c, H., Koprivanac, N., Selanec, I., 2006b. Fe-exchanged zeolite as the effective heterogeneous Fenton-type catalyst for the organic pollutant minimization: UV irradiation assistance. Chemosphere 65, 65e73. Kuznetsova, E.V., Savinov, E.N., Vostrikova, L.A., Parmon, V.N., 2004. Heterogeneous catalysis in the fenton-type system FeZSM-5/H2O2. Appl. Catal. B Environ. 51, 165e170. Kwon, B.G., Lee, D.S., Kang, N., Yoon, J., 1999. Characteristics of p-chlorophenol oxidation by Fenton’s reagent. Water Res. 33, 2110e2118. Lan, H., Wang, A., Liu, R., Liu, H., Qu, J., 2015. Heterogeneous photo-Fenton degradation of acid red B over Fe2O3 supported on activated carbon fiber. J. Hazard. Mater. 285, 167e172. ndez-Ib ~ ez, P., Maldonado, M.I., Pe rez-Estrada, L., Lapertot, M., Pulgarín, C., Ferna an Oller, I., Gernjak, W., Malato, S., 2006. Enhancing biodegradability of priority substances (pesticides) by solar photo-Fenton. Water Res. 40, 1086e1094. Legrini, O., Oliveros, E., Braun, A.M., 1993. Photochemical processes for water treatment. Chem. Rev. 93, 671e698. Li, H., Li, Y., Xiang, L., Huang, Q., Qiu, J., Zhang, H., Sivaiah, M.V., Baron, F., Barrault, J., Petit, S., Valange, S., 2015. Heterogeneous photo-Fenton decolorization of Orange II over Al-pillared Fe-smectite: response surface approach, degradation pathway, and toxicity evaluation. J. Hazard. Mater. 287, 32e41. Li, R., Yang, C., Chen, H., Zeng, G., Yu, G., Guo, J., 2009. Removal of triazophos pesticide from wastewater with Fenton reagent. J. Hazard. Mater. 167, 1028e1032. Li, W.C., 2014. Occurrence, sources, and fate of pharmaceuticals in aquatic environment and soil. Environ. Pollut. 187, 193e201. Li, Y., Lu, Y., Zhu, X., 2006. Photo-Fenton discoloration of the azo dye X-3B over pillared bentonites containing iron. J. Hazard. Mater. 132, 196e201. Liou, R.-M., Chen, S.-H., Hung, M.-Y., Hsu, C.-S., Lai, J.-Y., 2005. Fe (III) supported on resin as effective catalyst for the heterogeneous oxidation of phenol in aqueous solution. Chemosphere 59, 117e125. Liu, Z.-h., Kanjo, Y., Mizutani, S., 2009. Removal mechanisms for endocrine disrupting compounds (EDCs) in wastewater treatment d physical means, biodegradation, and chemical advanced oxidation: a review. Sci. Total Environ. 407, 731e748. Machulek, A., Moraes, J.E.F., Vautier-Giongo, C., Silverio, C.A., Friedrich, L.C., Nascimento, C.A.O., Gonzalez, M.C., Quina, F.H., 2007. Abatement of the inhibitory effect of chloride anions on the photo-fenton process. Environ. Sci. Technol. 41, 8459e8463. Machulek, A., Vautier-Giongo, C., Moraes, J.E.F., Nascimento, C.A.O., Quina, F.H., 2006. Laser flash photolysis study of the photocatalytic step of the photo-fenton reaction in saline solution. Photochem. Photobiol. 82, 208e212. , L., 2015. FentonMackulak, T., Mosný, M., Grabic, R., Golovko, O., Koba, O., Birosova like reaction: a possible way to efficiently remove illicit drugs and pharmaceuticals from wastewater. Environ. Toxicol. Pharmacol. 39, 483e488. Maezono, T., Tokumura, M., Sekine, M., Kawase, Y., 2011. Hydroxyl radical concentration profile in photo-Fenton oxidation process: generation and consumption of hydroxyl radicals during the discoloration of azo-dye Orange II. Chemosphere 82, 1422e1430. Mameri, Y., Debbache, N., Benacherine, M.e.m., Seraghni, N., Sehili, T., 2016. Heterogeneous photodegradation of paracetamol using Goethite/H2O2 and Goethite/oxalic acid systems under artificial and natural light. J. Photochem. Photobiol. A Chem. 315, 129e137. , V., Brix, H., 2012. Occurrence and Matamoros, V., Arias, C.A., Nguyen, L.X., Salvado behavior of emerging contaminants in surface water and a restored wetland. Chemosphere 88, 1083e1089. ndez-Arriaga, F., Esplugas, S., Gime nez, J., 2010. Degradation of the emerging Me contaminant ibuprofen in water by photo-Fenton. Water Res. 44, 589e595. Michael, I., Hapeshi, E., Michael, C., Fatta-Kassinos, D., 2010. Solar Fenton and solar
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688 TiO2 catalytic treatment of ofloxacin in secondary treated effluents: evaluation of operational and kinetic parameters. Water Res. 44, 5450e5462. Mirzaei, A., Chen, Z., Haghighat, F., Yerushalmi, L., 2016. Removal of pharmaceuticals and endocrine disrupting compounds from water by zinc oxide-based photocatalytic degradation: a review. Sustain. Cities Soc. 27, 407e418. Mirzaei, A., Ebadi, A., Khajavi, P., 2013. Kinetic and equilibrium modeling of single and binary adsorption of methyl tert-butyl ether (MTBE) and tert-butyl alcohol (TBA) onto nano-perfluorooctyl alumina. Chem. Eng. J. 231, 550e560. Mompelat, S., Le Bot, B., Thomas, O., 2009. Occurrence and fate of pharmaceutical products and by-products, from resource to drinking water. Environ. Int. 35, 803e814. Nakada, N., Shinohara, H., Murata, A., Kiri, K., Managaki, S., Sato, N., Takada, H., 2007. Removal of selected pharmaceuticals and personal care products (PPCPs) and endocrine-disrupting chemicals (EDCs) during sand filtration and ozonation at a municipal sewage treatment plant. Water Res. 41, 4373e4382. Navalon, S., Alvaro, M., Garcia, H., 2010. Heterogeneous Fenton catalysts based on clays, silicas and zeolites. Appl. Catal. B Environ. 99, 1e26. Nidheesh, P.V., Gandhimathi, R., Ramesh, S.T., 2013. Degradation of dyes from aqueous solution by Fenton processes: a review. Environ. Sci. Pollut. Res. 20, 2099e2132. Niu, H., Dizhang, Meng, Z., Cai, Y., 2012. Fast defluorination and removal of norfloxacin by alginate/Fe@Fe3O4 core/shell structured nanoparticles. J. Hazard. Mater. 227e228, 195e203. Niu, H., Zhang, D., Zhang, S., Zhang, X., Meng, Z., Cai, Y., 2011. Humic acid coated Fe3O4 magnetic nanoparticles as highly efficient Fenton-like catalyst for complete mineralization of sulfathiazole. J. Hazard. Mater. 190, 559e565. , A.G., 2005. Influence of the iron source on the Nogueira, R.F.P., Silva, M.R.A., Trovo solar photo-Fenton degradation of different classes of organic compounds. Sol. Energy 79, 384e392. rez, J.A., 2011. Combination of Advanced Oxidation Oller, I., Malato, S., S anchez-Pe Processes and biological treatments for wastewater decontaminationda review. Sci. Total Environ. 409, 4141e4166. Pal, A., Gin, K.Y.-H., Lin, A.Y.-C., Reinhard, M., 2010. Impacts of emerging organic contaminants on freshwater resources: review of recent occurrences, sources, fate and effects. Sci. Total Environ. 408, 6062e6069. Papoutsakis, S., Miralles-Cuevas, S., Gondrexon, N., Baup, S., Malato, S., Pulgarin, C., 2015. Coupling between high-frequency ultrasound and solar photo-Fenton at pilot scale for the treatment of organic contaminants: an initial approach. Ultrason. Sonochem. 22, 527e534. ringer, P., Pulgarin, C., 2000. Photochemical versus Parra, S., Sarria, V., Malato, S., Pe coupled photochemicalebiological flow system for the treatment of two biorecalcitrant herbicides: metobromuron and isoproturon. Appl. Catal. B Environ. 27, 153e168. rez-Estrada, L.A., Malato, S., Agüera, A., Ferna ndez-Alba, A.R., 2007. Degradation Pe of dipyrone and its main intermediates by solar AOPs: identification of intermediate products and toxicity assessment. Catal. Today 129, 207e214. rez-Estrada, L.A., Malato, S., Gernjak, W., Agüera, A., Thurman, E.M., Ferrer, I., Pe ndez-Alba, A.R., 2005a. Photo-Fenton degradation of Diclofenac: identiFerna fication of main intermediates and degradation pathway. Environ. Sci. Technol. 39, 8300e8306. rez-Estrada, L.A., Maldonado, M.I., Gernjak, W., Agüera, A., Fern Pe andez-Alba, A.R., Ballesteros, M.M., Malato, S., 2005b. Decomposition of diclofenac by solar driven photocatalysis at pilot plant scale. Catal. Today 101, 219e226. Perisic, D.J., Gilja, V., Stankov, M.N., Katancic, Z., Kusic, H., Stangar, U.L., Dionysiou, D.D., Bozic, A.L., 2016. Removal of diclofenac from water by zeoliteassisted advanced oxidation processes. J. Photochem. Photobiol. A Chem. 321, 238e247. Pignatello, J.J., Liu, D., Huston, P., 1999. Evidence for an additional oxidant in the photoassisted fenton reaction. Environ. Sci. Technol. 33, 1832e1839. Prieto-Rodríguez, L., Oller, I., Klamerth, N., Agüera, A., Rodríguez, E.M., Malato, S., 2013. Application of solar AOPs and ozonation for elimination of micropollutants in municipal wastewater treatment plant effluents. Water Res. 47, 1521e1528. ~ ones, D.H., Alvarez, Quin P.M., Rey, A., Beltr an, F.J., 2015. Removal of emerging contaminants from municipal WWTP secondary effluents by solar photocatalytic ozonation. A pilot-scale study. Sep. Purif. Technol. 149, 132e139. Rache, M.L., García, A.R., Zea, H.R., Silva, A.M.T., Madeira, L.M., Ramírez, J.H., 2014. Azo-dye orange II degradation by the heterogeneous Fenton-like process using a zeolite Y-Fe catalystdkinetics with a model based on the Fermi’s equation. Appl. Catal. B Environ. 146, 192e200. Rad, L.R., Haririan, I., Divsar, F., 2015. Comparison of adsorption and photo-Fenton processes for phenol and paracetamol removing from aqueous solutions: single and binary systems. Spectrochim. Acta Part A Mol. Biomol. Spectrosc. 136 (Part B), 423e428. , D., Malato, S., 2009. Solar photoRadjenovi c, J., Sirtori, C., Petrovi c, M., Barcelo catalytic degradation of persistent pharmaceuticals at pilot-scale: kinetics and characterization of major intermediate products. Appl. Catal. B Environ. 89, 255e264. Rahim Pouran, S., Abdul Aziz, A.R., Wan Daud, W.M.A., 2015. Review on the main advances in photo-Fenton oxidation system for recalcitrant wastewaters. J. Ind. Eng. Chem. 21, 53e69. dar, F.J., Pe rez-Cadenas, A.F., Moreno-Castilla, C., Ramirez, J.H., Maldonado-Ho Costa, C.A., Madeira, L.M., 2007. Azo-dye Orange II degradation by heterogeneous Fenton-like reaction using carbon-Fe catalysts. Appl. Catal. B Environ. 75, 312e323.
687
Ravina, M., Campanella, L., Kiwi, J., 2002. Accelerated mineralization of the drug Diclofenac via Fenton reactions in a concentric photo-reactor. Water Res. 36, 3553e3560. ndez, G., Ledesma, B., Alvarez, n, F.J., 2009. PhotoRodríguez, E., Ferna P., Beltra catalytic degradation of organics in water in the presence of iron oxides: influence of carboxylic acids. Appl. Catal. B Environ. 92, 240e249. ndez, G., Alvarez, n, F.J., 2011. Rodríguez, E.M., Ferna P.M., Hern andez, R., Beltra Photocatalytic degradation of organics in water in the presence of iron oxides: effects of pH and light source. Appl. Catal. B Environ. 102, 572e583. nez, J., Esplugas, S., 2016a. Enhancement of Romero, V., Acevedo, S., Marco, P., Gime Fenton and photo-Fenton processes at initial circumneutral pH for the degradation of the b-blocker metoprolol. Water Res. 88, 449e457. nez, J., Esplugas, S., 2016b. Romero, V., Gonz alez, O., Bayarri, B., Marco, P., Gime Degradation of Metoprolol by photo-Fenton: comparison of different photoreactors performance. Chem. Eng. J. 283, 639e648. Sanchis, S., Polo, A.M., Tobajas, M., Rodriguez, J.J., Mohedano, A.F., 2013. Degradation of chlorophenoxy herbicides by coupled Fenton and biological oxidation. Chemosphere 93, 115e122. ringer, P., Pulgarin, C., 2003. Degradation of a biorecalcitrant Sarria, V., Deront, M., Pe dye precursor present in industrial wastewaters by a new integrated iron(III) photoassistedebiological treatment. Appl. Catal. B Environ. 40, 231e246. Shemer, H., Kunukcu, Y.K., Linden, K.G., 2006. Degradation of the pharmaceutical Metronidazole via UV, Fenton and photo-Fenton processes. Chemosphere 63, 269e276. Sirtori, C., Zapata, A., Gernjak, W., Malato, S., Lopez, A., Agüera, A., 2011. Solar photoFenton degradation of nalidixic acid in waters and wastewaters of different composition. Analytical assessment by LCeTOF-MS. Water Res. 45, 1736e1744. Sirtori, C., Zapata, A., Oller, I., Gernjak, W., Agüera, A., Malato, S., 2009a. Decontamination industrial pharmaceutical wastewater by combining solar photoFenton and biological treatment. Water Res. 43, 661e668. Sirtori, C., Zapata, A., Oller, I., Gernjak, W., Agüera, A., Malato, S., 2009b. Solar photofenton as finishing step for biological treatment of a pharmaceutical wastewater. Environ. Sci. Technol. 43, 1185e1191. Snyder, S.A., Adham, S., Redding, A.M., Cannon, F.S., DeCarolis, J., Oppenheimer, J., Wert, E.C., Yoon, Y., 2007. Role of membranes and activated carbon in the removal of endocrine disruptors and pharmaceuticals. Desalination 202, 156e181. Song, R., Bai, B., Puma, G.L., Wang, H., Suo, Y., 2015. Biosorption of azo dyes by raspberry-like Fe3O4@yeast magnetic microspheres and their efficient regeneration using heterogeneous Fenton-like catalytic processes over an up-flow packed reactor. React. Kinet. Mech. Catal. 115, 547e562. Soon, A.N., Hameed, B.H., 2011. Heterogeneous catalytic treatment of synthetic dyes in aqueous media using Fenton and photo-assisted Fenton process. Desalination 269, 1e16. Sum, O.S.N., Feng, J., Hub, X., Yue, P.L., 2005. Photo-assisted fenton mineralization of an azo-dye acid black 1 using a modified laponite clay-based Fe nanocomposite as a heterogeneous catalyst. Top. Catal. 33, 233e242. Sun, S.-P., Lemley, A.T., 2011. p-Nitrophenol degradation by a heterogeneous Fentonlike reaction on nano-magnetite: process optimization, kinetics, and degradation pathways. J. Mol. Catal. A Chem. 349, 71e79. Sun, S.-P., Zeng, X., Lemley, A.T., 2013. Nano-magnetite catalyzed heterogeneous Fenton-like degradation of emerging contaminants carbamazepine and ibuprofen in aqueous suspensions and montmorillonite clay slurries at neutral pH. J. Mol. Catal. A Chem. 371, 94e103. Tang, W.Z., Huang, C.P., 1996. 2,4-Dichlorophenol oxidation kinetics by Fenton’s reagent. Environ. Technol. 17, 1371e1378. Tekbas¸, M., Yatmaz, H.C., Bektas¸, N., 2008. Heterogeneous photo-Fenton oxidation of reactive azo dye solutions using iron exchanged zeolite as a catalyst. Microporous Mesoporous Mater. 115, 594e602. Tekin, H., Bilkay, O., Ataberk, S.S., Balta, T.H., Ceribasi, I.H., Sanin, F.D., Dilek, F.B., Yetis, U., 2006. Use of Fenton oxidation to improve the biodegradability of a pharmaceutical wastewater. J. Hazard. Mater. 136, 258e265. Tijani, J.O., Fatoba, O.O., Babajide, O.O., Petrik, L.F., 2016. Pharmaceuticals, endocrine disruptors, personal care products, nanomaterials and perfluorinated pollutants: a review. Environ. Chem. Lett. 14, 27e49. Titouhi, H., Belgaied, J.-E., 2016. Heterogeneous Fenton oxidation of ofloxacin drug by iron alginate support. Environ. Technol. 1e13. , A.G., de Paiva, V.A.B., Machado, A.E.H., de Oliveira, C.A., Santos, R.O., 2013. Trovo Degradation of the antibiotic chloramphenicol by photo-Fenton process at labscale and solar pilot plant: kinetic, toxicity and inactivation assessment. Sol. Energy 97, 596e604. , A.G., Melo, S.A.S., Nogueira, R.F.P., 2008. Photodegradation of the pharmaTrovo ceuticals amoxicillin, bezafibrate and paracetamol by the photo-Fenton processdapplication to sewage treatment plant effluent. J. Photochem. Photobiol. A Chem. 198, 215e220. , A.G., Nogueira, R.F.P., Agüera, A., Fernandez-Alba, A.R., Sirtori, C., Malato, S., Trovo 2009. Degradation of sulfamethoxazole in water by solar photo-Fenton. Chemical and toxicological evaluation. Water Res. 43, 3922e3931. , A.G., Pupo Nogueira, R.F., Agüera, A., Fernandez-Alba, A.R., Malato, S., 2011. Trovo Degradation of the antibiotic amoxicillin by photo-Fenton process e chemical and toxicological assessment. Water Res. 45, 1394e1402. , A.G., Pupo Nogueira, R.F., Agüera, A., Fernandez-Alba, A.R., Malato, S., 2012. Trovo Paracetamol degradation intermediates and toxicity during photo-Fenton treatment using different iron species. Water Res. 46, 5374e5380. Turan-Ertas, T., Gurol, M.D., 2002. Oxidation of diethylene glycol with ozone and
688
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
modified Fenton processes. Chemosphere 47, 293e301. n ~ez, J., Zaror, C., Salazar, R.A., Pe rezVel asquez, M., Santander, I.P., Contreras, D.R., Ya Moya, M., Mansilla, H.D., 2014. Oxidative degradation of sulfathiazole by Fenton and photo-Fenton reactions. J. Environ. Sci. Health, Part A 49, 661e670. Velichkova, F., Julcour-Lebigue, C., Koumanova, B., Delmas, H., 2013. Heterogeneous Fenton oxidation of paracetamol using iron oxide (nano)particles. J. Environ. Chem. Eng. 1, 1214e1222. Veloutsou, S., Bizani, E., Fytianos, K., 2014. Photo-Fenton decomposition of bblockers atenolol and metoprolol; study and optimization of system parameters and identification of intermediates. Chemosphere 107, 180e186. Vergili, I., Gencdal, S., 2015. Applicability of combined Fenton oxidation and nanofiltration to pharmaceutical wastewater. Desalination Water Treat. 56, 3501e3509. Yuan, S., Gou, N., Alshawabkeh, A.N., Gu, A.Z., 2013. Efficient degradation of contaminants of emerging concerns by a new electro-Fenton process with Ti/MMO cathode. Chemosphere 93, 2796e2804. nchez-Pe rez, J.A., Oller, I., Maldonado, M.I., 2010. Scale-up Zapata, A., Malato, S., Sa strategy for a combined solar photo-Fenton/biological system for remediation of pesticide-contaminated water. Catal. Today 151, 100e106. Zapata, A., Oller, I., Gallay, R., Pulgarín, C., Maldonado, M.I., Malato, S., Gernjak, W., 2008. Comparison of photo-fenton treatment and coupled photo-fenton and
biological treatment for detoxification of pharmaceutical industry contaminants. J. Adv. Oxid. Technol. 11, 261e269. nchez-Pe rez, J.A., Mantzavinos, D., Maldonado, M.I., Zapata, A., Velegraki, T., Sa Malato, S., 2009. Solar photo-Fenton treatment of pesticides in water: effect of iron concentration on degradation and assessment of ecotoxicity and biodegradability. Appl. Catal. B Environ. 88, 448e454. Zazo, J.A., Casas, J.A., Mohedano, A.F., Rodríguez, J.J., 2006. Catalytic wet peroxide oxidation of phenol with a Fe/active carbon catalyst. Appl. Catal. B Environ. 65, 261e268. Zazo, J.A., Pliego, G., Blasco, S., Casas, J.A., Rodriguez, J.J., 2011. Intensification of the fenton process by increasing the temperature. Ind. Eng. Chem. Res. 50, 866e870. Zhang, H., Choi, H.J., Huang, C.-P., 2005. Optimization of Fenton process for the treatment of landfill leachate. J. Hazard. Mater. 125, 166e174. Zhang, X., Bai, B., Li Puma, G., Wang, H., Suo, Y., 2016a. Novel sea buckthorn biocarbon SBC@b-FeOOH composites: efficient removal of doxycycline in aqueous solution in a fixed-bed through synergistic adsorption and heterogeneous Fenton-like reaction. Chem. Eng. J. 284, 698e707. Zhang, Y., Liu, C., Xu, B., Qi, F., Chu, W., 2016b. Degradation of benzotriazole by a novel Fenton-like reaction with mesoporous Cu/MnO2: combination of adsorption and catalysis oxidation. Appl. Catal. B Environ. 199, 447e457.
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Review
Removal of pharmaceuticals from water by homo/heterogonous Fenton-type processes e A review Amir Mirzaei, Zhi Chen*, Fariborz Haghighat, Laleh Yerushalmi Department of Building, Civil and Environmental Engineering, Concordia University, Montreal, H3G 1M8, Canada
h i g h l i g h t s Safe, environmentally-benign and relatively cheap reagents of Fenton-type reactions. Comprehensive review of controlling parameters on the Fenton-type reactions. Narrow pH range and iron-containing sludge as the homogenous Fenton limitations. The slow reaction kinetics is the major drawback of heterogeneous Fenton reaction. Recent strategies to address these limitations are presented.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 27 September 2016 Received in revised form 1 February 2017 Accepted 3 February 2017 Available online 4 February 2017
The presence of emerging contaminants such as pharmaceuticals in natural waters has raised increasing concern due to their frequent appearance and persistence in the aquatic ecosystem and the threat to health and safety of aquatic life, even at trace concentrations. Conventional water treatment processes are known to be generally inadequate for the elimination of these persistent contaminants. Therefore, the use of advanced oxidation processes (AOPs) which are able to efficiently oxidize organic pollutants has attracted a great amount of attention. The main limitation of AOPs lies in their high operating costs associated with the consumption of energy and chemicals. Fenton-based processes, which utilize nontoxic and common reagents and potentially can exploit solar energy, will considerably reduce the removal cost of recalcitrant contaminants. The disadvantages of homogeneous Fenton processes, such as the generation of high amounts of iron-containing sludge and limited operational range of pH, have prompted much attention to the use of heterogeneous Fenton processes. In this review, the impacts of some controlling parameters including the H2O2 and catalyst dosage, solution pH, initial contaminants concentrations, temperature, type of catalyst, intensity of irradiation, reaction time and feeding mode on the removal efficiencies of hetero/homogeneous Fenton processes are discussed. In addition, the combination of Fenton-type processes with biological systems as the pre/post treatment stages in pilot-scale operations is considered. The reported experimental results obtained by using Fenton and photo-Fenton processes for the elimination of pharmaceutical contaminants are also compiled and evaluated. © 2017 Elsevier Ltd. All rights reserved.
Handling Editor: Xiangru Zhang Keywords: Water treatment Emerging contaminants Pharmaceuticals Toxicity Photo-Fenton Heterogeneous Fenton process
Contents 1. 2.
3.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 666 Mechanism of Fenton-type reactions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 667 2.1. Homogeneous Fenton processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 667 2.2. Heterogeneous Fenton processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 669 2.2.1. Iron oxide minerals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 669 2.2.2. Support material in heterogeneous Fenton processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 669 Major controlling parameters in the Fenton reaction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 674
* Corresponding author. E-mail address: [email protected] (Z. Chen). http://dx.doi.org/10.1016/j.chemosphere.2017.02.019 0045-6535/© 2017 Elsevier Ltd. All rights reserved.
666
4. 5. 6. 7.
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
3.1. Effect of reagents concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 674 3.2. Effect of initial contaminants concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 678 3.3. Effect of catalyst type . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 678 3.4. Effect of pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 679 3.5. Effect of intensity of radiation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 680 3.6. Effect of water matrix . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 681 3.7. Effect of the salinity of substrate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 681 3.8. Effect of feeding mode . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 682 3.9. Effect of temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 682 3.10. Effect of reaction time . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 682 3.11. Deactivation of catalyst . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 683 3.12. By-products formation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 683 Toxicity and biodegradability tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 683 Large-scale application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 684 Summary of findings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 684 Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 685 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 685
Nomenclature AAP ALC AMX AOP ATL BLB BZA BZF CAP CBZ CPC CWPO DCF DOC DW EC EDDP EDTA FeOx GAC HA HP IBP
acetaminophen amphetamine-like compound amoxicillin advanced oxidation process b-blocker atenolol black blue lamp benzotriazole bezafibrate chloramphenicol carbamazepine compound parabolic collector catalytic wet peroxide oxidation diclofenac dissolved organic carbon distilled water emerging contaminant 2-ethylene 1,5-dimethyl3,3-diphenylpyrrolidine ethylenediaminetetraacetic acid ferrioxalate granular activated carbon humic acid high pressure ibuprofen
1. Introduction Emerging contaminants (ECs) such as pharmaceuticals (PhACs) appear in the water and wastewater treatment plants mainly through excretion and/or improper disposal of outdated or unused medication (Boix et al., 2016). The presence of these contaminants has raised concerns since they are potentially toxic to aquatic organisms even at trace concentrations (ng/L or mg/L) (Yuan et al., 2013; Mirzaei et al., 2016). Besides, some emerging contaminants exhibit non-target effects as well as mixture toxicity in the environment. This means that PhACs have the potential to show specific effects, which are irrelevant to their therapeutic purposes. In addition, their incomplete mineralization may lead to the forma et al., tion of additional toxic chemicals (Escher et al., 2006; Catala 2015). These facts, as well as the low biodegradability of PhACs, are
L-H LP MET MP MR MWTP NF NXA OFX PAC PCT PCPs PhACs ph-F PILCs PPG RES S.E. SMX STP TC TOC TSS WWTPs
LangmuireHinshelwood low pressure metoprolol medium pressure molar ratio municipal wastewater treatment plant nanofiltration nalidixic acid ofloxacin powder activated carbon paracetamol personal care products pharmaceuticals photo-Fenton pillared interlayered clays procaine penicillin G resorcinol synthetic MWTP effluent sulfamethoxazole sewage treatment plant tetracycline total organic carbon total suspended solids wastewater treatment plants
significant concerns for public health and require urgent attention rez-Estrada et al., 2005b). Therefore, wastewaters containing (Pe PhACs should be treated prior to their discharge into the surface waters. In most cases, due to the poor removal of emerging contaminants in conventional treatment systems, these chemical compounds end up in soil, surface waters or even in drinking water (Mompelat et al., 2009; Hu et al., 2011; Matamoros et al., 2012; Veloutsou et al., 2014). Studies have demonstrated that about 64% of ECs are removed by less than 50% while 9% are not removed at all by conventional biological treatment processes (Chi et al., 2013). Several review papers have addressed the occurrence, fate and transport of ECs in different environmental compartments (Mompelat et al., 2009; Pal et al., 2010; Huerta-Fontela et al., 2011; Li, 2014; Tijani et al., 2016). Although some physical processes such as adsorption are effective for the removal of organic pollutants,
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
they are non-destructive and merely cause a phase transfer of contaminants, producing highly contaminant secondary wastes which require further treatment (Dantas et al., 2006; Mirzaei et al., 2013; Prieto-Rodríguez et al., 2013). In addition, the adsorption capacity of ECs on adsorbents (mostly GAC) strongly depends on environmental parameters such as pH, adsorbent concentration, presence of humic acid and salinity (Liu et al., 2009). A severe decrease in the adsorption of ECs on activated carbon was reported for the effluent of WWTPs as compared to surface waters (Snyder et al., 2007). Moreover, the widely used thermal regeneration of spent activated carbon is costly and indirectly causes an additional environmental problem (Georgi and Kopinke, 2005). Membrane processes have the ability to produce high quality effluents and reduce or remove trace amounts of micropollutants including ECs (Snyder et al., 2007; Liu et al., 2009). However, membrane filtration processes are often costly at full-scale operation because of frequent fouling and the requirement for backwashing that may limit their application, especially in the treatment of wastewaters at high flow rates (Chi et al., 2013; Ganiyu et al., 2015). The use of chemical techniques such as advanced oxidation processes (AOPs) may be considered as an alternative to this problem (Yuan et al., 2013). AOPs do not generate waste and they can be applied as a pre-treatment process prior to the biological treatment which can enhance the biodegradability of refractory compounds (Soon and Hameed, 2011). As an example of chemical techniques, the high removal efficiencies of a broad range of PhACs by ozonation have been reported (Nakada et al., 2007). However, concerns related to the potential toxicity of the partially oxidized PhACs during ozonation and the large energy consumption during ozone generation are considered as major obstacles to the commercial application of ozonation process (Nakada et al., 2007; Chi et al., 2013). Among AOPs, Fenton-type reactions have been identified as effective methods which produce hydroxyl radicals by the reaction between iron salts and hydrogen peroxide (Hartmann et al., 2010; Maezono et al., 2011). The advantages of these processes are the safe and environmentally-benign nature of reagents and relatively simple operating principles as well as short reaction time and the rez-Estrada et al., 2005b; absence of mass transfer limitations (Pe squez et al., 2014). Kajitvichyanukul and Suntronvipart, 2006; Vela Mackulak et al. (2015) successfully removed 13 drugs to concentrations below detection limit from a wastewater treatment plant by Fenton and Fenton-like processes. In addition, the Fenton process can be used as an effective pre-treatment to improve the biodegradability of wastewater contaminants, especially during the treatment of concentrated wastewaters containing recalcitrant compounds such as those emerging from drug factories and/or hospitals (Tekin et al., 2006; Sanchis et al., 2013). The photo-Fenton process is similar to the Fenton process but it offers higher removal efficiencies by employing irradiation. In this process, the generation of hydroxyl radicals accelerates in comparison with the “simple” Fenton process (Herney-Ramirez et al., 2010). However, the homogeneous Fenton and photo-Fenton reactions produce a secondary pollutant (ferric ion) following the treatment of organic contaminants (Sum et al., 2005). Thus, numerous attempts have recently been made to develop heterogeneous catalysts with high photocatalytic activity, long-term stability, and low cost for the Fenton and photo-Fenton oxidation in wastewater treatment. The heterogeneous Fenton process is a promising method that can allow operation at neutral pH and ambient temperature without the need to neutralize the effluent after the treatment (Navalon et al., 2010). This is because the heterogeneous catalysts are easier to separate from the effluent and they are noncorrosive and environmentally-benign (Soon and Hameed, 2011). The efficiency and performance of Fenton-type
667
processes, especially the heterogeneous Fenton processes for the destruction of pharmaceutical compounds, as an important category of emerging contaminants, and their removal from water have not been extensively reviewed. Therefore, this review addresses recent developments in the application of Fenton-type processes and the controlling parameters for PhACs degradation, as well as changes in the toxicity and biodegradability of treated water. Finally, the application of Fenton type processes as a pre-treatment step before conventional biological processes at pilot-scale operations for the removal of PhACs is considered.
2. Mechanism of Fenton-type reactions 2.1. Homogeneous Fenton processes The Fenton reaction is based on the generation of highly reactive hydroxyl radicals by the combination of peroxides (mainly hydrogen peroxide) and ferrous ions in acidic condition according to Eq. (1) (Kavitha and Palanivelu, 2004; Escher et al., 2006). Hydroxyl radicals are strong species that can remove one electron from any substance present in the solution to form hydroxide anion. Hydroxyl radicals can also gain a hydrogen atom from hydrocarbons to compensate for their missing atom. Considering the bond energy equal to 109 kcal/mol of OeH bond, any other bond such as CeH in organic compounds with a bond energy lower than this value is thermodynamically susceptible to oxidation (Navalon et al., 2010). At the first stage of reaction, the degradation of contaminants is very fast. The rapid decomposition of pollutants is attributed to the rapid production of hydroxyl radicals, owing to the presence of high Fe2þ species in the solution based on Eq. (1). At the second stage, due to the consumption of Fe2þ ions and generation of Fe3þ, the rate of reaction will decrease. This is because the reaction of Fe3þ and H2O2 leads to the production of HO2 radicals (E ¼ 1.65 V) which are weaker oxidants compared to the OH radicals (E ¼ 2.80 V) and have a lower rate of production (Eq. (2)) (de Luna et al., 2013). The generation of hydroxyl radicals and degradation efficiencies of contaminants are enhanced by the irradiation of UV light in the photo-Fenton process. The photo-Fenton process is based on the production of hydroxyl radicals by catalytic decomposition of hydrogen peroxide in acidic media and under irradiation. In order to prevent the precipitation of ferrous ion, this process is conducted at a pH of around 3. In the photo-Fenton reaction, the produced Fe3þ ions act as light-absorbing species to generate another hydroxyl radical based on Eqs. (3) and (4). In Fenton and photo-Fenton processes, the iron sludge formed after the reaction should be removed. However, in the photo-Fenton process, the required amount of ferrous catalyst and the volume of generated sludge are considerably lower than those in the Fenton process (Hermosilla et al., 2009). In addition, exploiting UV or solar light in the photoFenton process has important effects on the disinfection of treated water, through the inactivation of microorganisms (Rahim Pouran et al., 2015). The following reactions (Eqs. (1)e(8)) may take place in Fenton and Photo-Fenton processes. A list of complete reactions which may contribute to the Fenton and Photo-Fenton processes can be found elsewhere (Kusi c et al., 2006a).
Fe2þ þ H2 O2 /Fe3þ þ OH þ OH
(1)
Fe3þ þ H2 O2 /Fe2þ þ HO2 þ Hþ
(2)
Fe3þ þ H2 O þ hy/Fe2þ þ OH þ Hþ
(3)
668
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
Table 1 Homogeneous Fenton process for the removal of PhACs. [PhAC]/Initial concentration
Type of catalyst
Experimental conditions
Remarks and findings
[Sulfathiazole]0 ¼ 47 mmol/L
FeSO4$7H2O (Sigma eAldrich)
[Amoxicillin]0 ¼ 1000 mg/L
FeSO4$7H2O (Sinopharm Chemical Reagent)
squez et al., 84% PhAC removal was observed after 8 min reaction. (Vela 2014) 30% TOC removal was achieved after 60 min reaction. 40% COD removal was reported after 15 min reaction. Oxamic acid as a byproduct remained stable after long periods of reaction. (Guo et al., 2015) 80% removal of AMX was achieved. The Fenton reaction successfully applied as a pretreatment process for the conventional activated sludge process.
[Acetaminophen]0 ¼ 5 mM
FeSO4.H2O (Merck)
[Amoxicillin]0 ¼ 104 mg/L And [Ampicillin]0 ¼ 105 mg/L And [Cloxacillin]0 ¼ 103 mg/L
FeSO4$7H2O (Essex)
Distilled water pH 3 [Fe2þ] ¼ 192 mmol/L [H2O2] ¼ 1856 mmol/L Reaction time ¼ 60 min Distilled water pH 4 [Fe2þ ]/[H2O2] ¼ 20 Temperature ¼ 40 C Reaction time ¼ 70 min Distilled water pH 3 [Fe2þ] ¼ 0.5 mM [H2O2] ¼ 25 mM [Fe2þ ]/[H2O2] ¼ 0.02 Reaction time ¼ 20 min Distilled water pH 3 [H2O2]:[COD] ¼ 3:1(MR) [H2O2]:[Fe2þ] ¼ 10:1 (MR) [COD]:[H2O2]:[Fe2þ] ¼ 1:3:0.3 (MR) Reaction time ¼ 50 min
Drug manufacturing sewage [COD]0 ¼ 900e7000 mg/L
FeSO4$7H2O (Merck)
[Dipyrone]0 ¼ 50 mg/L
Modelled effluent pH 3.5 Room temperature [H2O2]:[Fe2þ] ¼ 155 (MR) [H2O2] ¼ 0.3 M [Fe2þ] ¼ 2 mM Reaction time ¼ 30 min (Fe(NH4)2(SO4)2$6H2O) Distilled water (Merck) Temperature ¼ 25 C pH 3.5 [H2O2]0 ¼ 22.5 mM [Fe2þ ]0 ¼ 2.25 mM Reaction time ¼ 50 min
Drug manufacturing sewage [COD]0 ¼ 4100 mg/L
FeSO4$7H2O (Merck)
Drug manufacturing effluent, pH 3 [COD ]0/[H2O2]0 ¼ 1:2.2 [Fe2þ ]0/[H2O2]0 ¼ 1:50 Reaction time ¼ 90 min
[Etodolac]0 ¼ 511 mg/L [COD]0 ¼ 18,000 mg/L
FeSO4$7H2O (Merck)
[Metronidazole]0 ¼ 1 mg/L
(FeSO4$7H2O) (Fisher Scientific)
Pharmaceutical wastewater Temperature ¼ 25 C pH 3 [H2O2]0 ¼ 1 M [Fe2þ ]0 ¼ 0.05 M Reaction time ¼ 30 min Deionized water pH 3.5 [H2O2]0 ¼ 1 mg/L [Fe2þ ]0 ¼ 11.76 mM Reaction time ¼ 5 min Distilled water pH 3.52 Temperature ¼ 25 C [H2O2]0 ¼ 1.39 104 mol/L [Fe2þ ]0 ¼ 1.25 105 mol/L [Fe3þ ]0 ¼ 1.68 105 mol/L Reaction time ¼ 120 min
[Carbamazepine]0 ¼ 4.98 mg/L (FeSO4$7H2O) and (Fe2(SO4)3.7H2O) (Merck)
Fe3þ þ H2 O2 þ hy/Fe2þ þ HO2 þ Hþ
(4)
OH þ H2 O2 /H2 O þ HO2
(5)
The photo-Fenton process has been shown to be a cost-effective advanced oxidation process. However, the associated costs may be reduced substantially when solar energy is used instead of artificial
Ref.
99.6% removal of AAP was achieved. (de Luna et al., Second order rate constant was 1.68 mM min1. 2013) Increasing initial concentration of reagent was shown a positive effect on AAP removal.
Complete antibiotics degradation was achieved after (Elmolla and Chaudhuri, 2010) 2 min of reaction. 80% COD removal was achieved after 50 min of reaction. 53% DOC removal was achieved after 50 min of reaction. Improvement in biodegradability was shown after treatment. First order rate constant of PhACs degradation was 0.0144 min1 and t1/2 was 69.3 min. No significant differences were shown at room (Tekin et al., 2006) temperature and 50 C. 52% COD removal was reported after treatment. About 85% BOD5 removal was achieved. Improvement in biodegradability and reduction in toxicity were shown after the treatment. 94.1% PhAC removal was achieved after 45 min treatment. 73.5% PhAC removal was shown after 2.5 min of oxidation. 42.78% TOC removal was obtained in 5 min. Biodegradability (BOD5/COD) was improved from ~0.1 to 0.62 in 10 min. 87% COD removal was achieved after 1.5 h. BOD5/COD was improved from 0.26 to 0.5. Fenton was very effective for pre-treatment of this type of wastewater. 97.5 and 99.6% removal of paracetamol and diclofenac were observed, respectively. 99.9% PhAC removal was achieved after treatment. 82% COD removal was obtained after reaction. Fenton process was an effective pre-treatment method to NF process.
No degradation of PhAC was observed without ferrous ions. 76% metronidazole removal was obtained after 5 min. First order rate constant of PhAC degradation was 0.066 min1 and t1/2 was 1 min. H2O2 initial concentration was the most influential of the studied parameters, followed by Fe2þ concentration, pH, and, Fe3þ ion concentration. Total degradation of CPZ was achieved at optimum conditions. In this integrated process, Fenton’s reaction was more important than Fenton-like reaction.
(Giri and Golder, 2014)
(Badawy et al., 2009)
(Vergili and Gencdal, 2015)
(Shemer et al., 2006)
(Domínguez et al., 2012)
light (Elmolla and Chaudhuri, 2010). The homogeneous Fenton reactions are widely used in water treatment (Hermosilla et al., 2009). However, the limitations of homogeneous reactions are mainly related to the high concentration (50e80 ppm) of Fenþ ions needed for effective removal of contaminants, which is significantly higher than the acceptable concentration in the effluent of water treatment (about 2 ppm) to be discharged into the environment
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
(Guo et al., 2015). Therefore, additional treatment is necessary to reduce the concentration of Fenþ ions in the effluent. Moreover, the treatment of iron-containing sludge which has been accumulated at the end of process is expensive and usually requires a large amount of chemicals (Bobu et al., 2008; Rache et al., 2014). Tables 1 and 2 provide an overview of the PhACs degraded by homogeneous Fenton and photo-Fenton processes, respectively, and summarize the mineralization efficiency of these contaminants under optimum conditions. 2.2. Heterogeneous Fenton processes As stated before, the homogeneous Fenton process has been widely used because of its simplicity, possibility of using conventional equipment and operation at ambient pressures and temperatures. Nevertheless, due to certain limitations such as high iron loss to the environment, need for iron recovery before the discharge of effluents to the receiving waters to comply with the standards (about 2 ppm), high H2O2 consumption and high operation cost, the development of heterogeneous Fenton process as an alternative to the homogeneous Fenton process has attracted much attention (Hermosilla et al., 2009). In addition, homogeneous Fenton processes have limitations such as the narrow range of operating pH and the formation of different Fe3þ complexes by changing the pH of solution (Garrido-Ramírez et al., 2010). As mentioned before, homogeneous Fenton-based processes proceed optimally at acidic conditions. Therefore, subsequent neutralization and the resulting precipitation of Fe(OH)3 as sludge will require additional handling and treatment (Kuan et al., 2015). A further limitation of the use of homogeneous Fenton process is the complexation of iron ions by the substances present in real effluents such as EDTA, or the generation of possible by-products such as oxalic acid (Hartmann et al., 2010). Finally, in the homogeneous Fenton process, catalyst deactivation may occur by iron complexing agents such as phosphate anions (Kuznetsova et al., 2004) (see in Tables 3 and 4). Using a heterogeneous Fenton-like process offers advantages compared to the homogeneous Fenton reactions. The heterogeneous Fenton reactions aim to expand the feasible operating pH range and reduce the problems associated with the separation of high dosage of iron ions that remain after the treatment (ArzateSalgado et al., 2016). An example of a simple Fenton-like solid catalyst is magnetite (Fe3O4), which can be easily removed from the solution through its magnetism (Kuan et al., 2015). Therefore, recent researches have focused on the development of stable heterogeneous catalysts in an effort to minimize leaching, while increasing catalytic activities and long-term stability (Hermosilla et al., 2009; Soon and Hameed, 2011). Improvements in the heterogeneous catalytic activity not only enhance the feasibility of Fenton process for wastewater treatment, but they also eliminate the need to pre-adjust the solution’s pH or to neutralize it at the end of the process (Bobu et al., 2008). However, the presence of a small fraction of iron on the surface of catalysts in the heterogeneous Fenton processes results in slow reaction kinetics compared to the homogeneous Fenton processes (Hermosilla et al., 2009; Bae et al., 2013). While the homogeneous Fenton processes depend only on chemical reactions between the reagents and contaminants, in the heterogeneous Fenton reactions physical processes on the surface of catalysts in addition to chemical reactions control the overall outcome of the process. The physical processes include surface reactions and desorption while the chemical reactions refer to those between contaminants and the produced radicals (Soon and Hameed, 2011). In the heterogeneous Fenton reaction three possible mechanisms can occur: (i) chemisorption of
669
pharmaceuticals on the surface of catalysts; (ii) homogeneous transformation of hydrogen peroxide to hydroxyl radicals by iron leaching to the reaction solution and/or (iii) reaction of hydrogen peroxide with the iron species on the surface of catalysts and its decomposition to hydroxyl radicals (Rahim Pouran et al., 2015). Therefore, potential mass transfer limitations affect the resulting reaction rates (Kuan et al., 2015). In the heterogeneous Fenton process, the catalyst characteristics such as surface area, pore volume, density, porosity, pore size and distribution need to be taken into consideration. Hence, the catalyst surface can be designed or modified to achieve higher catalytic performance such as activity, selectivity and stability (Soon and Hameed, 2011). In heterogeneous Fenton processes, the iron ions can leach from the catalyst during the reaction. This phenomenon not only causes a loss in activity with time, it also produces secondary metal ion contaminants (Hartmann et al., 2010). Currently, most studies on heterogeneous Fenton processes focus on the stability and reaction kinetics rather than the process mechanism such as the leaching of iron ions into the aqueous phase and its effect on the process. The difficulties and complications in identifying pertinent chemical pathways have hindered the identification of relevant mechanisms. Dissolved homogeneous ions can be deactivated and they can precipitate as a solid phase, especially due to pH variations as a result of the generation of intermediate compounds during the reaction (Kuan et al., 2015). Extensive efforts have been made to find an effective solid catalyst for the iron ions to overcome major disadvantages of the homogeneous Fenton process. 2.2.1. Iron oxide minerals Iron oxide minerals such as magnetite (Fe3O4), hematite (Fe2O3), goethite (a-FeOOH), pyrite (FeS2) and lepidocrocite (g-FeOOH), which are common constituents of soil, can serve as potential environmentally-benign material for the elimination of emerging contaminants. Moreover, these minerals are widespread in the environment and can be utilized in in situ removal processes (Sun and Lemley, 2011; Bae et al., 2013; Velichkova et al., 2013). Due to their specific properties, such as stability, reusability and ease of separation compared to other iron minerals and ferrous salts, and their environmental compatibility, magnetite (Fe3O4) has attracted much attention for use in Fenton processes (Sun et al., 2013). Magnetite can be formed naturally via different pathways such as Fe2þ oxidation, iron metal corrosion and chemical and biological reduction of Fe3þ ions. Magnetite can oxidize organic compounds into less hazardous or more biodegradable forms. However, total mineralization can be achieved only by Fenton-like processes in the presence of hydrogen peroxide (Sun and Lemley, 2011). In addition, Fe3O4 nanoparticles have a tendency to agglomerate due to the intrinsic magnetic interactions and van der Waals forces. Agglomeration can reduce the surface/volume ratio and catalytic activities (Jaafarzadeh et al., 2015). It seems that pyrite (FeS2) is a very effective heterogeneous catalyst in the Fenton process. Using pyrite has two positive effects since it provides aqueous Fe2þ ions, while decreasing the pH of solution to the proper level (normally pH 3e4) (Bae et al., 2013). The use of immobilized iron on support material can also be considered as an effective method to overcome the abovementioned limitations of homo/heterogonous Fenton type processes (Sun and Lemley, 2011; Sun et al., 2013). 2.2.2. Support material in heterogeneous Fenton processes Supported metal (mostly iron) compounds and iron-coated particles show potential to overcome the problems associated with homogenous Fenton reactions such as the limited operational range of pH, release of iron into the effluent, removal of the iron
670
Table 2 Homogeneous photo-Fenton process for the removal of PhACs. Type of catalyst
Experimental conditions
Source of energy
[Metoprolol]0 ¼ 50 mg/L
FeSO4$7H2O (Panreac PA)
BLB (3 8 W, lmax ¼ 365 nm) or solar Deionized water pH 3 irradiation or artificial solar irradiation Temperature ¼ 25 C (Xe lamp 1 kW) or UVC lamp [Fe2þ] ¼ 10 mg/L [H2O2] ¼ 150 mg/L Reaction time ¼ 150 min
[ATL]0 ¼ 20 mg/L and [Metoprolol]0 ¼ 20 mg/L
Fe(ClO4)3 (Sigma eAldrich)
[Diclofenac]0 ¼ 50 mg/L
Fe2 (SO4)3 H2O
[Sulfathiazole]0 ¼ 47 mmol/L
FeSO4$7H2O (Sigma eAldrich)
[Amoxicillin]0 ¼ 104 mg/L, [Cloxacillin]0 ¼ 105 mg/L, [Ampicillin]0 ¼ 103 mg/L
FeSO4$7H2O (Essex)
[Sulfonamides]0 ¼ 200 mg/L
FeSO4$7H2O (Panreac PA)
[Sulfonamides]0 ¼ 200 mg/L
FeSO4$7H2O (Panreac PA)
Distilled water, pH about 2.9 Temperature ¼ 30 e35 C [iron ions] ¼ 2.8 e5 mg/L [H2O2] ¼ 95 e100 mg/L Reaction time ¼ 180 min Distilled water pH 7 (with no buffer capacity) Temperature ¼ 30 e40 C [H2O2] ¼ 15 mM [Catalyst] ¼ 0.05 mM Reaction time ¼ 110 min Distilled water, pH 3 [Fe2þ] ¼ 157 mmol/L [H2O2] ¼ 1219 mmol/ L Reaction time ¼ 60 min Distilled water, pH 3 [H2O2]/[Fe2þ ] ¼ 20 Temperature ¼ 22 C [COD]0 ¼ 520 mg/L [H2O2]/[COD] ¼ 1.5 Reaction time ¼ 50 min Distilled water, pH 2.8 Temperature ¼ 23 e31 C (not controlled) [Fe2þ] ¼ 10 mg/L [H2O2] ¼ 400 mg/L Reaction time ¼ 100 min Distilled water, pH 2.8
UV 125 W HP, Hg lamp (lmax ¼ 290 nm)
Remarks and findings
Ref.
Photolysis alone, hydrogen peroxide alone and photolysis with iron were not able to eliminate MET. 100% MET removal after 7 min and 81.2% TOC removal after 90 min were achieved by BLB lamp. By using artificial solar 97.3% of MET elimination after 7 min and 78.8% of TOC conversion after 120 min were observed. With UVC, 100% of MET elimination after 20 min and 17.6% of TOC conversion after 60 min were reported. 100% removal of PhACs was achieved after 1 min. Total mineralization was taken about 150 min. Conduction experiment in river and lake water was caused a delay in degradation and mineralization and increased H2O2 consumption. First order rate constant of ATL degradation was 0.024 min1, and t1/2 was 28.4 min. First order rate constant of metoprolol degradation was 0.048 min1, and t1/2 was 14.3 min.
(Romero et al., 2016b)
(Veloutsou et al., 2014)
Solar irradiation
rez-Estrada After 50 min of reaction, 100% removal of PhAC was achieved. (Pe About 98% DOC removal was also achieved after 110 min of reaction. et al., 2005b) The H2O2 amount needed for total oxidation was about 8 times more than theoretical stoichiometric demand.
6 20 W solarium lamps (365 nm) 3.5 mW cm2
UV Lamp 6W (365 nm)
Complete degradation of amoxicillin, cloxacillin and ampicillin were (Elmolla and achieved in 2 min. Chaudhuri, 2009) 58.4% DOC removal was achieved in 50 min. 80.8% COD removal was achieved in 50 min. Biodegradability (BOD5/COD) was improved from 0 to 0.4 after 50 min which is considered adequate for biological treatment.
LP UV lamp Hg vapor 88W (l ¼ 350e400 nm)
95% removal was achieved after 60 min reaction. 75% TOC removal was reported after 60 min reaction. 90% COD removal was achieved after 15 min reaction. Photo-Fenton reaction was twice faster than Fenton reaction.
Complete degradation of PhAC was achieved in 50 min. 56.4% TOC removal was shown after 94 min. 82.2% COD removal was achieved after 94 min. The energy consumption was estimated 2.5 kJ/L in this research.
87% mineralization was achieved. 96% COD removal was achieved after 94 min.
(Vel asquez et al., 2014)
(Gonzalez et al., 2009)
(Gonzalez et al., 2009)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
[PhAC]/Initial concentration
[Amoxicillin]0 ¼ 138 mg/L [Cloxacillin]0 ¼ 84 mg/L [COD]0 ¼ 670 mg/L [DOC]0 ¼ 145 mg/L [BOD5]0 ¼ 70 mg/L
FeSO4$7H2O (Essex)
[Diclofenac]0 ¼ 0.12 mM
FeCl3.6H2O (Fluka)
Effluent from antibiotic industry pH 3 [H2O2]/[COD] ¼ 2.5 [H2O2]/[Fe2þ] ¼ 20 [H2O2] ¼ 44.9 mM Reaction time ¼ 30 min Distilled water pH 2.8 Temperature ¼ 50 C [Fe ions] ¼ 14 mg/L [H2O2] ¼ 340 mg/L Reaction time ¼ 60 min Distilled water or effluent from STP
Solar radiation (300 < l < 500 nm) z121.6 W/m2
The energy consumption was estimated 19 kJ/L.
Solar radiation
UV lamps 10 W (lmax ¼ 254 nm)
Total degradation of paracetamol, amoxicillin, ampicillin and diclofenac were achieved after 60, 90, 120, and 120 min of irradiation, respectively. First order rate constant of paracetamol, amoxicillin, ampicillin and diclofenac degradation was 0.070 min1, 0.056 min1, 0.042 min1 and 0.51 min1, respectively. The biodegradability (BOD5/COD) was enhanced from 0.3 to 0.52. The removal of BOD5, TOC and COD were reported at 61%, 52% and 77%, respectively. Photo-Fenton was an efficient pre-treatment process in order to improve biodegradability. TOC, COD and BOD5 removal were decreased dramatically with increasing initial COD concentration.
15 W Blacklight lamp UVA (320e400 nm) 19 W m-2
The degradation was observed only when three key components of photo- (Batista and Nogueira, 2012) Fenton were present (Light, Fe and H2O2). 92 and 90% mineralization was achieved after 42 min of reaction in the case of sulfadiazine and sulfathiazole, respectively.
UV lamp 9 W (362 nm) 12 W m-2
LP UV (254 nm) 1.95 mW cm-2 & MP UV (200e400 nm) 1.9 mW cm-2
6% removal was obtained by direct photolysis with LP while 12% removal was achieved by MP. For both lamps about 60% removal was observed by adding 25 mg/L of H2O2 after 2.5 min. 94% removal was achieved by the photo-Fenton process under optimum conditions. First order rate constant of PhAC degradation was 0.23 min1, and t1/2 was 0.5 min. Complete removal of the antibiotics occurred in 1 min. 59 and 49% soluble COD and DOC removal were achieved after 30 min. Biodegradability (BOD5/COD) was improved to 0.44 after treatment.
UV lamp 6 W (365 nm)
UV 400 W (254 nm)
(Alalm et al., 2015)
(Kajitvichyanukul and Suntronvipart, 2006)
(Giri and Golder, 83.2% drug removal was shown at 2.5 min of oxidation. 2014) 96.4% PhAC removal was achieved after 45 min reaction. In the absences of ferrous ion, the PhAC removal was 74.4% in 45 min. 56.0% TOC removal was obtained in 5min. Biodegradability (BOD5/COD) was improved from ~0.1 to 1.51 after 10 min of reaction. (Shemer et al., 2006)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
Temperature ¼ 26 C [Fe2þ] ¼ 10 mg/L [H2O2] ¼ 550 mg/L Reaction time ¼ 100 min Distilled water, pH 3 [Paracetamol]0 ¼ [Amoxicillin]0 ¼ FeSO4$7H2O (Sigma [Ampicillin]0 ¼ [Diclofenac]0 ¼ eAldrich) [Fe2þ] ¼ 500 mg/L 100 mg/L [H2O2] ¼ 1500 mg/L Reaction time ¼ 120 min FeSO4$7H2O Hospital effluent Hospital effluent [TSS] ¼ 115 (mg/L); [COD] ¼ 1350 (Merck) pH 3 (mg/L); [Fe2þ] ¼ 135 mg/L [BOD5] ¼ 410 (mg/L); [COD]:[ Fe2þ] ¼ 1:0.1 [TOC] ¼ 1050 (mg/L) [COD]:[H2O2] ¼ 1:4 Temperature ¼ 25 e31 C Reaction time ¼ 120 min [Sulfadiazine]0 ¼ 25 mg/L, Prepared Distilled water, pH [Sulfathiazole]0 ¼ 25 mg/L (K3Fe(C2O4)3$3H2O) 2.5 [Fe3þoxalalate] ¼ 0.2 mM [H2O2] ¼ 5 mmol/L Reaction time ¼ 45 min [Dipyrone]0 ¼ 50 mg/L (Fe(NH4)2(SO4)2$6H2O) Distilled water, (Merck) Temperature ¼ 25 C pH 3.5 [H2O2]0 ¼ 22.5 mM [Fe2þ ]0 ¼ 2.25 mM Reaction time ¼ 45 min [Metronidazole]0 ¼ 1 mg/L FeSO4$7H2O (Fisher Deionized water Scientific) pH 3.5 [H2O2]0 ¼ 1 mg/L [Fe2þ ]0 ¼ 11.76 mM Reaction time ¼ 5 min
(Elmolla and Chaudhuri, 2011)
A direct proportional between the DCF and TOC removal and UV light (Ravina et al., 2002) intensity was reported. Total DCF removal was achieved after only 0.5 min while about 50 min was required for total TOC removal.
et al., (Trovo 2008) 671
(continued on next page)
672
Table 2 (continued ) [PhAC]/Initial concentration
Type of catalyst
[Amoxicillin]0 ¼ 42 mg/L, or [Bezafibrate]0 ¼ 13 mg/L, or [Paracetamol]0 ¼ 20 mg/L
(K3Fe(C2O4)3$3H2O) or pH 2.5 [catalyst] ¼ 0.2 mM Fe(NO3)3 [H2O2] ¼ 5.0 mM (Mallinckrodt) Reaction time ¼ 30 min
[Tetracycline]0 ¼ 24 mg/L
(K3Fe(C2O4)3$3H2O) Mallinckrodt or Fe(NO3)3
[Diclofenac]0 ¼ 50 mg/L
[Paracetamol]0 ¼ 50 mg/L
[Ofloxacin]0 ¼ 10 mg/L
Drug formulation effluent [COD]0 ¼ 600 mg/L
Distilled water or Effluent from STP pH 2.5 [catalyst] ¼ 0.2 mM [H2O2] ¼ 10 mM Reaction time ¼ 60 min FeSO4$7H2O Distilled water pH 7 Temperature ¼ 30 e40 C [H2O2] ¼ 15 mM [Catalyst] ¼ 0.05 mM Reaction time ¼ 110 min FeSO4$7H2O (Panreac Distilled water PA) pH 2.8 Temperature ¼ 25 C [H2O2] ¼ 300 mg/L [Catalyst] ¼ 10 mg/L FeSO4$7H2O (POCH SA) Distilled water or (K3Fe(C2O4)3$3H2O) pH 2.5e2.8 Temperature ¼ 25 C (Mallinkrodt) [H2O2] ¼ 120 mg/L [Catalyst] ¼ 0.05 mM Reaction time ¼ 300 min FeSO4$7H2O WWTP effluent (Riedel-de Haen) (secondary treated) pH 3 Temperature ¼ 25 C [H2O2] ¼ 2.714 mM [Catalyst] ¼ 5 mg/L Reaction time ¼ 120 min Fe(NO3)3$9H2O
Drug formulation effluent pH 3 [H2O2] ¼ 25 mM [Catalyst] ¼ 1.5 mM Reaction time ¼ 30 min
Source of energy
Remarks and findings
Ref.
Black-light lamp (15 W fluorescent (for 98% removal of BZF and PCT was achieved after 5 min of reaction using FeOx; AMX and PCT), lmax ¼ 365, 19 W m2) whereas, by using Fe(NO3)3, 89 and 53% removal was achieved respectively & solar light (20 W m2 for BZF) at the same time. Total degradation of AMX was achieved after 0.5 min of reaction using both catalysts. TOC removal rate was higher when using FeOx for all considered compounds. First order rate constant of BZF and PCT, was 0.94 and 0.70 min1, respectively at optimum conditions. Black-light lamp (15 W fluorescent, Total removal of TC was achieved under black-light by using Fe(NO3)3 after (Bautitz and 1 min. Also, 72% TOC removal was obtained after 60 min reaction, while by Nogueira, 2007) lmax ¼ 365, 19 W m2) or solar light using FeOx, more reaction time (8 min) was needed for 100% removal of TC. (15e20 W. m2) Under solar irradiation, 100% removal was achieved by FeOx in 0.5 min while, Fe(NO3)3 needed 3 min for total removal in the same condition.
rez-Estrada (Pe et al., 2005a)
Solar irradiation
Total DCF removal was achieved after 60 min exposure to sunlight. 100% DOC removal (as total mineralization) was shown after 100 min. About 18 intermediates were identified during the mineralization.
UV lamp lmax ¼ 365 3 (8 W)
lez et al., (Gonza About total removal was reported at the end of the reaction. (BOD5/COD) was improved with increasing H2O2 dose and reached to the 2007) 0.25 by using 400 mg/L H2O2. The effluent was less toxic after the treatment.
Solar radiation simulator (1100 W xenon arc lamp)
et al., No photolysis of PCT was observedat natural pH (4.2) after 300 min (Trovo irradiation. 2012) 100% PCT removal was achieved after 120 min by FeSO4 and 180 min by FeOx. 79% TOC removal was reported by FeSO4 and 58% by FeOx after 300 min.
Solar radiation simulator (1 kW Xenon lamp)
Less than 6% OFX removal was observed after 120 min by photolysis. Total removal of OFX was achieved after 90 min by Fenton reaction while by photo-Fenton reaction 30 min was required. Photo-Fenton was more efficient than TiO2 not only for the OFX degradation but also for the DOC removal. At optimum conditions, 50% and 10% DOC removal were achieved by PhotoFenton and TiO2, respectively. First order rate constant of OFX degradation by the photo-Fenton process was 0.1128 min1, and t1/2 was 6.14 min. 56 and 44% COD removals were achieved by photo-Fenton and Fenton process, respectively. The BOD5/COD ratio was increased from 0.10 to 0.45 and 0.24 after application of the photo-Fenton and Fenton, respectively. 35 and 42% TOC removal were reported by Fenton and phot-Fenton processes respectively.
UV-A 125 W (365 nm)
(Michael et al., 2010)
(Arslan-Alaton and Gurses, 2004)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
[Sulfamethoxazole]0 ¼ 200 mg/L
Experimental conditions
[Metoprolol]0 ¼ 50 mg/L and [Resorcinol]0 ¼ 50 mg/L
[Sulfamethoxazole]0 ¼ 50 mg/L
[Ibuprofen]0 ¼ 0.87 mM
FeSO4.7H2O (Panreac)
Deionized water pH 6.2 (natural) Temperature ¼ 25 C [H2O2] ¼ 150 mg/L [Catalyst] ¼ 10 mg/L Reaction time ¼ 180 min FeSO4$7H2O (Sigma Distilled water eAldrich) Or Sea water pH 2.5e2.8 Temperature ¼ 25 C [H2O2] ¼ 210 mg/L [Catalyst] ¼ 5.2 mg/L Reaction time ¼ 7 h FeSO4$7H2O (Panreac) Distilled water pH 3 Temperature ¼ 30 C [H2O2] ¼ 0.32 mM [Catalyst] ¼ 1.2 mM Reaction time ¼ 120 min FeSO4$7H2O (POCH SA) or (K3Fe(C2O4)3$3H2O) (Mallinkrodt)
Combination of six emerging contaminants
Fe(ClO4)3$H2O (Sigma eAldrich)
Pharmaceutical laboratory effluent FeSO4$7H2O Merck
Distilled water pH 2.5e2.8 Temperature ¼ 25 C [H2O2] ¼ 120 mg/L [Catalyst] ¼ 0.05 mM Reaction time ¼ 240 min Secondary effluent from a MWTP pH 3 [H2O2]/[Fe3þ] ¼ 6.1 (mass ratio) [Catalyst] ¼ 2.8 mg/L Reaction time ¼ 5 h Diluted waste water Ambient temperature [TOC]0 ¼ 400 ppm pH 2.7 [H2O2] ¼ 2500 ppm [Catalyst] ¼ 20 ppm Reaction time ¼ 120 min
Solar radiation simulator (1100 W xenon arc lamp)
Xe lamp 1 kW 6.9 mE s1 (290e400 nm)
Solar radiation simulator (1100 W xenon arc lamp)
Solar radiation
Solar radiation
Adding resorcinol (ligand iron complex) to the solution let to perform the reaction in neutral pH condition. By Fenton reaction, 100 and 92% removal were achieved for MET and RES after 20 min respectively. By photo-Fenton reaction, 100 and 94.4% removal were achieved for MET and RES after 3 min respectively. Biodegradability increased at the end of the processes. No hydrolysis was reported at the pH studied (2.5, 4.8 and 9.0) after 48 h. After the process, the toxicity was increased from 16 to 86% in seawater. Total degradation of SMX was achieved after 16 min in distilled water and 105 min in seawater.
UV/H2O2 process led to the 30% IBP removal after 2 h, however, TOC did not decrease. Maximum 60% and 10% IBP and TOC removal were achieved by Fenton process, About 100% IBP (after 60 min) and 40% TOC removal were reported by the photo-Fenton reaction. Biodegradability was improved at the end of the process (BOD5 was reached to less than 1 mg/L from 25 mg/L). No AMX removal by photolysis was observed at natural pH (6.2) after 6 h irradiation. 100% removal of AMX was achieved after 5 min by FeOx and 15 min by FeSO4. After 240 min, between 73 and 81% TOC removal was reported. 73 and 81% TOC removal was obtained in the presence of FeSO4 and FeOx, respectively. Total ECs removal were achieved in less than 90 min. At the optimum point, 2.5-fold increase of the BOD5/COD ratio was shown after treatment. About 35% TOC removal was obtained after 5 h treatment.
(Romero et al., 2016a)
et al., (Trovo 2009)
ndez-Arriaga (Me et al., 2010)
et al., (Trovo 2011)
~ ones et al., (Quin 2015)
sito et al., (Expo About 80% TOC removal was achieved after treatment. Some acids like acetate, formate and oxalate were formed during the 2016) oxidation.
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
[Amoxicillin]0 ¼ 50 mg/L
BLB 8W (365 nm)
673
674
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
sludge after the treatment, and slow reaction kinetics of the heterogeneous Fenton processes (Hu et al., 2011). Therefore, increasing attention has been paid to the development of support medium for the heterogeneous Fenton catalysts such as pillared clays (Feng et al., 2006; Li et al., 2006, 2015; Bobu et al., 2008), activated carbon (Georgi and Kopinke, 2005; Dantas et al., 2006; Zazo et al., 2006; Jaafarzadeh et al., 2015; Lan et al., 2015), alumina (Bautista et al., 2010a; Kunde and Yadav, 2015) and zeolite (Kusi c et al., 2006b; Kasiri et al., 2008; Tekbas¸ et al., 2008; Aleksi c et al., 2010; Perisic et al., 2016). Using mesoporous material as the catalyst support in Fenton or Fenton-like processes is a promising approach due to the synergistic effect of oxidation and adsorption (Song et al., 2015). In addition, the incorporation of iron oxide nanoparticles into porous support materials such as activated carbon reduces the strong tendency of agglomeration of nanoparticles due to intraparticle interactions. Moreover, immobilization of metal catalysts into porous adsorbents reduces the risk of iron release in the treated effluent (Jonidi Jafari et al., 2016). Zhang et al. (2016b) used mesoporous manganese oxide with a high surface area as a support material for copper in Fenton-like processes for the removal of BZA at neutral pH. A simultaneous adsorption on activated carbon along with oxidation by Fenton reaction of TC is also reported (Jonidi Jafari et al., 2016). Activated carbon can be considered as a promising support for the heterogeneous Fenton reaction owing to its high surface area, wide availability, high stability in acidic or alkaline environments and low cost (Jaafarzadeh et al., 2015; Lan et al., 2015). Exploiting carbon materials as the support in heterogeneous Fenton processes is an innovative approach because it provides simultaneous adsorption and catalytic oxidation of soluble contaminants (Dantas et al., 2006). Whether the adsorption process is beneficial or not is not fully clear. Most investigators state that the adsorption offers advantages and will improve the efficiency of oxidation (Ramirez et al., 2007). Nevertheless, Georgi and Kopinke (2005) considered adsorption as a disadvantage because they claimed that the contaminants adsorbed on the activated carbon were nearly nonreactive and could not be degraded by hydroxyl radicals. It should be noted that the catalytic activity of activated carbon even without using iron for the decomposition of H2O2 to hydroxyl radicals has been demonstrated (Ramirez et al., 2007). The catalytic activity of activated carbon is related to its surface chemistry. However, the oxidation of contaminants by the hydroxyl radicals that are produced by bare activated carbon in the presence of H2O2 needs a long reaction time to achieve a sufficient extent of removal. Therefore, despite several advantages of bare activated carbon as a Fenton-like catalyst, such as the lack of iron leachate or lower price, using iron metal in carbon structure is crucial when shorter operation times are needed (Ramirez et al., 2007). In another work, porous sea buckthorn was utilized as the support material for (bFeOOH) nanoparticles for the removal of doxycycline antibiotic at neutral pH condition (Zhang et al., 2016a). Porous sea buckthorn is a cost-free agricultural waste residue and can serve as porous support in Fenton processes to enhance the removal efficiency, owing to the synergistic effect of adsorption and Fenton-like reactions. Besides, Fenton processes can serve as an in situ regeneration step for bioadsorbents, since the generated hydroxyl radicals can effectively oxidize accumulated organic contaminants on the adsorbents (Song et al., 2015; Zhang et al., 2016a). Zeolites are hydrated aluminosilicate materials with external and internal surface areas up to several hundred square meters per gram, high cation exchange capacity and cage-like structures. Zeolites in natural or synthesized forms are used in the industry as the catalyst support, adsorbent and ion exchange resins (Tekbas¸ et al., 2008). Due to their high surface area and porosity, zeolites can act as an adsorbent in order to remove contaminants from aquatic
media (Kasiri et al., 2008). Therefore, by using zeolites as a support medium in the heterogeneous photo-Fenton processes, the adsorption process should also be taken into account besides homo/heterogeneous Fenton reaction and direct photolysis of H2O2 and contaminants (Kusic et al., 2006b). In addition, the presence of a strong electrostatic field inside the pores of the zeolite slows down the leaching of iron from Fe-zeolite catalyst during the reaction (Kasiri et al., 2008). Moreover, the preparation of Fe-zeolite catalyst is relatively simple due to the zeolite’s cation exchange capability (Kusi c et al., 2006b). Pillared clays, also known as pillared interlayered clays (PILCs) or cross-linked clays, are among the most-widely studied microporous materials that can be used as support media due to their unique characteristics such as inertness, high surface area, abundance, low cost as well as the simplicity of the pillaring process (Li et al., 2006, 2015). These materials are prepared by exchanging the interlayer cations of layered clays with inorganic polyoxocations (Bobu et al., 2008). Mixed metal oxide pillared clays such as FeeAl or FeeCr may improve the catalyst performance by increasing the surface area or by modifying surface properties such as the acidity (Li et al., 2006). Good catalyst activity and stability of Fe-clay catalyst in removing ciprofloxacin by the heterogeneous photoFenton process was reported (Bobu et al., 2008). Similar to pillared clays and zeolite, alumina can also act as a support and adsorbent. Due to its characteristics, including chemical inertness, mesoporous nature, high surface area and thermal stability, alumina is widely used in industrial operations (Kunde and Yadav, 2015). 3. Major controlling parameters in the Fenton reaction 3.1. Effect of reagents concentration The maintenance of proper concentrations of hydrogen peroxide and iron is important in the homogeneous Fenton process. An excess or deficiency of these reagents can considerably decrease the efficiency of process (Bautitz and Nogueira, 2007). In addition, the use of appropriate concentrations of reagents will prevent extra costs of operation that may result from the use of excess reagents, and will reduce the encountered difficulties in removing the excess iron, according to the effluent standards. In Fenton-based processes, the presence of H2O2 is essential because it is the main source of OH radicals during the process (Bobu et al., 2008). However, due to the radical scavenging effect of H2O2 at high concentrations (Eq. (5) and Eq. (6)), and the auto-decomposition of H2O2 to O2 and H2O (Eq. (7)), the optimum concentration of H2O2 should be determined in Fenton-based processes (Badawy et al., 2006; Li et al., 2006; Domínguez et al., 2012). Moreover, high concentrations of H2O2 may reduce the COD removal efficiency (Elmolla and Chaudhuri, 2009). The consumption of H2O2 is also a critical issue due to its major contribution to the operating cost of Fenton processes (Zazo et al., 2011; Vergili and Gencdal, 2015). It should be noted that the unconverted H2O2 cannot be recovered and has to be eliminated before discharging the final effluent into the water bodies, due to potential toxicity effects (Bautista et al., 2010a). The optimum concentration of hydrogen peroxide depends on the type and concentration of contaminants and Ferrous ion (Elmolla and Chaudhuri, 2010).
HO2 þ OH /H2 O þ O2
(6)
2H2 O2 /2H2 O þ O2
(7)
The use of an adequate concentration of reagents in the Fenton process is important. However, the H2O2/iron molar ratio is the key
Table 3 Heterogeneous Fenton-like process for the removal of PhACs. [PhAC]/Initial concentration
Type of catalyst
Experimental conditions
Source of energy
Remarks and findings
Ref.
[Paracetamol]0 ¼ 100 mg/L
MGN1 (Fe3O4 powder < 50 nm); MGN2 (Fe3O4 powder < 5 nm); MGM (g-Fe2O3 powder < 50 nm) (All from Sigma-Aldrich)
Distilled water, pH 2.6 Temperature ¼ 60 C [Catalyst] ¼ 6 g/L [H2O2] ¼ 153 mM Reaction time ¼ 5 h
Without external energy
(Velichkova et al., 2013)
Combination of 11 detected drugs of abuse [TOC]0 ¼ 60 mg/L
Powder silica-supported iron oxide (Fe2O3/SBA-15)
Raw river water pH 3 Temperature ¼ 22 C [Catalyst] ¼ 0.1e0.6 g/L [H2O2] ¼ 15e60 mg/L Reaction Time ¼ 6 h
UVeVisible 150 W MP mercury lamp (l > 313 nm)
[Paracetamol]0 ¼ 20 mg/L
Synthesized CoFe2O4
UV lamps 4 15 W (lmax ¼ 365 nm)
[Paracetamol]0 ¼ 104 mol/L
Synthesized Goethite (aFeOOH)
Distilled water, pH 3 Temperature ¼ 45 C [Catalyst] ¼ 200 mg/L [H2O2] ¼ 50 mM Reaction time ¼ 60 min Ultra-pure water, pH 3 [H2O2] ¼ 5 103 mol/L [catalyst]0 ¼ 1 g/L Temperature ¼ 20 C Reaction time ¼ 180 min
[Diclofenac]0 ¼ 30 mg/L
Metallurgical slag (60.90% Fe2O3)
Distilled water, pH 7 [H2O2] ¼ 180 mg/L [catalyst]0 ¼ 26.6 mg/L Temperature ¼ 35 C Reaction time ¼ 300 min
Simulated sunlight (300e500 nm) 500 W/m2
[Ofloxacin]0 ¼ 30 mg/L
Synthesized alginate-iron catalyst (4% Iron)
Distilled water, pH 3 [H2O2] ¼ 4.067 mM [catalyst]0 ¼ 400 mg/L Temperature ¼ 25 C Reaction time ¼ 180 min
Without external energy
[Carbamazepine]0 ¼ 15 mg/L or [Ibuprofen]0 ¼ 15 mg/L
Nano-Fe3O4 (Nanostructured & Amorphous MaterialsInc)
Without external energy
[Diclofenac]0 ¼ 20 mg/L
Synthesized Fe-doped CeO2
Distilled water, pH 7.0 [H2O2] ¼ 600 mM [catalyst]0 ¼ 1.84 g/L Temperature ¼ 23 C Reaction time ¼ 12 h Distilled water, pH 5 [H2O2] ¼ 10 mM
Without catalyst, paracetamol was not oxidized by H2O2. For three catalysts, total PhAC degradation was obtained after 5 h oxidation. 43%, 34% and 39% mineralization were achieved with MGN1, MGN2 and MGM, respectively. More than 70% removal of all drugs were achieved using only UVevis light. Using 0.1 g/L of catalyst without UVevis and H2O2 led to strong reduction of the opioid (75e83%), moderate reduction of benzodiazepines (32e56%) however a strong increase of cocaine (225e159%) and a moderate increase of ALCs (51e11%) were observed. The best result (>75% removal of all drugs and 86% of TOC removal) was achieved by a combination of UV and 0.6 g/L catalyst. More than 99% removal was achieved. First order rate constant of EC degradation was 0.0967 min1 The presence of phenol in solution was reduced the efficiency of removal. No detectable adsorption onto Goethite surface was observed. With the 1 mM sensibility detection, Fe2þ leachate was not detected. No degradation was shown after 3 h of direct photolysis. At optimum conditions, the first order reaction rate was 2.33 101 h Complete degradation of PhAC was achieved in 90 min. First order rate constant of PhAC degradation was 0.046 min1. 87% mineralization was achieved after 300 min treatment. 98% EC removal was achieved after treatment. About 10% decrease of support activity was shown after three successive runs. A low (1.2%) iron leaching has been reported. Increasing temperature has a substantial effect on PhAC elimination. The first order rate constant was 0.0356 (min1). 90.0% and 81.1% removal were achieved for CBZ and IBP, respectively. The first-order degradation rate of CBZ and IBP were 0.182 h1 and 0.127, respectively.
Without external energy
Removal efficiency of PhAC with H2O2 alone was only 2.2% in 40 min.
(Rad et al., 2015)
(Mameri et al., 2016)
(Arzate-Salgado et al., 2016)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
LP mercury UV lamps (365 nm) 1840 mW cm2
(Catal a et al., 2015)
(Titouhi and Belgaied, 2016)
(Sun et al., 2013)
(Chong et al., 2016) 675
(continued on next page)
[PhAC]/Initial concentration
676
Table 3 (continued ) Type of catalyst
Experimental conditions
Source of energy
[catalyst]0 ¼ 0.5 g/L Reaction time ¼ 40 min
Synthesized Fe-ZSM5
Distilled water, pH 4 [H2O2] ¼ 50 mM [catalyst]0 ¼ 2 mM Temperature ¼ 25 C Reaction time ¼ 120 min
UVA lamps (lmax ¼ 365 nm) 3.04 107 E s1
[Norfloxacin]0 ¼ 50 mg/L
Synthesized Fe3O4@ALG/Fe
Without external energy
[Ciprofloxacin]0 ¼ 0.15 mM
Synthesized Feclaynanocomposites
[Tetracycline]0 ¼ 20 mg/L
Synthesized nano carbon/Fe3O4
Deionized water, pH 3.5 Temperature ¼ 40 C [Catalyst] ¼ 0.4 g/L [H2O2] ¼ 0.98 mM Reaction time ¼ 6 h Distilled water, pH 3 Temperature ¼ 25 C [catalyst] ¼ 1 g/L [H2O2] ¼ 60 mM Reaction time ¼ 30 min Deionized water, pH 3 Temperature ¼ 25 C [catalyst] ¼ 0.3 g/L [H2O2] ¼ 60 mM Reaction Time ¼ 5 h
[Sulfathiazole]0 ¼ 50 mg/L
Synthesized nano Fe3O4/HA
Deionized water, pH 3.5 Temperature ¼ 40 C [Catalyst] ¼ 3 g/L [H2O2] ¼ 0.39 M Reaction time ¼ 6 h
Without external energy
[Diclofenac]0 ¼ 5 mg/L
Pyrite (Ward’s natural science) or FeSO4$7H2O (Sigma eAldrich)
Deionized water pH 4 Temperature ¼ 25 C [Pyrite] ¼ 0.5 mM [H2O2] ¼ 80 mM Reaction time ¼ 200 s
Without external energy
HP Hg lamp 125 W 9 104 mE m2 s1
Without external energy
85.25% PhAC removal was achieved after 40 min. An excellent chemical stability with negligible leaching ions was reported for this catalyst. Both adsorption (early stage) and oxidation (final stage) were contributed in the DCF and TOC removal. More than 80% and 98.9% of DCF was removed after 2 and 120 min reaction, respectively. 66.4% TOC removal was achieved after 15 min. (BOD5/COD) was improved significantly from 0.043 to 0.63 in 120 min. Catalytic efficiency was increased to 2e4 folds by modifying Fe3O4. 100% of NOF and 90% of TOC is removed within 60 min. Total CFX degradation was achieved after 30 min. About 57% TOC removal also was reported. About 90% COD removal was achieved after 30 min. TC could not be degraded by H2O2 alone. 44.8% adsorption on PAC was achieved after 180 min. By Fenton-like process, 94% TC removal was obtained after 240 min. The pseudo-first order rate constant was 0.019 (min1) for TC degradation. Leachate iron from catalyst was less than allowed iron concentration in the effluent. The catalyst was shown high reusability and activity during the four sequential runs. Process could be conducted in a wide range of pH value in Fe3O4/HA-H2O2 system. Modification of catalyst with HA enhances the removal efficiency about 3.4 times. About total PhAC degradation was achieved after 1 h and 90% TOC removal was reported after 6 h. The pseudo-first order rate constant for PhAC and TOC removal were 0.034 and 0.0048 min1, respectively. 100% removal of DCF was observed in 120s with pyrite Fenton system. 65% of DCF was removed by classic Fenton system in 180s. The pseudo-first order rate constant was 0.164 (s1) for pyrite Fenton system.
Ref.
(Perisic et al., 2016)
(Niu et al., 2012)
(Bobu et al., 2008)
(Jaafarzadeh et al., 2015)
(Niu et al., 2011)
(Bae et al., 2013)
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
[Diclofenac]0 ¼ 0.1 mM
Remarks and findings
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
677
Table 4 Selected articles on application of Fenton-type process in pilot and industrial scale. Considered contaminants
Experimental conditions
Remarks
Ref.
UV254, Fenton and ph-Fenton processes were used as the posttreatment stage for the removal of recalcitrant ECs from MWTP effluent. By using a conventional biological process, the degradation efficiency of CBZ, DCF, SMX, benzotriazole and mecoprop were 0, 34, 86, 0 and 0% respectively. During 33 s irradiation by UV254 (alone) an overall mineralization of 26% and 9.6% was achieved for all examined ECs. The natural dissolved iron content (1.6 mg/L) was good enough to remove ECs higher than 80%. The Fenton process limitations such as low pH range and high reagent concentration were overcome by using powerful UV lamps. Mineralization efficiencies of 0%e22% was achieved. [Chloramphenicol]0 ¼ 229 mg/L Reactor Type: CPC Primary laboratory-scale tests were performed to find the optimum Illuminated volume: 12 L [DOC]0 ¼ 111 mg/L condition for pilot-scale operations. Light source: Sunlight (41.5 W/m2) During the first 40 min, due to the formation of carboxylic acids, the [COD]0 ¼ 289 mg/L toxicity was increased, then it was decreased. [Fe2þ] ¼ 10 mg/L 100% CAP removal was achieved by Fenton process after 40 min. [H2O2] ¼ 400 mg/L After 60 min of reaction, 92% and 98% DOC and COD removal were pH 2.5e2.8 achieved, respectively. Temperature ¼ 45e55 C Solar ph-Fenton process was used as the post-treatment after the [NXA]0 ¼ 38 mg/L Pharmaceutical wastewater bioreactor operation. [Fe2þ] ¼ 20 mg/L [DOC]0 ¼ 725 mg/L No removal of NXA was shown by conventional biological treatment. [COD]0 ¼ 3400 mg/L [H2O2] ¼ 300 mg/L After 25 min illumination in solar ph-Fenton, NXA was totally Reactor Type: CPC degraded and DOC concentration reached 20 mg/L. Light source: Sunlight, The overall efficiency of the combined system for the removal of DOC Illuminated volume: 22 L was 97%. Total volume: 40 L Using biological processes before the ph-Fenton, significantly reduced Light source: Sunlight, the operation costs which is a major drawback of using AOPs alone. pH 2.6e2.8 No toxicity was shown based on the D. magna bioassay after treatment by ph-Fenton process. [AAP]0 ¼ 10 mg/L Primary experiments confirmed that no hydrolysis and photolysis of Distilled water & S.E. 2þ PhACs occurred without the catalyst. [ATL]0 ¼ 10 mg/L [Fe ] ¼ 5 mg/L Total degradation of AAP and ATL were reported after 12 and 3.8 min [H2O2] ¼ 10e20 mg/L of illumination by solar ph-Fenton in DW, respectively. Reactor Type: CPC Total degradation of AAP and ATL were reported after 21.8 and 30 min Light source: Sunlight of illumination by solar ph-Fenton in S.E., respectively. Illuminated volume: 36 L The ph-Fenton treatment was more efficient for the degradation of Total volume: 150 L ATL and ACTP than the photocatalytic treatment with TiO2. pH 2.6e2.8 Solar ph-Fenton, ozonation and solar/TiO2 were used as the postCombination of 66 ECs include MWTP 2þ 16 PhACs treatment after the secondary biological treatment process and the [Fe ] ¼ 5 mg/L [DOC]0 ¼ 13e23 mg/L removal efficiency showed the following trend: Solar ph[H2O2] ¼ 60 mg/L Fenton > ozonation > solar/TiO2. [COD]0 ¼ 32e63 mg/L Reactor Type: CPC Light source: Sunlight Over 98% ECs removal was achieved in only a few minutes of reaction Illuminated volume: 44.6 L by the ph-Fenton process. Total volume: 75 L ph-Fenton treatment produced an effluent with extremely low (non Temperature ¼ 35 C toxic) inhibition (<23% V. fischeri). Solar ph-Fenton is a potential advanced treatment in MWTPs with pH 2.8 treatment costs of <0.4 V/m3. Total removal of 4-MAA and 10% TOC removal were observed during [dipyrone]0 ¼ 50 mg/L Demineralised water the dark Fenton reaction within 15 min. [Fe2þ] ¼ 2 mg/L By applying the ph-Fenton process, 75% mineralization was rapidly [H2O2] ¼ 200e500 mg/L achieved in the first 30 min. Reactor Type: CPC By using solar/TiO2, 4-MAA was degraded at a considerably lower rate Light source: Sunlight Illuminated volume: 22 L than that observed during the ph-Fenton process. Total volume: 35 L V. fischeri assays demonstrated that in ph-Fenton treatment solutions, Temperature ¼ 30e40 C toxicity did not increase during the process. pH 2.8 A heterogeneous Fenton process was used to remove ECs from ECs include PhACs and MWTP biologically pre-treated municipal wastewater. hormones [H2O2] ¼ 200 mg/L [COD]0 ¼ 93.9 mg/L Reactor Type: continuous stir tank Performing the oxidation process at the natural pH of wastewater is a major advantage compared to the homogeneous Fenton process, reactor which is limited to pH in the range of 2e4. Total volume: 31.34 L >90% of the hormones and >40% of PhACs removal were reported. Temperature ¼ 12e20 C BOD was reduced to less than 1 mg/L after treatment. In addition, 30% Natural pH e40% TOC, 50% TSS, 68e91% turbidity and 90% phosphate reduction residence time ¼ 3 h were achieved. There was less than 4% iron leachate, suggesting a good stability of the catalyst. 22 ECs include 15 PhACs and 5 MWTP effluent pesticides [Fe2þ] ¼ 1.6 mg/L (natural) [H2O2] ¼ 20e50 mg/L Light source: 5 150 W UV254 lamps, Reactor volume (pilot): 37 L pH 6e7 (natural)
factor to achieve high removal efficiencies. In other words, at the optimum H2O2/iron ratio, the required iron ion and the H2O2 doses
(De la Cruz et al., 2013)
et al., 2013) (Trovo
(Sirtori et al., 2009b)
(Radjenovi c et al., 2009)
(Prieto-Rodríguez et al., 2013)
rez-Estrada et al., 2007) (Pe
(Chi et al., 2013)
can be reduced considerably (Tekin et al., 2006). The removal efficiencies of contaminants and the TOC, as well as the reaction
678
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
pathway, are affected by the ratio between various reactants ndez-Arriaga et al., 2010). In general, the optimal H2O2/iron (Me ratio depends upon the water matrix to be oxidized (Tekin et al., 2006). It is reported that the H2O2/Fe2þ ratio higher than 10 may lower the removal efficiency of contaminants due to the scavenging of hydroxyl radicals, according to Eq. (5) (Bautitz and Nogueira, 2007). Based on Eq. (1), it is clear that the formation of OH radicals in Fenton reaction also depends on the concentration of Fe2þ ions. The higher solution concentration of Fe2þ can increase the production of hydroxyl radicals, which produces a rapid removal of contaminants in the solution (Li et al., 2006; de Luna et al., 2013). However, further increases in iron ion dosage above the optimum concentration will not only lead to limited improvement in the degradation of contaminants, but will also result in the turbidity of solution due to the presence of excess iron (Badawy et al., 2006; Alalm et al., 2015). This turbidity prevents the penetration of UV light into the solution and consequently the absorption of light required for photolysis. In addition, a high concentration of Fe2þ ions may react with the produced OH radicals as a radical scavenger according to Eq. (8) (Tekin et al., 2006).
high concentrations of contaminants require higher concentrations of hydroxyl radicals for their effective removal. Therefore, the increase in the input concentration of contaminants, while maintaining a constant level of other operating parameters, will reduce the efficiency of degradation (Tekbas¸ et al., 2008; Li et al., 2015). In addition, the number of active sites, where the radical reactions take place, will become limited when the contaminant concentration is too high in the heterogeneous Fenton reaction. Moreover, the high concentration of contaminants will reduce the absorption of light by hydrogen peroxide in photocatalytic Fenton reaction, thereby hindering the OH radical generation (Michael et al., 2010; Li et al., 2015). At higher contaminant concentrations, the oxidation reaction requires higher reaction times and/or higher dosage of reactants for higher removal efficiency of contaminants. These phenomena indicate that in the Fenton processes, the optimum concentration of reagents should be determined based on the initial concentration of organic contaminants to ensure their effective removal.
Fe2þ þ OH /Fe3þ þ OH
Although, FeSO4.7H2O is commercially available for use in homogeneous Fenton and photo-Fenton processes, other types of ferrous catalysts can be used, but they will substantially affect the removal efficiencies, especially under solar light irradiation (Bautitz et al., 2011, 2012). For instance, during and Nogueira, 2007; Trovo the removal of bezafibrate and paracetamol, the use of FeOx in the photo-Fenton process as an alternative reagent to produce Fe2þ ions is highly effective, especially under solar radiation, because it strongly absorbs the solar light spectrum between 250 and 500 nm and has a high quantum efficiency of Fe2þ production (4Fe2þ ¼ 1:24 at 300 nm). Under irradiation, the Fe3þ-polycarboxylate complex will react according to the following reactions to produce Fe2þ ions et al., 2008): (Trovo
(8)
It has been shown that the photo-Fenton reaction requires a lower concentration of reagents to achieve the same degradation squez et al., 2014). This is rate compared to the Fenton reaction (Vela due to the generation of additional hydroxyl radicals and the reduction of Fe3þ ions according to Eq. (3) and Eq. (4), respectively. In heterogeneous Fenton processes, the reactions between iron ions and H2O2 at the catalyst surface similar to that in the homogenous type is expected to occur for the generation of OH radicals, according to Eqs. (1) and (2) (Aleksi c et al., 2010). Considering the similarities in heterogeneous and homogeneous Fenton-type processes, a similar impact of H2O2 concentration on the removal efficiency of contaminants is expected. Therefore, according to Eq. (5), the excess H2O2 concentration leads to the production of less reactive species such as HO2 due to the scavenging effect, causing a reduction in the removal efficiency. In addition, a higher concentration of solid catalyst will not only provide higher surface area and active sites, but it will also increase the concentration of dissolved iron ions in the solution (Bae et al., 2013). The excess amount of solid catalyst prevents efficient light absorption in the heterogeneous photocatalytic reaction by blocking the penetration of photons into the solution (Li et al., 2006; Kasiri et al., 2008). In the heterogeneous photocatalytic Fenton process, four main reactants, namely solid catalyst, H2O2, UV irradiation and oxygen are present. The catalyst and H2O2 are the most controlling parameters, since the absence of catalyst or H2O2 will result in negligible rates of reaction. The presence of UV irradiation and air (oxygen) are important since they will increase the removal rates (Tekbas¸ et al., 2008). 3.2. Effect of initial contaminants concentration The concentration of contaminants in the effluent of wastewater treatment plants varies on a daily basis. Therefore, it is essential to assess the dependence of removal efficiency on the input concentration of pollutants (Titouhi and Belgaied, 2016). Considering very short lifetime of hydroxyl radicals, the increase in the concentration of contaminants will raise the possibility of collision between OH species and organic matter, leading to an improvement in the rate of removal (Kasiri et al., 2008). On the other hand, increasing the input concentration of contaminants has been shown to reduce their removal efficiency (Michael et al., 2010). This occurs because
3.3. Effect of catalyst type
2þ ½FeðC2 O4 Þ3 þ C2 O þ 3C2 O2 4 /Fe 4 þ 2CO2
(9)
3 þ hv/Fe2þ þ 2C2 O2 FeðC2 O4Þ3 4 þ C 2 O4
(10)
2 C2 O 4 þ O2 /O2 þ 3C2 O4 þ 2CO2
(11)
Using FeOx instead of Fe(NO3)3 in the photo-Fenton process under solar irradiation was very effective for the removal of organic contaminants (Nogueira et al., 2005). The high absorption of ferrioxalate makes a large portion of the solar spectrum available for sito et al., 2016). However, the generation of hydroxyl radicals (Expo in spite of the existing beneficial effects, using FeOx as the source of iron in photo-Fenton process increases the carbon load in the solution. Oxalate can be completely converted to CO2 under optimal conditions. While using Fe(NO3)3 as the source of iron was favored for tetracycline degradation under black-light irradiation (Bautitz and Nogueira, 2007), the use of FeOx was preferred for the degradation of 4-chlorophenol under solar irradiation (Nogueira et al., 2005). Therefore, the iron source should be chosen based on the chemical characteristic of the target contaminants and the type of irradiation source (Nogueira et al., 2005; Bautitz and Nogueira, 2007). Similarly, the type and structure of the catalyst can considerably affect the efficiency of heterogeneous Fenton process. A higher TOC removal efficiency of pharmaceuticals by nano-magnetite compared to submicron-structured magnetite was reported (Velichkova et al., 2013). The nanostructure catalysts provide more surface area and a higher number of active sites which decompose hydrogen peroxide, while producing a higher concentration of
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
leached iron (Sun and Lemley, 2011). The generation of a higher concentration of leached iron ions may promote homogeneous Fenton reaction, which leads to a higher removal efficiency of contaminants. On the other hand, it can increase the generation of sludge due to the precipitation of iron. The type of catalyst may also affect the efficiency of removal. For instance, maghemite powder (g-Fe2O3) has been shown to produce higher reaction kinetics for the decomposition of H2O2 to hydroxyl radicals compared to magnetite (Fe3O4) (Velichkova et al., 2013). In addition, due to the higher electron mobility in the spinel Fe3O4, the efficiency of hydroxyl radicals generation is higher than a-Fe2O3 (Rodríguez et al., 2011). It should be kept in mind that the higher activity could be mainly attributed to the higher iron leaching. In contrast to magnetite, in hematite and goethite, iron is initially present only as Fe3þ. Therefore, in the presence of hematite and goethite, the Fe2þ should be produced by irradiation based on Eq. (3) and/or Eq. (4) (Rodríguez et al., 2011). Pyrite not only introduces Fe2þ into the solution, it also reduces the pH of solution based on Eqs. (12) and (13). These two functions are favorable for the homogeneous Fenton process (Bae et al., 2013). þ 2FeS2 þ 7O2 þ 2H2 O/2Fe2þ þ 4SO2 4 þ 4H
(12)
þ 2FeS2 þ 15H2 O2 /2Fe3þ þ 4SO2 4 þ 2H þ 14H2 O
(13)
The differences between zero point charges (zpc) of different iron catalysts can change the mechanism of reaction at various pH values. Considering that the iron catalyst can act as a semiconductor and can produce pairs of electron/holes (e/hþ) under sufficient light energy, the different band gap in various types of iron catalysts can affect the efficiency of reaction. The superior photodegradation of paracetamol was reported in a combination of goethite-H2O2-UV system due to the synergy between the two processes (Mameri et al., 2016). Under UV irradiation, hematite and goethite with 2.2 eV band gap can act as a photocatalyst. The generated electron/hole pairs on the goethite surface can react with contaminants or dissolved oxygen on the goethite active sites and produce strong oxidative radical species. On the other hand, goethite can react with H2O2 according to the photo-Fenton reaction to produce more OH. Researchers have demonstrated the synergism between iron oxides and TiO2 catalyst in the photo-Fenton system (Rodríguez et al., 2009, 2011). This is because the presence of iron oxide may reduce the energy gap of TiO2 photocatalyst, facilitating the electron transfer from the valence to the conduction band of TiO2, which leads to the generation of more hydroxyl radicals. Moreover, a combination of hematite and TiO2 not only reduces the recombination rate of generated electron/hole pairs on TiO2 surface, but also increases the formation of Fe2þ from Fe3þ by capturing the produced electrons on the conduction band of titanium oxide. Besides, the reduction of Fe3þ by capturing the produced electrons on the conduction band of titania may increase iron leaching into the solution. The leached iron should later be removed from the effluent in order to meet the standard for iron, which will increase the operating costs of the process. However, this synergy was not reported in the case of magnetite by Rodríguez et al. (2009). The iron content of catalyst has a significant influence on the degradation of pharmaceuticals. The increase of Fe loading rate beyond the optimum value may have a negative effect on the efficiency of process because the Fe dispersion will become less homogeneous (Bautista et al., 2010b; Titouhi and Belgaied, 2016). In general, the higher iron content of the catalyst offers a better distribution, while increasing the iron concentration in solution (leachate). Considering these two effects, a trade-off should be made in order to find the optimum iron content of the catalyst (Lan et al., 2015).
679
3.4. Effect of pH The pH of solution has a decisive influence on the performance of homogeneous Fenton reaction (Kwon et al., 1999; Shemer et al., 2006; Batista and Nogueira, 2012). The pH of around 3 was mentioned as the optimal value for the Fenton reaction in many studies, while values over 5 have been shown to cause the formation of ferrous hydroxide according to Eq. (14) (Kwon et al., 1999; Gogate and Pandit, 2004; Escher et al., 2006; Hartmann et al., 2010; Vel asquez et al., 2014). Due to the poor solubility of ferrous hydroxide in water, its precipitation will not only consume the ferrous ions and inhibit the Fenton reaction, but it will also reduce the transmission of radiation into the water by increasing the turbidity of solution (Alalm et al., 2015).
Fe2þ þ 2OH /FeðOHÞ2 Y
(14)
The oxidation potential considerably changes at different pH values (E0 ¼ 2.8 V at pH 0 and E0 ¼ 1.59 V at pH 14). Another reason for the decline in the efficiency of photo-degradation at pH > 4 is attributed to the low rate of H2O2 decomposition, which leads to a decrease in the production of OH (Badawy et al., 2006). At pH > 4 hydrogen peroxide is preferentially decomposed into H2O and O2 (Hartmann et al., 2010; Titouhi and Belgaied, 2016). According to Eq. (1), at high pH values, due to the presence of OH species, the reaction shifts to the left and reduces the formation of free radicals. In basic media, hydrogen peroxide is decomposed to oxygen without generating hydroxyl radicals, according to Eq. (15) and Eq. (16) (Zhang et al., 2005).
H2 O2 þ OH 4HO 2 þ H2 O
(15)
H2 O2 þ HO 2 4O2 þ H2 O þ OH
(16)
It has been concluded that at the pH range of 2e4, more Fe(OH)þ complex is formed, which has a higher activity in comparison with Feþ2 in Fenton and photo-Fenton reactions. Alalm et al. (2015) reported that while the complete removal of amoxicillin was obtained at pH 3 after 90 min of irradiation, 80% removal was achieved at neutral condition. The same trend was reported in the case of ampicillin, diclofenac and paracetamol. On the other hand, at a strong acidic condition, the rate of formation of Fe2þ from Fe3þ reduces according to Eq. (4). This happens because at strongly acidic media, where the Hþ species are abundant, the reaction will be shifted towards the left and will reduce the formation of OH radicals (Li et al., 2009). pH values lower than 2.5 increase the formation of [FeOH]2þ, [Fe2þ (H2O)]2þ and [H3O2]þ complex, which react slowly with hydrogen peroxide and generate lower amounts of hydroxyl radicals, thus decreasing the removal efficiencies (Gogate and Pandit, 2004; Shemer et al., 2006; Tekin et al., 2006; Titouhi and Belgaied, 2016). Considering Eq. (17), at very acidic condition (pH < 3) the abundant hydrogen ions (Hþ) can act as a radical scavenger and will reduce the efficiency of reaction. In this reaction, the electron can be gained from the ferrous ion based on Eq. (17) (Tang and Huang, 1996; Shemer et al., 2006; Michael et al., 2010).
OH þ Hþ þ e /H2 O
(17)
At strongly acidic media (pH about 1), H2O2 molecules cannot be decomposed to produce OH radicals. This probably occurs because under this condition, H2O2 will solvate a Hþ to form an oxonium ion ðH3 Oþ 2 Þ (Domínguez et al., 2012). Due to the electrophilic behavior of oxonium ion, the reactivity with ferrous ions is reduced (Kwon et al., 1999). It should be noted that ferric iron precipitates at the pH > 4 and can act as a photocatalyst to generate electronehole
680
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
pairs via excitation by UV irradiation. Therefore, further possible reactions are the oxidation of adsorbed contaminants on the surface of the precipitated ferrous hydroxide and production of addirez-Estrada et al., 2005b). The reactions tional oxidative radicals (Pe in Eqs. (18e26) are possible in the latter case:
hþ þ organic matterðOMÞ/OM /oxidized products vb
(18)
e cb þ O2 ðadsÞ/O2
(19)
O 2 þ H2 O2 /OH þ OH þ O2
(20)
e CB þ H2 O2 /OH þ OH
(21)
þ H2 OðadsÞ/Hþ þ OH hþ vb
(22)
hþ þ HO ðadsÞ/OH vb
(23)
OH þ H2 O2 /H2 O þ HO2
(24)
HO2 þ OH /H2 O þ O2
(25)
H2 O2 þ hþ /HO2 þ Hþ
(26)
Additionally, the pH value can affect the reaction pathways by changing the structure and characteristics of contaminants such as their solubility. The pollutant is in molecular form when the pH is less than its pKa, whereas at pH > pKa the compound loses a proton and becomes negatively charged. For example, diclofenac is insoluble at low pH values (around 3), which is a favorable condition for Fenton reaction. Therefore, due to the incompatibility between diclofenac insolubility and the photo-Fenton process performance rez-Estrada at low pH, the optimization of pH value is essential (Pe et al., 2005b). It is possible that the buffer capacity of water will alter the reaction pathway and will partly overcome the iron and contaminant precipitation. For instance, diclofenac degradation in water without buffering capacity was favorable because starting the reaction at neutral pH, where diclofenac is very soluble, led to its high degradation. During the treatment, pH changes from 7 to about 4 due to the production of intermediate carboxylic acids. As a result, iron precipitation will slow down, yielding an overall rez-Estrada et al., 2005b). accelerated degradation (Pe Recently, Romero et al. (2016a) proposed a method to resolve the problems related to the formation of ferrous hydroxide at pH value higher than 4. They used resorcinol as a chemical additive in order to form Fe-RES (Ligand-Fe) which can surround the iron and hinder the precipitation reaction at neutral pH values. At the end of reaction, the target contaminant and the chemical additive were both degraded. By using this method, the degraded concentration of the target contaminant and the TOC removal efficiency were improved considerably compared to the conventional photoFenton process (pH 2.8) (Romero et al., 2016a). However, this method requires more investigation to determine the involved reactions and the implicated chemical pathway. Although the homogeneous Fenton process is efficient at the acidic pH of around 3, heterogeneous type processes can be effectively operated at milder pH conditions. A substantial dependence of the heterogeneous Fenton reaction on the solution’s pH was reported (Feng et al., 2006). Similar to the results obtained in the homogeneous Fenton process, the removal efficiency increases in the heterogeneous Fenton process by lowering the pH value. This is because of the enhanced leaching of iron ions from the solid catalyst at lower pH values which imposes additional limitations on the
process (Tekbas¸ et al., 2008). Under strongly acidic conditions (pH lower than 2), the removal efficiency declines due to the formation of H3 Oþ 2 ions, thus decreasing its reactivity with ferrous ion (Aleksic et al., 2010). It should be stressed that in the heterogeneous Fenton processes, the leaching of iron ions into the solution can be associated with the removal of contaminants, which will increase the removal efficiency of the process. Thus, increasing the pH of solution can restrain the homogeneous Fenton reaction through the formation of stable ferrous hydroxide, according to Eq. (14). Using the ZSM5 zeolite as the support medium in the heterogeneous Fenton processes can extend the pH range of the homogeneous Fenton reaction. Such effect could be related to the specific characteristics of ZSM5 zeolite, where iron cations are under the influence of a strong electrostatic field in the zeolite framework. The distribution of Fe cations over the zeolite structure with appropriate distribution of the negative charge can prevent or postpone the formation of iron hydroxides as pH approaches neutral conditions (Kasiri et al., 2008; Aleksi c et al., 2010). Moreover, Aleksi c et al. (2010) reported that the volume of leached iron from the zeolite support strongly decreases with the increase of pH value. This happens because at the strongly acidic conditions, the electrostatic force is weaker, thus allowing the leaching of iron ions. The same trend was reported for pillared clays used as the support medium in heterogeneous Fenton reactions (Feng et al., 2006). However, incorporating aluminum besides iron in the clay structure could markedly reduce the iron leachate even at low pH values due to the presence of a strong interaction between aluminum and iron (Li et al., 2006). 3.5. Effect of intensity of radiation The direct photolysis of contaminants occurs when the chemicals absorb light with enough energy (Elmolla and Chaudhuri, 2009). The direct photo-oxidation reaction is based on electronic excitation of the organic compounds, denoted in most cases an electron transfer from the excited-state C* (Eq. (27)) to molecular oxygen (Eq. (28)) or hemolysis (Eq. (29)) to produce radicals which will then react with oxygen (Eq. (30)) (Legrini et al., 1993). hy
C!C *
(27)
C * þ O2 /C $þ þ O$ 2
(28)
hy
R X!R$ þ X $
(29)
R$ þ O2 /RO2
(30)
Therefore, the efficiency of direct photolysis depends on the photon rate at the wavelength of excitation, the absorption cross sectional area of medium, the quantum yield of the process, and the concentration of dissolved molecular oxygen (Legrini et al., 1993). In most cases, due to the low absorption of chemical contaminants, the direct photolysis is insignificant or may lead to a high reaction time. For instance, no degradation of paracetamol was shown after 3 h of irradiation by 365 nm (Mameri et al., 2016). This is because paracetamol has a weak absorption in the range from 300 to 800 nm. The degradation efficiencies of amoxicillin, ampicillin and cloxacillin were reported to be 2.9%, 3.8% and 4.9%, respectively, after 5 h of irradiation under 365 nm wavelength lamp (Elmolla and Chaudhuri, 2009). The maximum absorption of metronidazole occurs at 310 nm. Therefore, applying medium-pressure UV lamp (200e400 nm) which has a substantial overlap with metronidazole absorption band, increased the removal rate constant by five-fold compared to using a low-pressure lamp (254 nm) (Shemer et al.,
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
2006). Ravina et al. (2002) used various types of UV lamps with different intensities and reported direct proportionality between the mineralization of diclofenac and the intensity of radiation. It is proven that the maximum absorbance of H2O2 occurs at 210e230 nm and the proteolysis is low at a wavelength higher than 300 nm (Pignatello et al., 1999; Elmolla and Chaudhuri, 2009). For instance, at UV-A region (365 nm), the production of hydrogen radicals by direct photolysis was very low due to the low molar absorptivity (~0.01 M-1 cm-1) (Rodríguez et al., 2011). At lower wavelengths, the direct proteolysis of chemical contaminants and H2O2 to hydroxyl radicals takes place simultaneously. In the Fenton reactions, free radicals can be produced in the absence of light energy by the reaction between ferrous salt and H2O2, according to Eq. (1). In this case, the production of hydroxyl radicals is limited since one molecule of Fe2þ produces only one OH radical. The illumination of Fenton reaction (photo-Fenton) is a proven method to overcome the limitation in the production of sufficient hydroxyl radicals (Eq. (3) and Eq. (4)). Considering Eq. (3), the regenerated Fe2þ is subsequently re-oxidized by H2O2 and produces new OH radicals. Consequently, the oxidation of organic contaminants is accelerated (Badawy et al., 2006). Therefore, the contaminants can be degraded more efficiently using fewer reagents or in less reacsquez et al., 2014). H2O2 can be decomposed to OH tion time (Vela radicals in the presence of UV light according to Eq. (4), while UV irradiation may excite contaminates to act as intermediate initiators to produce OH radicals in a chain reaction. At the pH between 2 and 3, [Fe(OH)(H2O)5]2þ and [Fe(OH)]2þ are predominant. These species are photosensitive to radiation in the range of 280e405 nm, making available the use of sunlight based on Eq. (31) (Vel asquez et al., 2014).
FeðOHÞðH2 OÞ5
2þ
2þ þ H2 O þ hy/ FeðH2 OÞ6 þ OH$
(31)
While the presence of solid catalyst increases the turbidity of solution and reduces the solution transparency and radiation flux, the heterogeneous photo-Fenton process may be more effective compared to the homogeneous photo-Fenton processes at quasi-neutral pH values. Higher removal efficiencies, higher mineralization and milder operating pH values have been reported as the advantages of photo- over dark-heterogeneous Fenton processes (Aleksi c et al., 2010). However, a thorough cost analysis should be performed in order to evaluate the viability of the overall process for the treatment of emerging contaminants in a water matrix.
3.6. Effect of water matrix The degradation of pharmaceuticals is dependent on the type of water matrix such as pure water, surface water and sewage treatment plant effluent. Therefore, in the case of pharmaceuticals which appear in the effluent of WWTPs, it is important to assess the et al., 2008). effect of this matrix on the removal efficiency (Trovo The WWTPs effluent usually contains chlorides, organic and inorganic carbon (carbonate and bicarbonate). The presence of these species can strongly reduce the removal efficiencies of pharmaceuticals by the scavenging hydroxyl radicals. Humic substances may act as an inner filter and severely decrease removal efficiencies (Legrini et al., 1993). The organic matters present in the real STPs can interfere in different manners and compete with the target components for degradation or by radiation attenuation in photoFenton reactions. The high carbonate content of solution can reduce the efficiency of removal due to its scavenging effect, according to Eqs. (32) and (33) (Bautitz and Nogueira, 2007).
681
$ $ HCO 3 þ OH /H2 O þ CO3
(32)
$ CO2 3 þ OH /OH þ CO3
(33)
3.7. Effect of the salinity of substrate A high concentration of inorganic salts, especially NaCl, has been detected in industrial wastewaters, such as those producing pharmaceuticals, pesticides, dyes and herbicides (Bacardit et al., 2007). 2 The presence of inorganic ions such as Cl ; SO2 4 ; HPO4 ; HPO4 may reduce the efficiency of Fenton processes. This may occur since anions react with hydroxyl radicals and generate anion radicals that are less reactive compared to hydroxyl radicals (Eqs. (34) and (35)) (Sirtori et al., 2009b). The interaction of chloride with Fe2þ/Fe3þ ions may produce complex chemical substances according to Eqs. (36)e(39) (Bacardit et al., 2007).
Cl þ OH$ 4ClOH$ 4Cl$ þ OH
(34)
ClOH$ þ Hþ 4ClOH2$ 4Cl$ þ H2 O
(35)
Fe2þ þ Cl /FeClþ
(36)
FeClþ þ Cl /FeCl02
(37)
Fe3þ þ Cl 4FeCl2þ
(38)
Fe3þ þ 2Cl /FeClþ 2
(39)
The chloride ions can react with Fe3þ under irradiation which may further produce organic radicals (R) or chloride radicals Cl, which are less reactive in comparison with the hydroxyl radical (Eqs. (40) and (41)). * e3þ
F
*
hy
ðaqÞCl !Fe2þ ðaqÞ þ Cl$
FeCl3 þ RH/FeCl2 þ HCl þ R$
Cl$ þ Cl /Cl$ 2
(40) (41) (42)
Where *Fe indicates the photoexcited species. These photoexcited species can also participate in the reaction but at a lower rate. In the distilled water, 76% removal of sulfamethoxazole by photolysis was reported after 7 h, while under the same condition, et al., 2009). only 14% removal was shown by using seawater (Trovo The inhabitation of Fenton and photo-Fenton reactions by a high concentration of sodium chloride strongly depends on the pH of solution. At the pH of equal or less than 2, the photolysis of Fe3þ complexes in the presence of Cl ions leads to the formation of the Cl$ 2 radical anion which is less reactive than OH radicals (Eq. (38), (40) and (42)). In addition, under these conditions, the scavenging of hydroxyl radical by the chloride ion (Eq. (34) and (35)) becomes important (Machulek et al., 2007). In the photo-Fenton process, at the initial stage of reaction where the pH of solution is adjusted to 3e3.5 (optimum pH for Fe(OH)2þ formation), the removal efficiency is only slightly affected by the presence of chloride ions, even at a high concentration (0.5 M). However, since the pH of solution is usually not controlled during the oxidation process, the formation of organic acids will cause a progressive acidification and drop in the pH value. In this condition, chloride ion begins to
682
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
scavenge any OH radicals formed in the system and will form Cl$ 2 species as a product. Therefore, there is an obvious influence of chloride ion on the photocatalytic oxidation process (Machulek et al., 2006). The inhibition of reaction can be circumvented by maintaining the solution pH at approximately 3 throughout the course of reaction, even in the presence of high chloride ion concentrations (0.5 M) (Machulek et al., 2006, 2007). Bacardit et al. (2007) demonstrated that the presence of high concentrations of NaCl reduced the rate of removal reaction in the photo-Fenton process. This is due to the fact that sodium chloride prevents or reduces the generation of hydroxyl radicals due to the chain reactions of chloride and iron ions. These investigators also reported that the presence of phosphate ðPO3 4 Þ completely inhibits the Fenton process and severely affects TOC removal. This should be expected since the addition of iron ions leads to the formation of iron-phosphate complexes, commonly used for phosphate precipitation and its removal in wastewater treatment plants. Finally, the introduction of acid or base for the adjustment of solution pH can affect the removal efficiency of contaminants. For instance, the acidification of solution by HCl introduces a high amount of conjugated anions (Cl) into the solution. These anions can reduce the efficiency of removal due to the drastic competition between the anions and contaminants for the hydroxyl radicals (Eq. (34)) (Kasiri et al., 2008). 3.8. Effect of feeding mode Several researchers have investigated the influence of the feeding mode of Fenton reagent to the reactor on the efficiency of process (Bowers et al., 1989; Turan-Ertas and Gurol, 2002; Zhang et al., 2005). Due to the competition between the species that participate in the reaction such as iron ions, radicals, hydrogen peroxide, target contaminants and produced intermediates, changing of the feeding mode may alter the reaction mechanism and the removal efficiency of pharmaceuticals. When H2O2 was added in two steps, the efficiency of COD removal was improved, as reported by Zhang et al. (2005). This effect might be due to the radical scavenging effect of H2O2 when introducing a large initial amount of hydrogen peroxide according to Eq. (5). Therefore, successive addition of the H2O2 will keep the hydrogen peroxide concentration at relatively low levels, decreasing the detrimental effect of hydroxyl radical scavenging (Turan-Ertas and Gurol, 2002; Zhang et al., 2005). In the solar reactor, the recirculation rate may affect the process efficiency. A faster recirculation of solution in the photo-reactor leads to the improvement of mixing inside the reactor. However, a very high recirculation rate can reduce the reactor’s residence time. The reduction of contact time between the contaminants and Fenton reagents will lead to the generation of fewer radicals and a reduced reaction rate (Ravina et al., 2002). 3.9. Effect of temperature Temperature is another important parameter in the Fenton processes. Theoretically, the increase of temperature should enhance the reaction kinetics (Velichkova et al., 2013). In the Fenton reaction, the increase of temperature accelerates the reaction rate between ferrous ions and H2O2, which causes the generation of additional hydroxyl radicals. The increase of reaction temperature can enhance the decomposition rate of H2O2 by a factor of 2 at each 10 C temperature increase (Hartmann et al., 2010). However, conducting the reaction at a temperature higher than 40 C may decrease the removal rate. It is generally accepted that at high temperatures, the decomposition of H2O2 to water and oxygen is possible (Guo et al., 2015). Zazo et al. (2011) reported that a
significant improvement in the mineralization of phenol was observed up to 90 C. They concluded that the increase of temperature could enhance hydrogen peroxide transformation into hydroxyl radicals. Therefore, a considerable improvement of the oxidation rate and the mineralization efficiency can result, leading to a reduction in the reagent dosage at high temperatures. The decrease of Fe2þ and hydrogen peroxide dose reduces sludge generation and improves the efficiency of process by minimizing scavenging reactions. Besides, it is expected that COD removal can be enhanced by the increase of temperature (Zhang et al., 2005). Nevertheless, Tekin et al. (2006) reported that no significant differences were observed in the outcome of reactions conducted at room temperature or at 40 C. It is worth mentioning that, although the increase of temperature imposes a large economical burden to the process in real STPs, it does not represent a major drawback for the treatment of many industrial effluents, which usually have a high temperature. Heat recovery from the treated stream allows saving energy (Zazo et al., 2011). Therefore, various effects of operation temperature must be taken into consideration when defining the optimal operating temperature for organic matter removal (Bautista et al., 2010b). Similar trends are expected in the heterogeneous Fenton process. The increase of operating temperature improves the efficiency of organic matter removal, as the higher temperatures allow the reagents to overcome the activation energy barrier based on the Arrhenius law, thus facilitating the oxidation reactions. However, high temperatures may introduce certain inconveniences such as increasing the possible presence of iron leachate in the solution (Rodríguez et al., 2009). For instance, increasing the temperature from 50 to 70 C reduced the reaction time by half during the removal of phenol, while at the same time increasing the loss of iron from the catalyst by three times compared to that at 50 C (Zazo et al., 2006). In addition, high temperature can cause the deterioration of support medium which will cause the loss of catalyst smoothness and its decomposition into fine particles (Titouhi and Belgaied, 2016). The reaction pathway can also change with temperature. For instance, the production of oxalic acid as a by-product of phenol oxidation was increased by increasing the temperature (Zazo et al., 2006). As mentioned before, high temperature may also enhance the thermal degradation rate of hydrogen peroxide. Therefore, the oxidation rate of organic contaminants should be a function of competition between thermal degradation of H2O2 and hydroxyl radical formation (Liou et al., 2005). 3.10. Effect of reaction time Because of the toxicity of intermediate by-products, the moment of reaction termination should be carefully evaluated if complete mineralization is not the goal. This implies that in certain cases, while the complete removal of target contaminant has been achieved after a short period of reaction, a major fraction of the contaminant has actually been transformed into the intermediates. Consequently, more reaction time is needed to ensure the mineralization of contaminants and their reliable removal (Li et al., 2015). For instance, the degradation of sulfathiazole led to the formation of maleic acid whose concentration reached a peak concentration squez et al., after 2 min and totally disappeared after 8 min (Vela 2014). Therefore, the evaluation of reaction time for the total mineralization of target contaminants and by-products is recommended. As stated before, the degradation rate of pharmaceutical depends on the presence and abundance of hydroxyl radicals. At the early stages of reaction, usually high degradation rates of contaminants are achieved due to the abundance of produced hydroxyl radicals. With the progress of reaction, the degradation rate
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
decreases due to the consumption of hydroxyl radicals and reduction in the concentration of contaminants (Alalm et al., 2015).
3.11. Deactivation of catalyst In the heterogeneous Fenton reaction, the performance of catalyst may decrease after consecutive runs. The main reasons for the deactivation of catalyst are the leaching of iron from the catalyst structure and/or the decay of active catalytic sites (Tekbas¸ et al., 2008). The leaching of iron which mainly occur at the low pH values will cause the reaction to shift to the homogeneous Fenton reaction. In this case, the removal rate usually increases. However, the capacity of catalyst in the subsequent stages of the process will substantially decrease. Moreover, the leaching of iron from a catalyst results in the formation of a secondary pollution (Nidheesh et al., 2013). The leachate of iron species from iron oxides depends on the structure of the catalyst. The corundum structure (aFe2O3) is more stable than Fe3O4 with spinel structure. Thus, the leachate from (a-Fe2O3) is considerably lower than Fe3O4 (Rodríguez et al., 2009; Nidheesh et al., 2013). The decrease of pH value increases the iron leaching, while the leachate from iron oxides is negligible at neutral pH (Rodríguez et al., 2009; Perisic et al., 2016). As mentioned before, the increase of temperature causes higher leaching of iron and consequently produces the deactivation of catalyst in successive cycles (Ramirez et al., 2007; Rodríguez et al., 2009). In the latter case, consecutive runs may cause surface blockage, which changes the porous structure and limits the accessibility of target compounds to the pores after reuse. It is possible that the adsorption of produced contaminants and residual organic matter at the surface of the catalyst, will lead to the formation of deposits on the surface (Nidheesh et al., 2013; Titouhi and Belgaied, 2016). Some modifications to the catalyst structure can improve the stability of catalyst by diminishing iron leachate during the oxidation. For instance, an excellent stability of heterogeneous Fenton catalyst using aluminum as mixing metal with iron in pillared clay was demonstrated (Li et al., 2006, 2015).
3.12. By-products formation Considering the non-selective nature of hydroxyl radicals, various by-products may be formed at low concentrations in radical-induced reactions (Radjenovi c et al., 2009). Therefore, during the treatment of effluents of wastewater treatment plants, complete mineralization is more desirable than the elimination of specific target contaminants. This is important since a complete removal of TOC will ensure complete mineralization of contaminants and potential intermediate compounds into CO2 and H2O (Feng et al., 2006). It should be noted that the produced intermediates can play incompatible roles as they can form complexes with iron ions and change the efficiency of degradation. For instance, organic intermediates accelerated the degradation of pchlorophenol by the reduction of ferric ion to ferrous ion (Kwon et al., 1999). Conversely, some intermediates can act as ferrous scavenger and prevent or delay the Fe2þ/Fe3þ regeneration cycle by forming stable complexes. The formation of a stable complex of organic intermediates with ferrous and ferric ions has been reported (Kwon et al., 1999). In some cases, the production of iron leachate and the deactivation of catalyst are related to the production of intermediate compounds. Some intermediate compounds such as oxalic acid can capture iron ions from the catalyst and form iron complexes. When these intermediates are mineralized into H2O and CO2, the iron ions return to the surface of catalyst.
683
4. Toxicity and biodegradability tests The degradation products and/or metabolites of emerging contaminants should be taken into consideration since the byproducts produced during the degradation processes are sometimes more toxic than the parent compounds (Ballesteros Martín et al., 2008; Oller et al., 2011; Prieto-Rodríguez et al., 2013; Veloutsou et al., 2014). A very high toxicity was reported in the case of isoproturon degradation by the photo-Fenton process, especially during the early stages of reaction (Parra et al., 2000). However, toxicity assessments demonstrated that photo-Fenton processes showed better performance in comparison with TiO2 rez-Estrada photocatalysis for the removal of pharmaceuticals (Pe et al., 2007). The generated intermediates may be more toxic to the microorganisms than the original contaminants (Oller et al., 2011). For that reason, a toxicity assessment is crucial when AOPs are used as a pretreatment step in hybrid systems (PrietoRodríguez et al., 2013). Toxicity assays may be applied as a major criterion for selecting the type of pretreatment process upstream of biotreatment operations (Lapertot et al., 2006). It is also important to mention that most AOPs reduce the pH of solution due to the formation of inorganic acids. In Fenton processes, the solution pH should be around 3. Hence, prior to conducting toxicity tests, the solution should be neutralized and chemical reagents such as hydrogen peroxide, catalysts, etc. must be removed from the media. Various tests such as D. magna, Selenastrum capricornutum, V. fischeri, Pseudomonas, Staphylococcus aureus, Escherichia coli, Phaeodactylum tricornutum, Pseudokirchneriella subcapitata and Lepidium sativum have been used for toxicity assessment (Oller et al., 2011). By using D. magna toxicity test, Arslan-Alaton and Gurses (2004) concluded that the photo-Fenton-like process is an effective method for the complete detoxification and partial oxidation of PPG production effluent. In another research, the high efficiency of Fenton process as a pretreatment stage for the elimination of acute toxicity and increased biodegradability of an effluent was shown by conducting activated sludge inhibition test (ISO 8192) (Alaton and Teksoy, 2007). The improvement of biodegradability and decrease of ecotoxicity to the marine bacteria V. fischeri by solar photo-Fenton treatment of ECs were reported (Zapata et al., 2009). It is worth mentioning that the water matrix can substantially alter the generated intermediates and consequently the toxicity of the treated water. For instance, D. magna bioassays showed that the degradation of SMX when performed in distilled water led to the decrease of toxicity from 85% to 20%. However, the toxicity increased from 16% to 86% when the reaction et al., 2009). During the removal was performed in seawater (Trovo of paracetamol by using FeSO4, a significant reduction of the sample toxicity was reported compared to using FeOx as the iron source et al., 2012). This is because the intermediate compounds (Trovo generated in distilled water are different from those generated in seawater. However, by conducting the V. Fischeri assay, Zapata et al. (2008) concluded that although the photo-Fenton process was very effective in the degradation of pharmaceuticals, the acute toxicity of effluents persisted until the end of treatment, while biodegradability was only slightly improved. Although toxicity tests provide valuable information regarding the effect of Fenton reactions on subsequent biological processes, further biodegradability tests (BOD5/COD) should be conducted to ensure the effectiveness of biological processes. This is important since bacteriaor fungi, which are commonly used in conventional biological water treatment processes, are sensitive to changes in the environmental parameters such as pH, temperature, dissolved oxygenconcentration, salinity, redox potential, and the presence of toxic compounds (Oller et al., 2011). Biodegradability is not directly related to toxicity, but it could be inferred that if toxicity enhances
684
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
during the treatment process, the biodegradability of solution should decrease (Lapertot et al., 2006). The enhancement in biodegradability of pharmaceuticals has been reported in various studies following Fenton-based reactions (Tekin et al., 2006; Badawy et al., 2009; Elmolla and Chaudhuri, 2009; Giri and Golder, 2014). For instance, biodegradability (BOD5/COD) was improved significantly from 0.043 to 0.63 after 120 min of exposure to Fenton reactions, in the case of DCF removal (Perisic et al., 2016). At the optimum condition, about 2.5 fold enhancement of (BOD5/ COD) ratio was reported after the treatment of a secondary effluent ~ ones et al., 2015). of a municipal wastewater treatment plant (Quin García-Ripoll et al. (2009) used Pseudomonas putida culture as a convenient, reproducible and reliable test for evaluating both biodegradability and toxicity, as an alternative to other timeconsuming tests for ECs solution partially treated by photoFenton process. 5. Large-scale application Extensive studies assessing Fenton-type processes for the removal of a wide range of pharmaceuticals from drinking water were reviewed and discussed in this manuscript. Most of the reviewed papers are based on bench-scale operations as there are only few reports on pilot-scale or industrial applications of Fenton technologies. Small-scale investigations often used a model system prepared by dissolving a single EC in distilled water and evaluated parameters such as COD, BOD5 and DOC in laboratory-scale operations (Zapata et al., 2010). Despite the high efficiency of Fenton processes in the removal of toxic and recalcitrant contaminants such as ECs, the high operating costs limit their commercial or industrial use. One attractive cost-cutting approach is the exploitation of renewable energy resources such as sunlight in photoFenton processes (Sirtori et al., 2009b, 2011; Zapata et al., 2010; et al., 2013). Another promPrieto-Rodríguez et al., 2013; Trovo ising solution is the integration of Fenton processes with the conventional biological treatment processes (Sirtori et al., 2009b, 2011). From this point of view, Fenton reactions are mostly utilized as a pre-treatment process in order to improve the biodegradability of contaminants. Compared to biological processes, AOPs are very effective in the complete mineralization of contaminants, but they are expensive when applied alone due to the requirement for energy and chemical reagents (Oller et al., 2011). Therefore, exploiting a chemical oxidation process such as Fentontype processes as the pre-treatment to transform the initially recalcitrant contaminants into more biodegradable intermediates is a promising approach with a significantly lower cost (Sarria et al., 2003; Lapertot et al., 2006). In a typical design approach, a Fentontype process oxidizes toxic substances, and once the required biodegradability has been achieved, the resulting effluent is transferred to a conventional biological treatment system (Sirtori et al., 2009a). Zapata et al. (2010) optimized the combination of photo-Fenton and biological processes based on H2O2 consumption in both pilot and full-scale plants. De la Cruz et al. (2013) successfully removed 22 ECs (more than 80%) from the effluent of secondary MWTP by ph-Fenton process at neutral pH condition and by using the iron content of effluent. The problem associated with the limitations of Fenton process, i.e. working at low pH value and high reagent consumption, were addressed by the design optimization of pilot setup at very low residence times (De la Cruz et al., 2012; De la Cruz et al., 2013). In addition, some treatment plants such as pharmaceutical wastewater treatment plants produce acidic effluents. Therefore, the natural pH level of these effluents are favorable for Fenton reactions without requiring any adjustment (Sirtori et al., 2009b). In another research, the combination of high frequency ultrasound and solar photo-Fenton process was used in
order to decrease both H2O2 consumption and the treatment time for ECs removal (Papoutsakis et al., 2015). In this case, a significant synergy was shown in the removal of both ECs and TOC. Nevertheless, the high cost of ultrasound process in comparison with photo-Fenton process still remains an obvious challenge during the treatment of real industrial effluents. Fenton-type processes can also be applied as the post-treatment step. Sirtori et al. (2009b) significantly reduced chemical consumption by employing the ph-Fenton process as a polishing step for the conventional biological stage. As a post-treatment step to polish off the effluent of rez-Estrada et al. (2007) used an iron sewage treatment plant, Pe catalyst at low concentration (2 mg/L) in order to avoid additional steps to reduce the iron content in the final effluent. Chi et al. (2013) utilized the heterogeneous Fenton process at pilot-scale for the removal of PhACs and hormones from secondary MWTP effluent for the first time. By using the post-treatment Fenton process, not only were the ECs removed successfully at natural pH and temperature, but also other parameters such as BOD, TOC, TSS, phosphate and turbidity were reduced; hence rendering this technology more competitive for full-scale applications. 6. Summary of findings The commercial availability of H2O2, its possible on-site storage, high solubility in water, thermal stability, production of two hydroxyl radicals by each molecule of H2O2, lack of mass transfer limitations associated with the dissolution of gas phase into the water (compared to O3), cost-effectiveness as a source of OH radicals, and the simple operation procedure highlight the advantages of using H2O2 as a the source of hydroxyl radicals (Kumar and Bansal, 2013). Nevertheless, there are certain limitations associated with the use of H2O2. They include the small absorption cross section of H2O2 at 254 nm which limits its application, especially when dissolved organic matters act as inner filters. The use of another source of irradiation such as Xedoped Hg arcs, which exhibit a strong emission in the spectral region of 210e240 nm, can resolve this problem since hydrogen peroxide has a higher molar absorption coefficient in this wavelength range. Conducting reactions in distilled water usually produces higher removal efficiencies in comparison with using surface waters or seawater. However, these results generally lead to erroneous large-scale designs and the overestimation of real process efficiency. Therefore, the presence of ions in real effluents and their contribution to increasing the removal efficiency of contaminants can be considered as a promising approach. On the other hand, applying the Fenton process in waters with high salinity may lead to the formation of by-products which are more toxic et al., 2009). than their parent compounds (Trovo The need for performing Fenton and photo-Fenton processes under acidic condition is often considered as one of their major disadvantages because it will impose additional costs for acidification and subsequent neutralization. Recently, it is proven that using ligands as a chemical additive can resolve this problem by forming complexes with iron ions (Romero et al., 2016a). Time-based kinetic relationships are commonly used to compare and evaluate the effectiveness of various removal processes. This is because other parameters such as energy demand and cost estimation can be determined based on the kinetic data. In the photo-Fenton process, the direct photolysis of chemical contaminants along with the reactions with hydrogen peroxide contribute to the degradation of contaminants. Despite the controlling effects of wavelength and irradiation intensity on the reaction rates of photolysis processes, these parameters are often ignored when kinetic models are developed.
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
Therefore, removal efficiencies cannot be properly evaluated by using the process dynamics, especially by using the data generated in bench-scale operations. Using the Fenton and photo-Fenton processes as a pretreatment step to biological processes produces promising results and has received attention in recent years. This approach will improve the biodegradability of organic pollutants and will enhance the efficiency of biological treatment (Kajitvichyanukul and Suntronvipart, 2006). However, the combined process will require the adjustment of pH to approximately 7 upstream of biological processes, proper handling of the chemical oxidant to prevent adverse effects on the biomass, and optimization of chemical dosage and reaction time. During the operation of homogeneous Fenton and photo-Fenton processes, the selection of iron source is important for effective wastewater treatment, and it should be selected based on the chemical characteristic of the target contaminants (Bautitz and Nogueira, 2007). If the composition of contaminants is unknown, a simple evaluation test at bench scale can be carried out in order to determine the most appropriate iron source. A major advantage of the photo-Fenton processes, besides their simplicity, is the possibility of operation at room temperature and at atmospheric pressure and the use of sunlight or near UV light as the energy sources. These operational aspects result in significant economic savings, especially in sunny countries. The use of Ferrioxalate as the source of iron is very efficient when using solar light irradiation, mostly due to its high absorption in the UVevis region and its high quantum yield of Fe2þ generation sito et al., 2016). (Expo In the case of effluents with high salinity, the photo-Fenton process is very slow and has a low economic viability, particularly when using an artificial radiation source (Bacardit et al., 2007). In these cases, employing solar photo-Fenton process is strongly advised. The use of heterogeneous Fenton-type process instead of the homogenous type offers significant advantages, such as the operation at milder pH, formation of significantly less sludge and elimination of the need for pH adjustment or neutralization. However, it has been demonstrated that the transparency of solution may be reduced by using suspended catalyst particles (Kaur et al., 2016). Therefore, the application of fixed catalysts, instead of suspended catalysts, and immobilization of iron onto a support medium are recommended.
7. Conclusion The presence of emerging contaminants such as pharmaceuticals in aquatic media has received considerable attention due to their potential health effects on human and their toxicity to various microorganisms. Conventional wastewater treatment processes such as flocculation, filtration, coagulation and sedimentation, are not efficient in the removal of pharmaceuticals from water. Therefore, these contaminants have the potential to accumulate and persist in the aquatic environment. Although pharmaceutical chemicals appear in the water at relatively low concentrations (ng/L to mg/L), they may impose acute toxicity effects on the environment. Advanced oxidation processes, which are based on the oxidation of organic contaminants by hydroxyl radicals, can be considered as an effective method to eliminate emerging contaminants from the environment. Among oxidation processes, Fenton processes are easy to operate, they are safe and relatively cheap, produce fast reactions and require environmentally-benign reagents. However, homogeneous Fenton processes suffer from certain limitations such as operation at a narrow pH range and the
685
generation of iron-containing sludge. In addition, the regeneration of iron ions is not feasible and the final effluent should be treated to meet the discharge standards for iron concentration. The immobilization of iron species on a support medium, in a process known as the heterogeneous Fenton, is an effective approach to overcome these limitations. In this process, the stability of catalyst and leaching of iron ions into the solution play significant roles and control the rate of reactions. Photo-Fenton process not only enhances the removal efficiency of pharmaceuticals, but also reduces the reagents consumption and waste generation. Although, Fenton processes do not always mineralize the contaminants, they can significantly enhance their biodegradability. Therefore, by integrating Fenton reactions with biological processes, it is possible to develop an advanced treatment process with higher efficiency and lower costs. Future research should focus on the use of solar energy, which can minimize energy costs, while reducing fossil fuel consumption and greenhouse gas emission, advancing sustainable development aspects of contaminant removal processes. References Alalm, M.G., Tawfik, A., Ookawara, S., 2015. Degradation of four pharmaceuticals by solar photo-Fenton process: kinetics and costs estimation. J. Environ. Chem. Eng. 3, 46e51. Alaton, I.A., Teksoy, S., 2007. Acid dyebath effluent pretreatment using Fenton’s reagent: process optimization, reaction kinetics and effects on acute toxicity. Dyes Pigm. 73, 31e39. Aleksi c, M., Kusi c, H., Koprivanac, N., Leszczynska, D., Bo zi c, A.L., 2010. Heterogeneous Fenton type processes for the degradation of organic dye pollutant in water d the application of zeolite assisted AOPs. Desalination 257, 22e29. Arslan-Alaton, I., Gurses, F., 2004. Photo-Fenton-like and photo-fenton-like oxidation of Procaine Penicillin G formulation effluent. J. Photochem. Photobiol. A Chem. 165, 165e175. rez, A.-A., Solís-Lo pez, M., Ramírez-Zamora, R.-M., Arzate-Salgado, S.-Y., Morales-Pe 2016. Evaluation of metallurgical slag as a Fenton-type photocatalyst for the degradation of an emerging pollutant: Diclofenac. Catal. Today 266, 126e135. €tzner, J., Chamarro, E., Esplugas, S., 2007. Effect of salinity on the Bacardit, J., Sto photo-fenton process. Ind. Eng. Chem. Res. 46, 7615e7619. Badawy, M.I., Ghaly, M.Y., Gad-Allah, T.A., 2006. Advanced oxidation processes for the removal of organophosphorus pesticides from wastewater. Desalination 194, 166e175. Badawy, M.I., Wahaab, R.A., El-Kalliny, A.S., 2009. Fenton-biological treatment processes for the removal of some pharmaceuticals from industrial wastewater. J. Hazard. Mater. 167, 567e574. Bae, S., Kim, D., Lee, W., 2013. Degradation of diclofenac by pyrite catalyzed Fenton oxidation. Appl. Catal. B Environ. 134e135, 93e102. rez, J.A., Acie n Ferna ndez, F.G., Casas Lo pez, J.L., Ballesteros Martín, M.M., S anchez Pe García-Ripoll, A.M., Arques, A., Oller, I., Malato Rodríguez, S., 2008. Combined photo-Fenton and biological oxidation for pesticide degradation: effect of phototreated intermediates on biodegradation kinetics. Chemosphere 70, 1476e1483. Batista, A.P.S., Nogueira, R.F.P., 2012. Parameters affecting sulfonamide photoFenton degradation e iron complexation and substituent group. J. Photochem. Photobiol. A Chem. 232, 8e13. Bautista, P., Mohedano, A.F., Casas, J.A., Zazo, J.A., Rodriguez, J.J., 2010a. Oxidation of cosmetic wastewaters with H2O2 using a Fe/g-Al2O3 catalyst. Water Sci. Technol. 61, 1631e1636. ndez, N., Casas, J.A., Rodriguez, J.J., 2010b. CataBautista, P., Mohedano, A.F., Mene lytic wet peroxide oxidation of cosmetic wastewaters with Fe-bearing catalysts. Catal. Today 151, 148e152. Bautitz, I.R., Nogueira, R.F.P., 2007. Degradation of tetracycline by photo-Fenton processdsolar irradiation and matrix effects. J. Photochem. Photobiol. A Chem. 187, 33e39. Bobu, M., Yediler, A., Siminiceanu, I., Schulte-Hostede, S., 2008. Degradation studies of ciprofloxacin on a pillared iron catalyst. Appl. Catal. B Environ. 83, 15e23. ~ ez, M., Fabregat-Safont, D., Morales, E., Pastor, L., Sancho, J.V., S Boix, C., Ib an anchezRamírez, J.E., Hern andez, F., 2016. Behaviour of emerging contaminants in sewage sludge after anaerobic digestion. Chemosphere 163, 296e304. Bowers, A.R., Gaddipati, P., Eckenfelder, W.W., Monsen, R.M., 1989. Treatment of toxic or refractory wastewaters with hydrogen peroxide. Water Sci. Technol. 21, 477e486. , M., Domínguez-Morueco, N., Migens, A., Molina, R., Martínez, F., Valc Catala arcel, Y., pez de Alda, M., Barcelo , D., Segura, Y., 2015. Elimination of Mastroianni, N., Lo drugs of abuse and their toxicity from natural waters by photo-Fenton treatment. Sci. Total Environ. 520, 198e205. Chi, G.T., Churchley, J., Huddersman, K.D., 2013. Pilot-scale removal of trace steroid hormones and pharmaceuticals and personal care products from municipal wastewater using a heterogeneous Fenton’s catalytic process. Int. J. Chem. Eng. 2013, 10.
686
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
Chong, S., Zhang, G., Zhang, N., Liu, Y., Zhu, J., Huang, T., Fang, S., 2016. Preparation of FeCeOx by ultrasonic impregnation method for heterogeneous Fenton degradation of diclofenac. Ultrason. Sonochem. 32, 231e240. , H.J., Rodrigues, A.E., Moreira, R.F.P.M., 2006. Dantas, T.L.P., Mendonça, V.P., Jose Treatment of textile wastewater by heterogeneous Fenton process using a new composite Fe2O3/carbon. Chem. Eng. J. 118, 77e82. De la Cruz, N., Esquius, L., Grandjean, D., Magnet, A., Tungler, A., de Alencastro, L.F., Pulgarín, C., 2013. Degradation of emergent contaminants by UV, UV/H2O2 and neutral photo-Fenton at pilot scale in a domestic wastewater treatment plant. Water Res. 47, 5836e5845. nez, J., Esplugas, S., Grandjean, D., de Alencastro, L.F., De la Cruz, N., Gime Pulgarín, C., 2012. Degradation of 32 emergent contaminants by UV and neutral photo-fenton in domestic wastewater effluent previously treated by activated sludge. Water Res. 46, 1947e1957. de Luna, M.D.G., Briones, R.M., Su, C.-C., Lu, M.-C., 2013. Kinetics of acetaminophen degradation by Fenton oxidation in a fluidized-bed reactor. Chemosphere 90, 1444e1448. lez, T., Palo, P., Cuerda-Correa, E.M., 2012. Fenton þ fentonDomínguez, J.R., Gonza like integrated process for carbamazepine degradation: optimizing the system. Ind. Eng. Chem. Res. 51, 2531e2538. Elmolla, E.S., Chaudhuri, M., 2009. Degradation of the antibiotics amoxicillin, ampicillin and cloxacillin in aqueous solution by the photo-Fenton process. J. Hazard. Mater. 172, 1476e1481. Elmolla, E.S., Chaudhuri, M., 2010. Comparison of different advanced oxidation processes for treatment of antibiotic aqueous solution. Desalination 256, 43e47. Elmolla, E.S., Chaudhuri, M., 2011. Combined photo-FentoneSBR process for antibiotic wastewater treatment. J. Hazard. Mater. 192, 1418e1426. Escher, B.I., Bramaz, N., Richter, M., Lienert, J., 2006. Comparative ecotoxicological hazard assessment of beta-blockers and their human metabolites using a mode-of-action-based test battery and a QSAR approach. Environ. Sci. Technol. 40, 7402e7408. sito, A.J., Dura n, A., Monteagudo, J.M., Acevedo, A., 2016. Solar photoExpo degradation of a pharmaceutical wastewater effluent in a semi-industrial autonomous plant. Chemosphere 150, 254e257. Feng, J., Hu, X., Yue, P.L., 2006. Effect of initial solution pH on the degradation of Orange II using clay-based Fe nanocomposites as heterogeneous photo-Fenton catalyst. Water Res. 40, 641e646. Ganiyu, S.O., van Hullebusch, E.D., Cretin, M., Esposito, G., Oturan, M.A., 2015. Coupling of membrane filtration and advanced oxidation processes for removal of pharmaceutical residues: a critical review. Sep. Purif. Technol. 156 (Part 3), 891e914. García-Ripoll, A., Amat, A.M., Arques, A., Vicente, R., Ballesteros Martín, M.M., rez, J.A.S., Oller, I., Malato, S., 2009. Confirming Pseudomonas putida as a reliPe able bioassay for demonstrating biocompatibility enhancement by solar photooxidative processes of a biorecalcitrant effluent. J. Hazard. Mater. 162, 1223e1227. Garrido-Ramírez, E.G., Theng, B.K.G., Mora, M.L., 2010. Clays and oxide minerals as catalysts and nanocatalysts in Fenton-like reactions d a review. Appl. Clay Sci. 47, 182e192. Georgi, A., Kopinke, F.-D., 2005. Interaction of adsorption and catalytic reactions in water decontamination processes: Part I. Oxidation of organic contaminants with hydrogen peroxide catalyzed by activated carbon. Appl. Catal. B Environ. 58, 9e18. Giri, A.S., Golder, A.K., 2014. Fenton, photo-fenton, H2O2 photolysis, and TiO2 photocatalysis for dipyrone oxidation: drug removal, mineralization, biodegradability, and degradation mechanism. Ind. Eng. Chem. Res. 53, 1351e1358. Gogate, P.R., Pandit, A.B., 2004. A review of imperative technologies for wastewater treatment I: oxidation technologies at ambient conditions. Adv. Environ. Res. 8, 501e551. lez, O., Sans, C., Esplugas, S., 2007. Sulfamethoxazole abatement by photoGonza Fenton: toxicity, inhibition and biodegradability assessment of intermediates. J. Hazard. Mater. 146, 459e464. Gonzalez, O., Sans, C., Esplugas, S., Malato, S., 2009. Application of solar advanced oxidation processes to the degradation of the antibiotic sulfamethoxazole. Photochem. Photobiol. Sci. 8, 1032e1039. Guo, R., Xie, X., Chen, J., 2015. The degradation of antibiotic amoxicillin in the Fenton-activated sludge combined system. Environ. Technol. 36, 844e851. Hartmann, M., Kullmann, S., Keller, H., 2010. Wastewater treatment with heterogeneous Fenton-type catalysts based on porous materials. J. Mater. Chem. 20, 9002e9017. Hermosilla, D., Cortijo, M., Huang, C.P., 2009. Optimizing the treatment of landfill leachate by conventional Fenton and photo-Fenton processes. Sci. Total Environ. 407, 3473e3481. Herney-Ramirez, J., Vicente, M.A., Madeira, L.M., 2010. Heterogeneous photo-Fenton oxidation with pillared clay-based catalysts for wastewater treatment: a review. Appl. Catal. B Environ. 98, 10e26. Hu, X., Liu, B., Deng, Y., Chen, H., Luo, S., Sun, C., Yang, P., Yang, S., 2011. Adsorption and heterogeneous Fenton degradation of 17a-methyltestosterone on nano Fe3O4/MWCNTs in aqueous solution. Appl. Catal. B Environ. 107, 274e283. Huerta-Fontela, M., Galceran, M.T., Ventura, F., 2011. Occurrence and removal of pharmaceuticals and hormones through drinking water treatment. Water Res. 45, 1432e1442. Jaafarzadeh, N., Kakavandi, B., Takdastan, A., Kalantary, R.R., Azizi, M., Jorfi, S., 2015. Powder activated carbon/Fe3O4 hybrid composite as a highly efficient heterogeneous catalyst for Fenton oxidation of tetracycline: degradation mechanism and kinetic. RSC Adv. 5, 84718e84728.
Jonidi Jafari, A., Kakavandi, B., Jaafarzadeh, N., Rezaei Kalantary, R., Ahmadi, M., Akbar Babaei, A., 2016. Fenton-like catalytic oxidation of tetracycline by AC@ Fe3O4 as a heterogeneous persulfate activator: adsorption and degradation studies. J. Ind. Eng. Chem. http://dx.doi.org/10.1016/j.jiec.2016.09.044. Kajitvichyanukul, P., Suntronvipart, N., 2006. Evaluation of biodegradability and oxidation degree of hospital wastewater using photo-Fenton process as the pretreatment method. J. Hazard. Mater. 138, 384e391. Kasiri, M.B., Aleboyeh, H., Aleboyeh, A., 2008. Degradation of Acid Blue 74 using FeZSM5 zeolite as a heterogeneous photo-Fenton catalyst. Appl. Catal. B Environ. 84, 9e15. Kaur, A., Umar, A., Kansal, S.K., 2016. Heterogeneous photocatalytic studies of analgesic and non-steroidal anti-inflammatory drugs. Appl. Catal. A General 510, 134e155. Kavitha, V., Palanivelu, K., 2004. The role of ferrous ion in Fenton and photo-Fenton processes for the degradation of phenol. Chemosphere 55, 1235e1243. Kuan, C.-C., Chang, S.-Y., Schroeder, S.L.M., 2015. Fenton-like oxidation of 4chlorophenol: homogeneous or heterogeneous? Ind. Eng. Chem. Res. 54, 8122e8129. Kumar, J., Bansal, A., 2013. Photocatalysis by nanoparticles of titanium dioxide for drinking water purification: a conceptual and state-of-art review. Mater. Sci. Forum 764, 130e150. Kunde, G.B., Yadav, G.D., 2015. Synthesis, characterization and application of ironaluminate nodules in advanced Fenton’s oxidation process. J. Environ. Chem. Eng. 3, 2010e2021. Kusi c, H., Koprivanac, N., Bo zi c, A.L., Selanec, I., 2006a. Photo-assisted Fenton type processes for the degradation of phenol: a kinetic study. J. Hazard. Mater. 136, 632e644. Kusi c, H., Koprivanac, N., Selanec, I., 2006b. Fe-exchanged zeolite as the effective heterogeneous Fenton-type catalyst for the organic pollutant minimization: UV irradiation assistance. Chemosphere 65, 65e73. Kuznetsova, E.V., Savinov, E.N., Vostrikova, L.A., Parmon, V.N., 2004. Heterogeneous catalysis in the fenton-type system FeZSM-5/H2O2. Appl. Catal. B Environ. 51, 165e170. Kwon, B.G., Lee, D.S., Kang, N., Yoon, J., 1999. Characteristics of p-chlorophenol oxidation by Fenton’s reagent. Water Res. 33, 2110e2118. Lan, H., Wang, A., Liu, R., Liu, H., Qu, J., 2015. Heterogeneous photo-Fenton degradation of acid red B over Fe2O3 supported on activated carbon fiber. J. Hazard. Mater. 285, 167e172. ndez-Ib ~ ez, P., Maldonado, M.I., Pe rez-Estrada, L., Lapertot, M., Pulgarín, C., Ferna an Oller, I., Gernjak, W., Malato, S., 2006. Enhancing biodegradability of priority substances (pesticides) by solar photo-Fenton. Water Res. 40, 1086e1094. Legrini, O., Oliveros, E., Braun, A.M., 1993. Photochemical processes for water treatment. Chem. Rev. 93, 671e698. Li, H., Li, Y., Xiang, L., Huang, Q., Qiu, J., Zhang, H., Sivaiah, M.V., Baron, F., Barrault, J., Petit, S., Valange, S., 2015. Heterogeneous photo-Fenton decolorization of Orange II over Al-pillared Fe-smectite: response surface approach, degradation pathway, and toxicity evaluation. J. Hazard. Mater. 287, 32e41. Li, R., Yang, C., Chen, H., Zeng, G., Yu, G., Guo, J., 2009. Removal of triazophos pesticide from wastewater with Fenton reagent. J. Hazard. Mater. 167, 1028e1032. Li, W.C., 2014. Occurrence, sources, and fate of pharmaceuticals in aquatic environment and soil. Environ. Pollut. 187, 193e201. Li, Y., Lu, Y., Zhu, X., 2006. Photo-Fenton discoloration of the azo dye X-3B over pillared bentonites containing iron. J. Hazard. Mater. 132, 196e201. Liou, R.-M., Chen, S.-H., Hung, M.-Y., Hsu, C.-S., Lai, J.-Y., 2005. Fe (III) supported on resin as effective catalyst for the heterogeneous oxidation of phenol in aqueous solution. Chemosphere 59, 117e125. Liu, Z.-h., Kanjo, Y., Mizutani, S., 2009. Removal mechanisms for endocrine disrupting compounds (EDCs) in wastewater treatment d physical means, biodegradation, and chemical advanced oxidation: a review. Sci. Total Environ. 407, 731e748. Machulek, A., Moraes, J.E.F., Vautier-Giongo, C., Silverio, C.A., Friedrich, L.C., Nascimento, C.A.O., Gonzalez, M.C., Quina, F.H., 2007. Abatement of the inhibitory effect of chloride anions on the photo-fenton process. Environ. Sci. Technol. 41, 8459e8463. Machulek, A., Vautier-Giongo, C., Moraes, J.E.F., Nascimento, C.A.O., Quina, F.H., 2006. Laser flash photolysis study of the photocatalytic step of the photo-fenton reaction in saline solution. Photochem. Photobiol. 82, 208e212. , L., 2015. FentonMackulak, T., Mosný, M., Grabic, R., Golovko, O., Koba, O., Birosova like reaction: a possible way to efficiently remove illicit drugs and pharmaceuticals from wastewater. Environ. Toxicol. Pharmacol. 39, 483e488. Maezono, T., Tokumura, M., Sekine, M., Kawase, Y., 2011. Hydroxyl radical concentration profile in photo-Fenton oxidation process: generation and consumption of hydroxyl radicals during the discoloration of azo-dye Orange II. Chemosphere 82, 1422e1430. Mameri, Y., Debbache, N., Benacherine, M.e.m., Seraghni, N., Sehili, T., 2016. Heterogeneous photodegradation of paracetamol using Goethite/H2O2 and Goethite/oxalic acid systems under artificial and natural light. J. Photochem. Photobiol. A Chem. 315, 129e137. , V., Brix, H., 2012. Occurrence and Matamoros, V., Arias, C.A., Nguyen, L.X., Salvado behavior of emerging contaminants in surface water and a restored wetland. Chemosphere 88, 1083e1089. ndez-Arriaga, F., Esplugas, S., Gime nez, J., 2010. Degradation of the emerging Me contaminant ibuprofen in water by photo-Fenton. Water Res. 44, 589e595. Michael, I., Hapeshi, E., Michael, C., Fatta-Kassinos, D., 2010. Solar Fenton and solar
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688 TiO2 catalytic treatment of ofloxacin in secondary treated effluents: evaluation of operational and kinetic parameters. Water Res. 44, 5450e5462. Mirzaei, A., Chen, Z., Haghighat, F., Yerushalmi, L., 2016. Removal of pharmaceuticals and endocrine disrupting compounds from water by zinc oxide-based photocatalytic degradation: a review. Sustain. Cities Soc. 27, 407e418. Mirzaei, A., Ebadi, A., Khajavi, P., 2013. Kinetic and equilibrium modeling of single and binary adsorption of methyl tert-butyl ether (MTBE) and tert-butyl alcohol (TBA) onto nano-perfluorooctyl alumina. Chem. Eng. J. 231, 550e560. Mompelat, S., Le Bot, B., Thomas, O., 2009. Occurrence and fate of pharmaceutical products and by-products, from resource to drinking water. Environ. Int. 35, 803e814. Nakada, N., Shinohara, H., Murata, A., Kiri, K., Managaki, S., Sato, N., Takada, H., 2007. Removal of selected pharmaceuticals and personal care products (PPCPs) and endocrine-disrupting chemicals (EDCs) during sand filtration and ozonation at a municipal sewage treatment plant. Water Res. 41, 4373e4382. Navalon, S., Alvaro, M., Garcia, H., 2010. Heterogeneous Fenton catalysts based on clays, silicas and zeolites. Appl. Catal. B Environ. 99, 1e26. Nidheesh, P.V., Gandhimathi, R., Ramesh, S.T., 2013. Degradation of dyes from aqueous solution by Fenton processes: a review. Environ. Sci. Pollut. Res. 20, 2099e2132. Niu, H., Dizhang, Meng, Z., Cai, Y., 2012. Fast defluorination and removal of norfloxacin by alginate/Fe@Fe3O4 core/shell structured nanoparticles. J. Hazard. Mater. 227e228, 195e203. Niu, H., Zhang, D., Zhang, S., Zhang, X., Meng, Z., Cai, Y., 2011. Humic acid coated Fe3O4 magnetic nanoparticles as highly efficient Fenton-like catalyst for complete mineralization of sulfathiazole. J. Hazard. Mater. 190, 559e565. , A.G., 2005. Influence of the iron source on the Nogueira, R.F.P., Silva, M.R.A., Trovo solar photo-Fenton degradation of different classes of organic compounds. Sol. Energy 79, 384e392. rez, J.A., 2011. Combination of Advanced Oxidation Oller, I., Malato, S., S anchez-Pe Processes and biological treatments for wastewater decontaminationda review. Sci. Total Environ. 409, 4141e4166. Pal, A., Gin, K.Y.-H., Lin, A.Y.-C., Reinhard, M., 2010. Impacts of emerging organic contaminants on freshwater resources: review of recent occurrences, sources, fate and effects. Sci. Total Environ. 408, 6062e6069. Papoutsakis, S., Miralles-Cuevas, S., Gondrexon, N., Baup, S., Malato, S., Pulgarin, C., 2015. Coupling between high-frequency ultrasound and solar photo-Fenton at pilot scale for the treatment of organic contaminants: an initial approach. Ultrason. Sonochem. 22, 527e534. ringer, P., Pulgarin, C., 2000. Photochemical versus Parra, S., Sarria, V., Malato, S., Pe coupled photochemicalebiological flow system for the treatment of two biorecalcitrant herbicides: metobromuron and isoproturon. Appl. Catal. B Environ. 27, 153e168. rez-Estrada, L.A., Malato, S., Agüera, A., Ferna ndez-Alba, A.R., 2007. Degradation Pe of dipyrone and its main intermediates by solar AOPs: identification of intermediate products and toxicity assessment. Catal. Today 129, 207e214. rez-Estrada, L.A., Malato, S., Gernjak, W., Agüera, A., Thurman, E.M., Ferrer, I., Pe ndez-Alba, A.R., 2005a. Photo-Fenton degradation of Diclofenac: identiFerna fication of main intermediates and degradation pathway. Environ. Sci. Technol. 39, 8300e8306. rez-Estrada, L.A., Maldonado, M.I., Gernjak, W., Agüera, A., Fern Pe andez-Alba, A.R., Ballesteros, M.M., Malato, S., 2005b. Decomposition of diclofenac by solar driven photocatalysis at pilot plant scale. Catal. Today 101, 219e226. Perisic, D.J., Gilja, V., Stankov, M.N., Katancic, Z., Kusic, H., Stangar, U.L., Dionysiou, D.D., Bozic, A.L., 2016. Removal of diclofenac from water by zeoliteassisted advanced oxidation processes. J. Photochem. Photobiol. A Chem. 321, 238e247. Pignatello, J.J., Liu, D., Huston, P., 1999. Evidence for an additional oxidant in the photoassisted fenton reaction. Environ. Sci. Technol. 33, 1832e1839. Prieto-Rodríguez, L., Oller, I., Klamerth, N., Agüera, A., Rodríguez, E.M., Malato, S., 2013. Application of solar AOPs and ozonation for elimination of micropollutants in municipal wastewater treatment plant effluents. Water Res. 47, 1521e1528. ~ ones, D.H., Alvarez, Quin P.M., Rey, A., Beltr an, F.J., 2015. Removal of emerging contaminants from municipal WWTP secondary effluents by solar photocatalytic ozonation. A pilot-scale study. Sep. Purif. Technol. 149, 132e139. Rache, M.L., García, A.R., Zea, H.R., Silva, A.M.T., Madeira, L.M., Ramírez, J.H., 2014. Azo-dye orange II degradation by the heterogeneous Fenton-like process using a zeolite Y-Fe catalystdkinetics with a model based on the Fermi’s equation. Appl. Catal. B Environ. 146, 192e200. Rad, L.R., Haririan, I., Divsar, F., 2015. Comparison of adsorption and photo-Fenton processes for phenol and paracetamol removing from aqueous solutions: single and binary systems. Spectrochim. Acta Part A Mol. Biomol. Spectrosc. 136 (Part B), 423e428. , D., Malato, S., 2009. Solar photoRadjenovi c, J., Sirtori, C., Petrovi c, M., Barcelo catalytic degradation of persistent pharmaceuticals at pilot-scale: kinetics and characterization of major intermediate products. Appl. Catal. B Environ. 89, 255e264. Rahim Pouran, S., Abdul Aziz, A.R., Wan Daud, W.M.A., 2015. Review on the main advances in photo-Fenton oxidation system for recalcitrant wastewaters. J. Ind. Eng. Chem. 21, 53e69. dar, F.J., Pe rez-Cadenas, A.F., Moreno-Castilla, C., Ramirez, J.H., Maldonado-Ho Costa, C.A., Madeira, L.M., 2007. Azo-dye Orange II degradation by heterogeneous Fenton-like reaction using carbon-Fe catalysts. Appl. Catal. B Environ. 75, 312e323.
687
Ravina, M., Campanella, L., Kiwi, J., 2002. Accelerated mineralization of the drug Diclofenac via Fenton reactions in a concentric photo-reactor. Water Res. 36, 3553e3560. ndez, G., Ledesma, B., Alvarez, n, F.J., 2009. PhotoRodríguez, E., Ferna P., Beltra catalytic degradation of organics in water in the presence of iron oxides: influence of carboxylic acids. Appl. Catal. B Environ. 92, 240e249. ndez, G., Alvarez, n, F.J., 2011. Rodríguez, E.M., Ferna P.M., Hern andez, R., Beltra Photocatalytic degradation of organics in water in the presence of iron oxides: effects of pH and light source. Appl. Catal. B Environ. 102, 572e583. nez, J., Esplugas, S., 2016a. Enhancement of Romero, V., Acevedo, S., Marco, P., Gime Fenton and photo-Fenton processes at initial circumneutral pH for the degradation of the b-blocker metoprolol. Water Res. 88, 449e457. nez, J., Esplugas, S., 2016b. Romero, V., Gonz alez, O., Bayarri, B., Marco, P., Gime Degradation of Metoprolol by photo-Fenton: comparison of different photoreactors performance. Chem. Eng. J. 283, 639e648. Sanchis, S., Polo, A.M., Tobajas, M., Rodriguez, J.J., Mohedano, A.F., 2013. Degradation of chlorophenoxy herbicides by coupled Fenton and biological oxidation. Chemosphere 93, 115e122. ringer, P., Pulgarin, C., 2003. Degradation of a biorecalcitrant Sarria, V., Deront, M., Pe dye precursor present in industrial wastewaters by a new integrated iron(III) photoassistedebiological treatment. Appl. Catal. B Environ. 40, 231e246. Shemer, H., Kunukcu, Y.K., Linden, K.G., 2006. Degradation of the pharmaceutical Metronidazole via UV, Fenton and photo-Fenton processes. Chemosphere 63, 269e276. Sirtori, C., Zapata, A., Gernjak, W., Malato, S., Lopez, A., Agüera, A., 2011. Solar photoFenton degradation of nalidixic acid in waters and wastewaters of different composition. Analytical assessment by LCeTOF-MS. Water Res. 45, 1736e1744. Sirtori, C., Zapata, A., Oller, I., Gernjak, W., Agüera, A., Malato, S., 2009a. Decontamination industrial pharmaceutical wastewater by combining solar photoFenton and biological treatment. Water Res. 43, 661e668. Sirtori, C., Zapata, A., Oller, I., Gernjak, W., Agüera, A., Malato, S., 2009b. Solar photofenton as finishing step for biological treatment of a pharmaceutical wastewater. Environ. Sci. Technol. 43, 1185e1191. Snyder, S.A., Adham, S., Redding, A.M., Cannon, F.S., DeCarolis, J., Oppenheimer, J., Wert, E.C., Yoon, Y., 2007. Role of membranes and activated carbon in the removal of endocrine disruptors and pharmaceuticals. Desalination 202, 156e181. Song, R., Bai, B., Puma, G.L., Wang, H., Suo, Y., 2015. Biosorption of azo dyes by raspberry-like Fe3O4@yeast magnetic microspheres and their efficient regeneration using heterogeneous Fenton-like catalytic processes over an up-flow packed reactor. React. Kinet. Mech. Catal. 115, 547e562. Soon, A.N., Hameed, B.H., 2011. Heterogeneous catalytic treatment of synthetic dyes in aqueous media using Fenton and photo-assisted Fenton process. Desalination 269, 1e16. Sum, O.S.N., Feng, J., Hub, X., Yue, P.L., 2005. Photo-assisted fenton mineralization of an azo-dye acid black 1 using a modified laponite clay-based Fe nanocomposite as a heterogeneous catalyst. Top. Catal. 33, 233e242. Sun, S.-P., Lemley, A.T., 2011. p-Nitrophenol degradation by a heterogeneous Fentonlike reaction on nano-magnetite: process optimization, kinetics, and degradation pathways. J. Mol. Catal. A Chem. 349, 71e79. Sun, S.-P., Zeng, X., Lemley, A.T., 2013. Nano-magnetite catalyzed heterogeneous Fenton-like degradation of emerging contaminants carbamazepine and ibuprofen in aqueous suspensions and montmorillonite clay slurries at neutral pH. J. Mol. Catal. A Chem. 371, 94e103. Tang, W.Z., Huang, C.P., 1996. 2,4-Dichlorophenol oxidation kinetics by Fenton’s reagent. Environ. Technol. 17, 1371e1378. Tekbas¸, M., Yatmaz, H.C., Bektas¸, N., 2008. Heterogeneous photo-Fenton oxidation of reactive azo dye solutions using iron exchanged zeolite as a catalyst. Microporous Mesoporous Mater. 115, 594e602. Tekin, H., Bilkay, O., Ataberk, S.S., Balta, T.H., Ceribasi, I.H., Sanin, F.D., Dilek, F.B., Yetis, U., 2006. Use of Fenton oxidation to improve the biodegradability of a pharmaceutical wastewater. J. Hazard. Mater. 136, 258e265. Tijani, J.O., Fatoba, O.O., Babajide, O.O., Petrik, L.F., 2016. Pharmaceuticals, endocrine disruptors, personal care products, nanomaterials and perfluorinated pollutants: a review. Environ. Chem. Lett. 14, 27e49. Titouhi, H., Belgaied, J.-E., 2016. Heterogeneous Fenton oxidation of ofloxacin drug by iron alginate support. Environ. Technol. 1e13. , A.G., de Paiva, V.A.B., Machado, A.E.H., de Oliveira, C.A., Santos, R.O., 2013. Trovo Degradation of the antibiotic chloramphenicol by photo-Fenton process at labscale and solar pilot plant: kinetic, toxicity and inactivation assessment. Sol. Energy 97, 596e604. , A.G., Melo, S.A.S., Nogueira, R.F.P., 2008. Photodegradation of the pharmaTrovo ceuticals amoxicillin, bezafibrate and paracetamol by the photo-Fenton processdapplication to sewage treatment plant effluent. J. Photochem. Photobiol. A Chem. 198, 215e220. , A.G., Nogueira, R.F.P., Agüera, A., Fernandez-Alba, A.R., Sirtori, C., Malato, S., Trovo 2009. Degradation of sulfamethoxazole in water by solar photo-Fenton. Chemical and toxicological evaluation. Water Res. 43, 3922e3931. , A.G., Pupo Nogueira, R.F., Agüera, A., Fernandez-Alba, A.R., Malato, S., 2011. Trovo Degradation of the antibiotic amoxicillin by photo-Fenton process e chemical and toxicological assessment. Water Res. 45, 1394e1402. , A.G., Pupo Nogueira, R.F., Agüera, A., Fernandez-Alba, A.R., Malato, S., 2012. Trovo Paracetamol degradation intermediates and toxicity during photo-Fenton treatment using different iron species. Water Res. 46, 5374e5380. Turan-Ertas, T., Gurol, M.D., 2002. Oxidation of diethylene glycol with ozone and
688
A. Mirzaei et al. / Chemosphere 174 (2017) 665e688
modified Fenton processes. Chemosphere 47, 293e301. n ~ez, J., Zaror, C., Salazar, R.A., Pe rezVel asquez, M., Santander, I.P., Contreras, D.R., Ya Moya, M., Mansilla, H.D., 2014. Oxidative degradation of sulfathiazole by Fenton and photo-Fenton reactions. J. Environ. Sci. Health, Part A 49, 661e670. Velichkova, F., Julcour-Lebigue, C., Koumanova, B., Delmas, H., 2013. Heterogeneous Fenton oxidation of paracetamol using iron oxide (nano)particles. J. Environ. Chem. Eng. 1, 1214e1222. Veloutsou, S., Bizani, E., Fytianos, K., 2014. Photo-Fenton decomposition of bblockers atenolol and metoprolol; study and optimization of system parameters and identification of intermediates. Chemosphere 107, 180e186. Vergili, I., Gencdal, S., 2015. Applicability of combined Fenton oxidation and nanofiltration to pharmaceutical wastewater. Desalination Water Treat. 56, 3501e3509. Yuan, S., Gou, N., Alshawabkeh, A.N., Gu, A.Z., 2013. Efficient degradation of contaminants of emerging concerns by a new electro-Fenton process with Ti/MMO cathode. Chemosphere 93, 2796e2804. nchez-Pe rez, J.A., Oller, I., Maldonado, M.I., 2010. Scale-up Zapata, A., Malato, S., Sa strategy for a combined solar photo-Fenton/biological system for remediation of pesticide-contaminated water. Catal. Today 151, 100e106. Zapata, A., Oller, I., Gallay, R., Pulgarín, C., Maldonado, M.I., Malato, S., Gernjak, W., 2008. Comparison of photo-fenton treatment and coupled photo-fenton and
biological treatment for detoxification of pharmaceutical industry contaminants. J. Adv. Oxid. Technol. 11, 261e269. nchez-Pe rez, J.A., Mantzavinos, D., Maldonado, M.I., Zapata, A., Velegraki, T., Sa Malato, S., 2009. Solar photo-Fenton treatment of pesticides in water: effect of iron concentration on degradation and assessment of ecotoxicity and biodegradability. Appl. Catal. B Environ. 88, 448e454. Zazo, J.A., Casas, J.A., Mohedano, A.F., Rodríguez, J.J., 2006. Catalytic wet peroxide oxidation of phenol with a Fe/active carbon catalyst. Appl. Catal. B Environ. 65, 261e268. Zazo, J.A., Pliego, G., Blasco, S., Casas, J.A., Rodriguez, J.J., 2011. Intensification of the fenton process by increasing the temperature. Ind. Eng. Chem. Res. 50, 866e870. Zhang, H., Choi, H.J., Huang, C.-P., 2005. Optimization of Fenton process for the treatment of landfill leachate. J. Hazard. Mater. 125, 166e174. Zhang, X., Bai, B., Li Puma, G., Wang, H., Suo, Y., 2016a. Novel sea buckthorn biocarbon SBC@b-FeOOH composites: efficient removal of doxycycline in aqueous solution in a fixed-bed through synergistic adsorption and heterogeneous Fenton-like reaction. Chem. Eng. J. 284, 698e707. Zhang, Y., Liu, C., Xu, B., Qi, F., Chu, W., 2016b. Degradation of benzotriazole by a novel Fenton-like reaction with mesoporous Cu/MnO2: combination of adsorption and catalysis oxidation. Appl. Catal. B Environ. 199, 447e457.