- Email: [email protected]
Sunlight-driven photo-transformation of bisphenol A by Fe(III) in aqueous solution: Photochemical activity and mechanistic aspects
Chemosphere 167 (2017) 353e359
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Sunlight-driven photo-transformation of bisphenol A by Fe(III) in aqueous solution: Photochemical activity and mechanistic aspects Meilan Pan a, Jie Ding b, Lin Duan a, Guandao Gao b, * a
Ministry of Education Key Laboratory of Pollution Processes and Environmental Criteria, Tianjin Key Laboratory of Environmental Remediation and Pollution Control, College of Environmental Science and Engineering, Nankai University, Tianjin, 300071, China b State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, 210023, China
h i g h l i g h t s
g r a p h i c a l a b s t r a c t
Natural sunlight can activate iron ion-related species and produce strong oxidants. Colloidal [Fe(OH)3]m as a core photocatalytic species were explored. O 2 , rather than OH, was identified as the key active radical in our process. Our researches introduced the novel process for oxidation of organic pollutants.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 4 August 2016 Received in revised form 22 September 2016 Accepted 28 September 2016 Available online 12 October 2016
Iron is one of the most abundant elements in aquatic environments, and plays important roles in the fate and transport of environmental contaminants. Previous studies on the photochemical properties of Fe(III) species have largely focused on complexes formed between Fe(III) and environmental ligands such as natural organic matter (NOM) under UV irradiation, whereas the potentially important roles of hydrolysis species of Fe(III) in Fe(III)-mediated photo-transformation of environmental contaminants under solar light are not fully understood. In this study, the solar light-driven photochemical activities of hydrolysis species of Fe(III) were further explored, using a system containing only 0.5 mM Fe2(SO4)3 and bisphenol A. The important role of colloidal [Fe(OH)3]m, formed from the hydrolysis of Fe3þ, as a core photochemical species of Fe(III) was proposed and verified. Interestingly, O 2 , rather than OH, was identified (via electron spin resonance) as the key active radical responsible for the degradation of bisphenol A. We propose that unlike Fe(OH)2þ, which under UV irradiation can yield OH even in sunlight (Fe(OH)2þ þ hv / Fe2þ þ OH), colloidal [Fe(OH)3]m produces O 2 ([Fe(OH)3]m þ 2O2 þ hv / Fe(II) þ 2O 2 þ H2O). The fact that Fe(III) can produce strong radicals in sunlight may have important environmental implications. © 2016 Published by Elsevier Ltd.
Handling Editor: Jun Huang Keywords: Sunlight-driven Photochemical activity Colloidal [Fe(OH)3]m Bisphenol A
1. Introduction * Corresponding author. School of the Environment, Nanjing University, Nanjing, 210023, China. E-mail address: [email protected] (G. Gao). http://dx.doi.org/10.1016/j.chemosphere.2016.09.144 0045-6535/© 2016 Published by Elsevier Ltd.
et
Iron is one of the most abundant elements on earth (McKnight al., 1988; Bigham and Nordstrom, 2000). In aquatic
354
M. Pan et al. / Chemosphere 167 (2017) 353e359
environments, iron is produced naturally from weathering of iron minerals such as magnetite, hematite, and goethite (Stumm and Sulzberger, 1992; Mitsunobu et al., 2010; Guo and Barnard, 2013), or released as the results of iron mine exploiting and broad applications in industrial fields (Rose and Ghazi, 1997; Edwards et al., 2000; Nordstrom et al., 2000). A number of studies have shown that iron plays important roles in the fate and transport of environmental contaminants in surface and ground waters (Barry et al., 2002; Zhang and Elliott, 2006; Cundy et al., 2008) by participating directly in redox and photochemical reactions(Zuo and Hoigne, 1992; Zuo, 1995; Zuo and Deng, 1997; Chen et al., 2013) or serving as electron mediators in the biodegradation of organic pollutants (Heijman et al., 1995; Kappler and Haderlein, 2003; Van der Zee and Cervantes, 2009; Watanabe et al., 2009). It has been shown that ferric ion, i.e., Fe(III), can oxidize organic compounds in acidic aqueous solutions (pH z 3) when irradiated with UV light, and OH is responsible for the photochemical activity (Evans and Uri, 1949; Bates and Uri, 1953). Previous studies on the photochemical properties of different Fe(III) species have emphasized on the roles of complexes formed between Fe(III) and ligands(Zuo and Hoigne, 1994; Zuo and Zhan, 2005; Zuo et al., 2005, 2013) such as citrate (Nansheng et al., 1998; Conte et al., 2014), oxalate (Zhou et al., 2004; Conte et al., 2014), or natural organic matter (NOM) (Rose and Waite, 2005; Sharp et al., 2006) under UV irradiation. It was proposed that ligand-to-metal charge transition and electron transfer from the ligand to the central metal ion could occur in such complexes and consequently result in the photochemical activity of Fe(III). In aqueous solution Fe(III) has a strong tendency to undergo hydrolysis and form complexes with OH. In fact, the logarithm complex constants between Fe(III) and OH range from 11.9 to 29.3 (Schnitzer and Hansen, 1970; Stumm and Morgan, 2012), much greater than those of Fe(III)NOM complexes (6.1) and the complexes between Fe(III) and model NOMs such as succinic acid (7.5) and citric acid (7.2) (Schnitzer and Hansen, 1970; Stumm and Morgan, 2012). Thus, hydrolysis species of Fe(III), whose absorbance coefficients and quantum yields are comparable to those of FeNOM (Schnitzer and Hansen, 1970; Stumm and Morgan, 2012), can often be the dominant Fe(III) species (Schnitzer and Hansen, 1970; Stumm and Morgan, 2012) even in the presence of relatively high concentration of NOM (10e100 mg/L) (see the results of Visual MINTEQ 3.0 modeling in Supporting Information (SI)). To date, only Fe(OH)2þ has been identified as an important photoactive hydrolysis species of Fe(III), which can produce OH when irradiated with UV (Evans and Uri, 1949; Bates and Uri, 1953), whereas the photochemical activities of other hydrolysis species of Fe(III) are largely unknown. Furthermore, UV only accounts for a very small fraction of solar light. Thus, it is important to understand the photochemical activities of other hydrolysis species of Fe(III) under solar light. The objective of the present study was to further understand the sunlight-driven photochemical activities of hydrolysis species of Fe(III), as well as the underlying mechanisms. The phototransformation of bisphenol A (BPA)(Zuo and Zhu, 2014), a common endocrine disrupter and model compound for phototransformation studies, by Fe(III) under simulated solar light and natural sunlight was examined, in the absence of any external ligands. The important role of [Fe(OH)3]m, a previously unidentified photoactive Fe(III) species, was demonstrated, and the active radical involved was verified with both electron spin resonance (ESR) and a chemiluminescent probe. The fact that Fe(III) can be activated under solar light and thus produce strong radicals may have important environmental implications.
2. Materials and methods 2.1. Materials and reagents BPA (purity>99%), used as target compound, was purchased from Aldrich. Fe2(SO4)3 was purchased from Guangfu Fine Chemical Research Institute, Tianjin. Stock solution of BPA (20 mg L1) and Fe2(SO4)3 (112 mg L1 Fe(III)) were prepared by dissolving BPA or Fe2(SO4)3, respectively, in DI water (pH 6.0). 2.2. Thermodynamic modeling and z potential of aqueous Fe(III) Aqueous speciation of the medium was calculated using visual MINTEQ (version 3.1) to estimate complexation between iron and medium components under the conditions tested. Concentrations of Fe2(SO4)3 was 0.5 mM, and the temperature was 25 C, and the ionic strength was calculated from mass balance at the different pH value. Principal component analysis was used to identify significant iron species for further investigation. The z potential of aqueous Fe(III) related species after/before addition of NaCl for destroying probable colloidal parts was measured with a ZetaPALS (Brookhaven, American) at 25 C for exploring the relationship between z and photochemical performance, then discovered the function of the colloidal species from Fe(III). 2.3. Experimental procedures of photochemical reaction The photochemical degradation experiments were carried out in a XPA-7 photochemical reactor (Xujiang Electromechanical plant, Nanjing, China). Simulated sunlight irradiation was provided by a 500 W xenon lamp (Institute of Electric Light Source, Beijing) other than UV light (300 W or 500 W super-high pressure mercury lamp adopted by previous researches), which was positioned in the cylindrical quartz container. The reaction system was cooled by circulating water and maintained at room temperature. In quartz test tubes, 25 mL stock solution of Fe2(SO4)3 was added followed by adding 25 mL of BPA aqueous solution (pH~6.0). The solution was magnetically stirred during the reaction. Approximately 2 mL samples were taken at selected time interval and then filtered through 13 mm 0.22 mm membrane. Thus the solution was stored to be analyzed by High Performance Liquid Chromatography (HPLC) (Waters 1525). Toxicity of both BPA and corresponding intermediates were assessed via the luminescent bacteria using the acute toxicity detector (BHP9511, Hamamatsu Photonics, China). The freeze-dried luminescent bacterium P. phosphoreum were supplied by Institute of Soil Science, Chinese Academy of Sciences, and they were awaken by the treatment of 2% (w/v) NaCl solution. BPA in the reaction solution was measured by HPLC with the UV detector at 258 nm wavelength. The analytical conditions were followed: the column was 4.6 150 mm (Waters XBridge™ C18, 5 mm column); the mobile phase was acetonitrile/water 55:45 (v/ v); flow rate was set as 0.80 mL min1; column temperature was set at 40 C. The reaction intermediates of BPA were monitored and identified by ultra performance liquid chromatography-mass spectra (UPLC-MS) system (Waters Xevo TQ-S) with full scan (m/z 100e300) in positive electrospray ionization (ESI) mode with a dwell time of 400 m s. 2.4. EPR spectrosc and identification of involved opy and identification of involved radicals A Brucker EPR spectrometer (EMX-6/1) was used for measuring signals of free radicals and those spin-trapped by 5,5-dimethyl-1pyrroline-N-oxide (DMPO). The temperature was controlled with a standard temperature accessory and monitored before and after
M. Pan et al. / Chemosphere 167 (2017) 353e359
each measurement. 2-methy-6-[p-methoxyphenyl]-3,7 -dihydroimidazo[1,2-a] pyrazin-3-one (MCLA) was used as a chemiluminescent probe in consideration of the low concentration and short lifetime of O 2 , which selectively reacts with O2 in a purposebuilt detector called a flow injection analysis (FIA) system (Waterville Analytical, Waterville, ME). A 200 mM MCLA stock solution was prepared and subsequently stored at 4 C. The analytical reagent contained 1.0 mM MCLA buffered with 0.05 M sodium acetate, and the final pH of this reagent was adjusted to 5.5 using HCl. This reagent was prepared 1 day before using to ensure its stability. During the measurement, the MCLA reagent and samples were pumped into the flow cell where the reaction occurred. The reaction be tween O 2 and MCLA results in a chemiluminescent signal at 455 nm then was detected using a photomultiplier (PMT). 3. Results and discussion 3.1. Photochemical activity of Fe(III) Aqueous Fe(III) exhibited significant photochemical activities under simulated solar light, as indicated by the degradation of BPA (Fig. 1 and SI Fig. S1-3). In the presence of 0.5 mM Fe2(SO4)3 approximately 87% BPA was degraded in 16 h, whereas no apparent degradation was observed without solar light irradiation or in the absence of Fe2(SO4)3. Furthermore, the experiments conducted at different Fe(III) concentrations (0.05, 0.25, 0.5 and 1 mM) showed that the higher the Fe(III) concentration, the greater the BPA degradation (Fig. 1B). The reactions at different Fe(III) concentrations all followed pseudo-first-order reaction kinetic, and the fitted reaction rate constants are 0.0053, 0.020, 0.041 and 0.12 h1,
355
respectively. The intermediates of BPA transformation were monitored with UPLCeMS (Fig. 1C and Figs. S2 and S3), and the results show that BPA (m/z 227) was transformed into at least four intermediates (m/z ¼ 257, 241, 213 and 135). 3.2. Roles of different Fe(III) species in photodegradation of BPA In acidic and circumneutral aquatic environments, Fe(III) can exist as many different species (Xu et al., 2013), and it is important to understand which species were responsible for the photochemical activity of Fe(III) observed in this study. In the absence of external ligands, Fe(III) can form a series of hydrolysis species via the following pathways (Flynn, 1984): (a) primary hydrolysis giving rise to low-molecular-weight complexes (FeðOHÞ2þ , FeðOHÞþ 2, Fe2 ðOHÞ4þ 2 , etc.); (b) formation and aging of polynuclear polymers ð3nmÞþ ð3m2nxÞþ (Fen ðOHÞm ðH2 OÞx or Fem On ðOHÞx ); (c) precipitation of ferric oxides and hydroxides (amorphous Fe(OH)m, FeOOH, and Fe2O3) (see SI eq. S(1)S6 for detailed reactions). Moreover, the formation of colloidal Fe(OH)3 is also anticipated, and colloidal Fe(OH)3 can further aggregate to form Fe(OH)3 precipitate (Feng and Nansheng, 2000). The speciation of Fe(III) under the test conditions was analyzed using Visual MINTEQ (version 3.1) (Moberly et al., 2010; Ren et al., 2014), and the modeled species include six dissolved complexes, colloidal Fe(III), and precipitate (SI Fig. S5 and Table S1), consistent with the three pathways mentioned above. We verified the existence of colloidal Fe(III) with a simple Tyndall experiment, in which light scattering from the Tyndall effect was observed for aqueous Fe(III) (Fig. 2C and SI Fig. S5C). Through UPLCeMS, we also detected a series of ½FeðOHÞ3 m nFeOþ xH2 O and mðFe3þ Þ2 O3 nH2 O derived from colloidal Fe(III) with different m
Fig. 1. The Photochemical Reaction Condition in Photochemical Oxidation of BPA. A) Photochemical activity of samples (Fe2(SO4)3) in absence or presence of solar light. [Fe2(SO4)3] ¼ 0.5 mM, [BPA] ¼ 10 mg/L, T ¼ 25 C, and ambient pH, B) effect of initial concentration Fe2(SO4)3 on BPA removal and the corresponding kinetic simulation, C) UPLC-MS chromatogram (number, m/z) for the photochemical oxidation of BPA under simulated solar light irradiation.
356
M. Pan et al. / Chemosphere 167 (2017) 353e359
Fig. 2. The Roles of Fe3þ and Colloidal Fe(OH)3 Existence in Photochemical Oxidation of BPA Irradiated under Solar Light. A) Speciation of Fe3þ in aqueous solution disordered by different concentration NaCl simulated by visual MINTEQ, B) the effect of (NaCl) on the degradation of BPA, [NaCl] ¼ 10 mM, and C) Tyndall effects of different solution.
and n (SI Table S2). To understand the relative contributions of different Fe(III) species in the degradation of BPA, we added salt (NaCl) to disrupt the hydrolysis equilibrium of Fe(III). As demonstrated with Visual MINTEQ, the mass fractions of Fe(OH)2þ, Fe(OH)3, and [Fe(OH)3]m decreased with the increasing NaCl dose (Fig. 2A). Adding NaCl can inhibit colloid formation by compressing the electronic double layers, as indicated by the decrease of z potential (SI Fig. S5), and thus, result in decrease in concentrations of colloidal species such as [Fe(OH)3]m. Interestingly, degradation of BPA was significantly inhibited after adding 10 mM NaCl (Fig. 2B). Furthermore, the inhibited photochemical activities of Fe(III) correlated well with the amount of NaCl dose (SI Fig. S6). This indicates that besides of the previously identified photoactive Fe(OH)2þ, colloidal [Fe(OH)3]m may also be an important photoactive Fe(III) species. Moreover, when the colloidal Fe(III) and possible precipitates from colloidal Fe(III) were filtered out with 0.25-mm membrane filters (the mean hydrodynamic diameters of colloidal [Fe(OH)3]m were greater than 250 nm, SI Fig. S6), the photochemical activity of the resultant filtrate was reduced (SI Fig. S5D). This further indicates the contribution of colloidal [Fe(OH)3]m. 3.3. Radical formation mechanism and photochemical processes In previous studies Fe(OH)2þ has been considered the major photoactive hydrolysis species of Fe(III), and under acidic condition and UV irradiation OH is the primary radical formed (Evans and Uri, 1949; Bates and Uri, 1953), Fe(OH)2þ þ hv/ Fe(II) þ OH [Fe2(OH)2(H2O)4]
4þ
(1)
þ
þ H þ hv / Fe(II) þ Fe(III) þ H2O þ OH (2)
þ
4 Fe(II)þ O2 þ 4 H / 4 Fe(III) þ 2 H2O Fe(III) þ H2O / Fe(OH)
2þ
þ
þH
(3) (4)
To understand the radical formation mechanisms in this study, we used DMPO spin-trapped EPR spectroscopy to monitor the active radicals formed during the degradation processes. As shown in Fig. 3, the characteristic intensity 1:1:1:1:1:1 of DMPO-O-2 was observed in the methanol media (He and Goldsmith, 2012). The measured concentration of O 2 was ca. 6 nM. Furthermore, the characteristic intensity 1:2:2:1 of the DMPO-OH adducts were not observed, indicating that little OH was produced. Thus, it appears that Fe(OH)2þ was not the primary Fe(III) species responsible for the degradation of BPA under the test conditions. The detection of O 2 indicates that O2 was involved in the radical formation. Thus, we compared the degradation of BPA under O2rich and O2-deficit conditions. Degradation of BPA was almost completely inhibited when the solution was purged with Ar, whereas O2-enrichment accelerated the degradation of BPA (Fig. 3). Apparently, DO was a necessary factor for the significant photochemical oxidation of BPA in this study. Had Fe(OH)2þ been the main photoactive Fe(III) species for the oxidation of BPA (eqs. (1)e(4)), O2 would not be needed, in that ~0.39 mmol L1 OH could have been produced even in the absence of DO according (eq. (1)), and accordingly, 0.013e0.39 mmol/L BPA could have been transformed or mineralized. However, only 0.0066 mmol/L BPA (15%) was reduced in the absence of DO (Fig. 3A), indicating that the processes delineated in eqs. (1)e(4) are not the predominant mechanisms involved in this study. On the basis of the findings in this study, in particular, the observed important roles of colloidal Fe(III) species and O 2 , we propose that the high photochemical activity of Fe(III) is mainly exerted via the following mechanisms (Eqs. (5)e(9)), [Fe(OH)3]m þ 2O2 þ hv / Fe(II) þ 2O2-þ H2O
(5)
O 2 þ Fe(III) / Fe(II) þ O2
(6)
þ O 2 þ Fe(II) þ 2 H / Fe(III) þ H2O2
(7)
M. Pan et al. / Chemosphere 167 (2017) 353e359
357
Fig. 3. The Roles of Dissolved Oxygen and the Related Radicals in Photochemical Oxidation of BPA. A) Photochemical oxidation of BPA in absence or presence of DO, B) ESR signals of the DMPO-O-2 in methanol media and DMPO-OH in aqueous solution, respectively, in which filled squares and stars are the signs of DMPO and DMPO-O-2, respectively, and C) steady state O 2 concentrations in-situ.
Fe(II) þ Hþ þ H2O2 / Fe(III) þ OH
(8)
Fe(III) þ H2O2 þ OH / Fe(II) þ HO2
(9)
It has been reported that iron oxides or minerals, such as nano Fe3O4, can activate molecular oxygen and produce O 2 (Ardo et al., 2015) via the similar mechanisms mentioned above. Note that in the real water condition, radical transformation can occur among O 2 , OH, HO2 etc. following reactions shown in SI eqs. S(7)e(17). It is possible that these radicals collectively resulted in the photodegradation of BPA, as described in Scheme 1. 3.4. Photochemical activity of Fe(III) in natural sunlight We also tested the degradation of BPA by Fe(III) in natural sunlight (Fig. 4A). The UV-vis spectra (200e700 nm) of natural sunlight (Nankai University campus in Tianjin, 117:15 E 39:13 N, China, on June 6, 2014, recorded using a spectrometer (USBe4000, Ocean Optics Inc.)) and the simulated solar light used in this study are compared in Fig. 4B. Interestingly, 90% BPA was degraded in 10 h and as high as 95% removal of BPA was observed in 14 h, and the degradation rate even exceeded that observed under simulated solar light due to that natural sunlight showed higher light intensity than simulated sunlight, moreover, the former contains few UV light from 300 to 400 nm as indicated in Fig. 4B. Note that there was no noticeable removal during the night (from 8:00 p.m. to 8:00 a.m. the next day), further proving that significant transformation of BPA by Fe(III) can occur in natural environments. The fact that natural sunlight can activate naturally occurring Fe(III) species and produce oxidants such as O 2 further proves that Fe(III) species may significantly affect the fate, transport
Scheme 1. Schematic of Photochemical Mechanism of Fe3þ in Photochemical Oxidation of BPA Irradiated by Simulated Solar Light.
358
M. Pan et al. / Chemosphere 167 (2017) 353e359
spin resonance (ESR)das the key active radical responsible for the degradation of bisphenol A (BPA). This was because the colloidal [Fe(OH)3]m prefers to produce O 2 þ H O) whereas Fe(OH)2þ ([Fe(OH)3]m þ 2O2 þ hv / Fe(II) þ 2O 2 2 tends to yield OH primarily (FeOH2þ þ hv / Fe2þ þ OH). Briefly, iron ion in water should attract both environmental researchers and mangers more attention. Our researches maybe also introduce the novel process for oxidation and removal of organic pollutants from the point source iron-containing via irradiated by solar light with iron ion as green photocatalysts. Acknowledgments This work was supported by the National Natural Science Foundation of China (Grant 21577069), Key Project of Chinese National Programs for Fundamental Research and Development (973 Program) (Grant 2014CB932001), and Tianjin Municipal Science and Technology Commission (15JCZDJC40000). We thank Liping Yang for her assistance in analysis of UPLC-MS, Yanyan Gong for her assistance with MINTEQ analysis, and Yao Li for his assistant with O 2 measurement. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.chemosphere.2016.09.144. References
Fig. 4. The Roles of Solar light in Photochemical Oxidation of BPA. A) Function of (simulated) solar light, and B) the spectrum of the simulated solar light (500 W Xe lamp) and natural solar light (Tianjin, Jun-06, 2014) and strong UV light (500 W Hg light).
and effects of environmental contaminants. For example, transformation of organic contaminants by Fe(III) species may alter contaminant toxicity; in this study BPA was transformed to less toxic intermediates (SI Fig. S7). Findings of our study may also provide insights for the design of novel photochemical processes based on Fe(III) species. 4. Conclusions The natural sunlight can activate iron ion-related species and produce strong oxidants such as oxygen-involved radicals even without UV and ligands adopt by previous researchers, furthermore, these oxidants can oxidize organic pollutants and transform them. The important role of colloidal [Fe(OH)3]m as a core photocatalytic species from Fe(III) hydrolysis were explored and verified. Interestingly, O 2 , rather than OH, was identifieddvia electron
lieu, S., Ona-Nguema, G., Delarue, G., Brest, J., Pironin, E., Morin, G., 2015. Ardo, S., Ne Oxidative degradation of nalidixic acid by nano-magnetite via Fe2þ/O2Mediated reactions. Environ. Sci. Technol. 49, 4506e4514. Barry, D., Prommer, H., Miller, C., Engesgaard, P., Brun, A., Zheng, C., 2002. Modelling the fate of oxidisable organic contaminants in groundwater. Adv. Wat. Resour. 25, 945e983. Bates, H., Uri, N., 1953. Oxidation of aromatic compounds in aqueous solution by free radicals produced by photo-excited electron transfer in iron Complexes1. J. Am. Chem. Soc. 75, 2754e2759. Bigham, J., Nordstrom, D.K., 2000. Iron and aluminum hydroxysulfates from acid sulfate waters. Rev. Mineral. Geochem 40, 351e403. Chen, Y., Zhang, K., Zuo, Y.G., 2013. Direct and indirect photodegradation of estriol in the presence of humic acid, nitrate and iron complexes in water solutions. Sci. Total Environ. 463, 802e809. Conte, L.O., Querini, P., Albizzati, E.D., Alfano, O.M., 2014. Photonic and quantum efficiencies for the homogeneous photo-Fenton degradation of herbicide 2, 4-D using different iron complexes. J. Chem. Technol. Biotechnol. 89, 1967e1974. Cundy, A.B., Hopkinson, L., Whitby, R.L., 2008. Use of iron-based technologies in contaminated land and groundwater remediation: a review. Sci. Total Environ. 400, 42e51. Edwards, K.J., Bond, P.L., Druschel, G.K., McGuire, M.M., Hamers, R.J., Banfield, J.F., 2000. Geochemical and biological aspects of sulfide mineral dissolution: lessons from Iron Mountain, California. Chem. Geol. 169, 383e397. Evans, M., Uri, N., 1949. Photochemical polymerization in aqueous solution. Nature 164, 404, 404. Feng, W., Nansheng, D., 2000. Photochemistry of hydrolytic iron (III) species and photoinduced degradation of organic compounds. A minireview. Chemosphere 41, 1137e1147. Flynn Jr., C.M., 1984. Hydrolysis of inorganic iron (III) salts. Chem. Rev. 84, 31e41. Guo, H., Barnard, A.S., 2013. Naturally occurring iron oxide nanoparticles: morphology, surface chemistry and environmental stability. J. Mater. Chem. A 1, 27e42. He, Y., Goldsmith, C.R., 2012. Observation of a ferric hydroperoxide complex during the non-heme iron catalysed oxidation of alkenes and alkanes by O 2. Chem. Commun. 48, 10532e10534. Heijman, C.G., Grieder, E., Holliger, C., Schwarzenbach, R.P., 1995. Reduction of nitroaromatic compounds coupled to microbial iron reduction in laboratory aquifer columns. Environ. Sci. Technol. 29, 775e783. Kappler, A., Haderlein, S.B., 2003. Natural organic matter as reductant for chlorinated aliphatic pollutants. Environ. Sci. Technol. 37, 2714e2719. McKnight, D., Kimball, B., Bencala, K., 1988. Iron photoreduction and oxidation in an acidic mountain stream. Science 240, 637e640. Mitsunobu, S., Takahashi, Y., Terada, Y., Sakata, M., 2010. Antimony (V) incorporation into synthetic ferrihydrite, goethite, and natural iron oxyhydroxides. Environ. Sci. Technol. 44, 3712e3718. Moberly, J.G., Staven, A., Sani, R.K., Peyton, B.M., 2010. Influence of pH and inorganic phosphate on toxicity of zinc to Arthrobacter sp. isolated from heavy-metal-
M. Pan et al. / Chemosphere 167 (2017) 353e359 contaminated sediments. Environ. Sci. Technol. 44, 7302e7308. Nansheng, D., Feng, W., Fan, L., Mei, X., 1998. Ferric citrate-induced photodegradation of dyes in aqueous solutions. Chemosphere 36, 3101e3112. Nordstrom, D.K., Alpers, C.N., Ptacek, C.J., Blowes, D.W., 2000. Negative pH and extremely acidic mine waters from Iron Mountain, California. Environ. Sci. Technol. 34, 254e258. Ren, Z., Dong, Y., Liu, Y., 2014. Enhanced glyphosate removal by montmorillonite in the presence of Fe (III). Ind. Eng. Chem. Res. 53, 14485e14492. Rose, A.L., Waite, T.D., 2005. Reduction of organically complexed ferric iron by superoxide in a simulated natural water. Environ. Sci. Technol. 39, 2645e2650. Rose, S., Ghazi, A.M., 1997. Release of sorbed sulfate from iron oxyhydroxides precipitated from acid mine drainage associated with coal mining. Environ. Sci. Technol. 31, 2136e2140. Schnitzer, M., Hansen, E., 1970. Organo-metallic interactions in soils: 8. An evaluation of methods for the determination of stability constants of metal-fulvic acid complexes. Soil Sci. 109, 333e340. Sharp, E.L., Jarvis, P., Parsons, S.A., Jefferson, B., 2006. The impact of zeta potential on the physical properties of ferric-NOM flocs. Environ. Sci. Technol. 40, 3934e3940. Stumm, W., Morgan, J.J., 2012. Aquatic Chemistry: Chemical Equilibria and Rates in Natural Waters. John Wiley & Sons. Stumm, W., Sulzberger, B., 1992. The cycling of iron in natural environments: considerations based on laboratory studies of heterogeneous redox processes. Geochim. Cosmochim. Acta 56, 3233e3257. Van der Zee, F.P., Cervantes, F.J., 2009. Impact and application of electron shuttles on the redox (bio) transformation of contaminants: a review. Biotechnol. Adv. 27, 256e277. Watanabe, K., Manefield, M., Lee, M., Kouzuma, A., 2009. Electron shuttles in biotechnology. Curr. Opin. Biotechnol. 20, 633e641. Xu, J., Li, J., Wu, F., Zhang, Y., 2013. Rapid photooxidation of as (III) through surface
359
complexation with nascent colloidal ferric hydroxide. Environ. Sci. Technol. 48, 272e278. Zhang, W.x., Elliott, D.W., 2006. Applications of iron nanoparticles for groundwater remediation. Rem. J. 16, 7e21. Zhou, D., Wu, F., Deng, N., Xiang, W., 2004. Photooxidation of bisphenol A (BPA) in water in the presence of ferric and carboxylate salts. Water Res. 38, 4107e4116. Zuo, Y.G., 1995. Kinetics of photochemical chemical cycling of iron coupled with organic-substances in-cloud and fog droplets. Geochim. Cosmochim. Acta 59, 3123e3130. Zuo, Y.G., Deng, Y.W., 1997. Iron(II)-catalyzed photochemical decomposition of oxalic acid and generation of H2O2 in atmospheric liquid phases. Chemosphere 35, 2051e2058. Zuo, Y.G., Hoigne, J., 1992. Formation of hydrogen-peroxide and depletion of oxalicacid in atmospheric water by photolysis of iron(III) oxalato complexes. Environ. Sci. Technol. 26, 1014e1022. Zuo, Y.G., Hoigne, J., 1994. Photochemical decomposition of oxalic, glyoxalic and pyruvic-acid catalyzed by iron in atmospheric waters. Atmos. Environ. 28, 1231e1239. Zuo, Y.G., Zhan, J., 2005. Effects of oxalate on Fe-catalyzed photoxidation of dissolved sulfur dioxide in atmospheric water. Atmos. Environ. 39, 27e37. Zuo, Y.G., Zhan, J., Wu, T.X., 2005. Effects of monochromatic UV-Visible light and sunlight on Fe(III)-catalyzed oxidation of dissolved sulfur dioxide. JAtC 50, 195e210. Zuo, Y.G., Zhang, K., Zhou, S., 2013. Determination of estrogenic steroids and microbial and photochemical degradation of 17 alpha-ethinylestradiol (EE2) in lake surface water, a case study. Environ. Sci.-Process Impacts 15, 1529e1535. Zuo, Y.G., Zhu, Z., 2014. Simultaneous identification and quantification of 4cumylphenol, 2,4-bis-(dimethylbenzyl)phenol and bisphenol A in prawn Macrobrachium rosenbergii. Chemosphere 107, 447e453.
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Sunlight-driven photo-transformation of bisphenol A by Fe(III) in aqueous solution: Photochemical activity and mechanistic aspects Meilan Pan a, Jie Ding b, Lin Duan a, Guandao Gao b, * a
Ministry of Education Key Laboratory of Pollution Processes and Environmental Criteria, Tianjin Key Laboratory of Environmental Remediation and Pollution Control, College of Environmental Science and Engineering, Nankai University, Tianjin, 300071, China b State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, 210023, China
h i g h l i g h t s
g r a p h i c a l a b s t r a c t
Natural sunlight can activate iron ion-related species and produce strong oxidants. Colloidal [Fe(OH)3]m as a core photocatalytic species were explored. O 2 , rather than OH, was identified as the key active radical in our process. Our researches introduced the novel process for oxidation of organic pollutants.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 4 August 2016 Received in revised form 22 September 2016 Accepted 28 September 2016 Available online 12 October 2016
Iron is one of the most abundant elements in aquatic environments, and plays important roles in the fate and transport of environmental contaminants. Previous studies on the photochemical properties of Fe(III) species have largely focused on complexes formed between Fe(III) and environmental ligands such as natural organic matter (NOM) under UV irradiation, whereas the potentially important roles of hydrolysis species of Fe(III) in Fe(III)-mediated photo-transformation of environmental contaminants under solar light are not fully understood. In this study, the solar light-driven photochemical activities of hydrolysis species of Fe(III) were further explored, using a system containing only 0.5 mM Fe2(SO4)3 and bisphenol A. The important role of colloidal [Fe(OH)3]m, formed from the hydrolysis of Fe3þ, as a core photochemical species of Fe(III) was proposed and verified. Interestingly, O 2 , rather than OH, was identified (via electron spin resonance) as the key active radical responsible for the degradation of bisphenol A. We propose that unlike Fe(OH)2þ, which under UV irradiation can yield OH even in sunlight (Fe(OH)2þ þ hv / Fe2þ þ OH), colloidal [Fe(OH)3]m produces O 2 ([Fe(OH)3]m þ 2O2 þ hv / Fe(II) þ 2O 2 þ H2O). The fact that Fe(III) can produce strong radicals in sunlight may have important environmental implications. © 2016 Published by Elsevier Ltd.
Handling Editor: Jun Huang Keywords: Sunlight-driven Photochemical activity Colloidal [Fe(OH)3]m Bisphenol A
1. Introduction * Corresponding author. School of the Environment, Nanjing University, Nanjing, 210023, China. E-mail address: [email protected] (G. Gao). http://dx.doi.org/10.1016/j.chemosphere.2016.09.144 0045-6535/© 2016 Published by Elsevier Ltd.
et
Iron is one of the most abundant elements on earth (McKnight al., 1988; Bigham and Nordstrom, 2000). In aquatic
354
M. Pan et al. / Chemosphere 167 (2017) 353e359
environments, iron is produced naturally from weathering of iron minerals such as magnetite, hematite, and goethite (Stumm and Sulzberger, 1992; Mitsunobu et al., 2010; Guo and Barnard, 2013), or released as the results of iron mine exploiting and broad applications in industrial fields (Rose and Ghazi, 1997; Edwards et al., 2000; Nordstrom et al., 2000). A number of studies have shown that iron plays important roles in the fate and transport of environmental contaminants in surface and ground waters (Barry et al., 2002; Zhang and Elliott, 2006; Cundy et al., 2008) by participating directly in redox and photochemical reactions(Zuo and Hoigne, 1992; Zuo, 1995; Zuo and Deng, 1997; Chen et al., 2013) or serving as electron mediators in the biodegradation of organic pollutants (Heijman et al., 1995; Kappler and Haderlein, 2003; Van der Zee and Cervantes, 2009; Watanabe et al., 2009). It has been shown that ferric ion, i.e., Fe(III), can oxidize organic compounds in acidic aqueous solutions (pH z 3) when irradiated with UV light, and OH is responsible for the photochemical activity (Evans and Uri, 1949; Bates and Uri, 1953). Previous studies on the photochemical properties of different Fe(III) species have emphasized on the roles of complexes formed between Fe(III) and ligands(Zuo and Hoigne, 1994; Zuo and Zhan, 2005; Zuo et al., 2005, 2013) such as citrate (Nansheng et al., 1998; Conte et al., 2014), oxalate (Zhou et al., 2004; Conte et al., 2014), or natural organic matter (NOM) (Rose and Waite, 2005; Sharp et al., 2006) under UV irradiation. It was proposed that ligand-to-metal charge transition and electron transfer from the ligand to the central metal ion could occur in such complexes and consequently result in the photochemical activity of Fe(III). In aqueous solution Fe(III) has a strong tendency to undergo hydrolysis and form complexes with OH. In fact, the logarithm complex constants between Fe(III) and OH range from 11.9 to 29.3 (Schnitzer and Hansen, 1970; Stumm and Morgan, 2012), much greater than those of Fe(III)NOM complexes (6.1) and the complexes between Fe(III) and model NOMs such as succinic acid (7.5) and citric acid (7.2) (Schnitzer and Hansen, 1970; Stumm and Morgan, 2012). Thus, hydrolysis species of Fe(III), whose absorbance coefficients and quantum yields are comparable to those of FeNOM (Schnitzer and Hansen, 1970; Stumm and Morgan, 2012), can often be the dominant Fe(III) species (Schnitzer and Hansen, 1970; Stumm and Morgan, 2012) even in the presence of relatively high concentration of NOM (10e100 mg/L) (see the results of Visual MINTEQ 3.0 modeling in Supporting Information (SI)). To date, only Fe(OH)2þ has been identified as an important photoactive hydrolysis species of Fe(III), which can produce OH when irradiated with UV (Evans and Uri, 1949; Bates and Uri, 1953), whereas the photochemical activities of other hydrolysis species of Fe(III) are largely unknown. Furthermore, UV only accounts for a very small fraction of solar light. Thus, it is important to understand the photochemical activities of other hydrolysis species of Fe(III) under solar light. The objective of the present study was to further understand the sunlight-driven photochemical activities of hydrolysis species of Fe(III), as well as the underlying mechanisms. The phototransformation of bisphenol A (BPA)(Zuo and Zhu, 2014), a common endocrine disrupter and model compound for phototransformation studies, by Fe(III) under simulated solar light and natural sunlight was examined, in the absence of any external ligands. The important role of [Fe(OH)3]m, a previously unidentified photoactive Fe(III) species, was demonstrated, and the active radical involved was verified with both electron spin resonance (ESR) and a chemiluminescent probe. The fact that Fe(III) can be activated under solar light and thus produce strong radicals may have important environmental implications.
2. Materials and methods 2.1. Materials and reagents BPA (purity>99%), used as target compound, was purchased from Aldrich. Fe2(SO4)3 was purchased from Guangfu Fine Chemical Research Institute, Tianjin. Stock solution of BPA (20 mg L1) and Fe2(SO4)3 (112 mg L1 Fe(III)) were prepared by dissolving BPA or Fe2(SO4)3, respectively, in DI water (pH 6.0). 2.2. Thermodynamic modeling and z potential of aqueous Fe(III) Aqueous speciation of the medium was calculated using visual MINTEQ (version 3.1) to estimate complexation between iron and medium components under the conditions tested. Concentrations of Fe2(SO4)3 was 0.5 mM, and the temperature was 25 C, and the ionic strength was calculated from mass balance at the different pH value. Principal component analysis was used to identify significant iron species for further investigation. The z potential of aqueous Fe(III) related species after/before addition of NaCl for destroying probable colloidal parts was measured with a ZetaPALS (Brookhaven, American) at 25 C for exploring the relationship between z and photochemical performance, then discovered the function of the colloidal species from Fe(III). 2.3. Experimental procedures of photochemical reaction The photochemical degradation experiments were carried out in a XPA-7 photochemical reactor (Xujiang Electromechanical plant, Nanjing, China). Simulated sunlight irradiation was provided by a 500 W xenon lamp (Institute of Electric Light Source, Beijing) other than UV light (300 W or 500 W super-high pressure mercury lamp adopted by previous researches), which was positioned in the cylindrical quartz container. The reaction system was cooled by circulating water and maintained at room temperature. In quartz test tubes, 25 mL stock solution of Fe2(SO4)3 was added followed by adding 25 mL of BPA aqueous solution (pH~6.0). The solution was magnetically stirred during the reaction. Approximately 2 mL samples were taken at selected time interval and then filtered through 13 mm 0.22 mm membrane. Thus the solution was stored to be analyzed by High Performance Liquid Chromatography (HPLC) (Waters 1525). Toxicity of both BPA and corresponding intermediates were assessed via the luminescent bacteria using the acute toxicity detector (BHP9511, Hamamatsu Photonics, China). The freeze-dried luminescent bacterium P. phosphoreum were supplied by Institute of Soil Science, Chinese Academy of Sciences, and they were awaken by the treatment of 2% (w/v) NaCl solution. BPA in the reaction solution was measured by HPLC with the UV detector at 258 nm wavelength. The analytical conditions were followed: the column was 4.6 150 mm (Waters XBridge™ C18, 5 mm column); the mobile phase was acetonitrile/water 55:45 (v/ v); flow rate was set as 0.80 mL min1; column temperature was set at 40 C. The reaction intermediates of BPA were monitored and identified by ultra performance liquid chromatography-mass spectra (UPLC-MS) system (Waters Xevo TQ-S) with full scan (m/z 100e300) in positive electrospray ionization (ESI) mode with a dwell time of 400 m s. 2.4. EPR spectrosc and identification of involved opy and identification of involved radicals A Brucker EPR spectrometer (EMX-6/1) was used for measuring signals of free radicals and those spin-trapped by 5,5-dimethyl-1pyrroline-N-oxide (DMPO). The temperature was controlled with a standard temperature accessory and monitored before and after
M. Pan et al. / Chemosphere 167 (2017) 353e359
each measurement. 2-methy-6-[p-methoxyphenyl]-3,7 -dihydroimidazo[1,2-a] pyrazin-3-one (MCLA) was used as a chemiluminescent probe in consideration of the low concentration and short lifetime of O 2 , which selectively reacts with O2 in a purposebuilt detector called a flow injection analysis (FIA) system (Waterville Analytical, Waterville, ME). A 200 mM MCLA stock solution was prepared and subsequently stored at 4 C. The analytical reagent contained 1.0 mM MCLA buffered with 0.05 M sodium acetate, and the final pH of this reagent was adjusted to 5.5 using HCl. This reagent was prepared 1 day before using to ensure its stability. During the measurement, the MCLA reagent and samples were pumped into the flow cell where the reaction occurred. The reaction be tween O 2 and MCLA results in a chemiluminescent signal at 455 nm then was detected using a photomultiplier (PMT). 3. Results and discussion 3.1. Photochemical activity of Fe(III) Aqueous Fe(III) exhibited significant photochemical activities under simulated solar light, as indicated by the degradation of BPA (Fig. 1 and SI Fig. S1-3). In the presence of 0.5 mM Fe2(SO4)3 approximately 87% BPA was degraded in 16 h, whereas no apparent degradation was observed without solar light irradiation or in the absence of Fe2(SO4)3. Furthermore, the experiments conducted at different Fe(III) concentrations (0.05, 0.25, 0.5 and 1 mM) showed that the higher the Fe(III) concentration, the greater the BPA degradation (Fig. 1B). The reactions at different Fe(III) concentrations all followed pseudo-first-order reaction kinetic, and the fitted reaction rate constants are 0.0053, 0.020, 0.041 and 0.12 h1,
355
respectively. The intermediates of BPA transformation were monitored with UPLCeMS (Fig. 1C and Figs. S2 and S3), and the results show that BPA (m/z 227) was transformed into at least four intermediates (m/z ¼ 257, 241, 213 and 135). 3.2. Roles of different Fe(III) species in photodegradation of BPA In acidic and circumneutral aquatic environments, Fe(III) can exist as many different species (Xu et al., 2013), and it is important to understand which species were responsible for the photochemical activity of Fe(III) observed in this study. In the absence of external ligands, Fe(III) can form a series of hydrolysis species via the following pathways (Flynn, 1984): (a) primary hydrolysis giving rise to low-molecular-weight complexes (FeðOHÞ2þ , FeðOHÞþ 2, Fe2 ðOHÞ4þ 2 , etc.); (b) formation and aging of polynuclear polymers ð3nmÞþ ð3m2nxÞþ (Fen ðOHÞm ðH2 OÞx or Fem On ðOHÞx ); (c) precipitation of ferric oxides and hydroxides (amorphous Fe(OH)m, FeOOH, and Fe2O3) (see SI eq. S(1)S6 for detailed reactions). Moreover, the formation of colloidal Fe(OH)3 is also anticipated, and colloidal Fe(OH)3 can further aggregate to form Fe(OH)3 precipitate (Feng and Nansheng, 2000). The speciation of Fe(III) under the test conditions was analyzed using Visual MINTEQ (version 3.1) (Moberly et al., 2010; Ren et al., 2014), and the modeled species include six dissolved complexes, colloidal Fe(III), and precipitate (SI Fig. S5 and Table S1), consistent with the three pathways mentioned above. We verified the existence of colloidal Fe(III) with a simple Tyndall experiment, in which light scattering from the Tyndall effect was observed for aqueous Fe(III) (Fig. 2C and SI Fig. S5C). Through UPLCeMS, we also detected a series of ½FeðOHÞ3 m nFeOþ xH2 O and mðFe3þ Þ2 O3 nH2 O derived from colloidal Fe(III) with different m
Fig. 1. The Photochemical Reaction Condition in Photochemical Oxidation of BPA. A) Photochemical activity of samples (Fe2(SO4)3) in absence or presence of solar light. [Fe2(SO4)3] ¼ 0.5 mM, [BPA] ¼ 10 mg/L, T ¼ 25 C, and ambient pH, B) effect of initial concentration Fe2(SO4)3 on BPA removal and the corresponding kinetic simulation, C) UPLC-MS chromatogram (number, m/z) for the photochemical oxidation of BPA under simulated solar light irradiation.
356
M. Pan et al. / Chemosphere 167 (2017) 353e359
Fig. 2. The Roles of Fe3þ and Colloidal Fe(OH)3 Existence in Photochemical Oxidation of BPA Irradiated under Solar Light. A) Speciation of Fe3þ in aqueous solution disordered by different concentration NaCl simulated by visual MINTEQ, B) the effect of (NaCl) on the degradation of BPA, [NaCl] ¼ 10 mM, and C) Tyndall effects of different solution.
and n (SI Table S2). To understand the relative contributions of different Fe(III) species in the degradation of BPA, we added salt (NaCl) to disrupt the hydrolysis equilibrium of Fe(III). As demonstrated with Visual MINTEQ, the mass fractions of Fe(OH)2þ, Fe(OH)3, and [Fe(OH)3]m decreased with the increasing NaCl dose (Fig. 2A). Adding NaCl can inhibit colloid formation by compressing the electronic double layers, as indicated by the decrease of z potential (SI Fig. S5), and thus, result in decrease in concentrations of colloidal species such as [Fe(OH)3]m. Interestingly, degradation of BPA was significantly inhibited after adding 10 mM NaCl (Fig. 2B). Furthermore, the inhibited photochemical activities of Fe(III) correlated well with the amount of NaCl dose (SI Fig. S6). This indicates that besides of the previously identified photoactive Fe(OH)2þ, colloidal [Fe(OH)3]m may also be an important photoactive Fe(III) species. Moreover, when the colloidal Fe(III) and possible precipitates from colloidal Fe(III) were filtered out with 0.25-mm membrane filters (the mean hydrodynamic diameters of colloidal [Fe(OH)3]m were greater than 250 nm, SI Fig. S6), the photochemical activity of the resultant filtrate was reduced (SI Fig. S5D). This further indicates the contribution of colloidal [Fe(OH)3]m. 3.3. Radical formation mechanism and photochemical processes In previous studies Fe(OH)2þ has been considered the major photoactive hydrolysis species of Fe(III), and under acidic condition and UV irradiation OH is the primary radical formed (Evans and Uri, 1949; Bates and Uri, 1953), Fe(OH)2þ þ hv/ Fe(II) þ OH [Fe2(OH)2(H2O)4]
4þ
(1)
þ
þ H þ hv / Fe(II) þ Fe(III) þ H2O þ OH (2)
þ
4 Fe(II)þ O2 þ 4 H / 4 Fe(III) þ 2 H2O Fe(III) þ H2O / Fe(OH)
2þ
þ
þH
(3) (4)
To understand the radical formation mechanisms in this study, we used DMPO spin-trapped EPR spectroscopy to monitor the active radicals formed during the degradation processes. As shown in Fig. 3, the characteristic intensity 1:1:1:1:1:1 of DMPO-O-2 was observed in the methanol media (He and Goldsmith, 2012). The measured concentration of O 2 was ca. 6 nM. Furthermore, the characteristic intensity 1:2:2:1 of the DMPO-OH adducts were not observed, indicating that little OH was produced. Thus, it appears that Fe(OH)2þ was not the primary Fe(III) species responsible for the degradation of BPA under the test conditions. The detection of O 2 indicates that O2 was involved in the radical formation. Thus, we compared the degradation of BPA under O2rich and O2-deficit conditions. Degradation of BPA was almost completely inhibited when the solution was purged with Ar, whereas O2-enrichment accelerated the degradation of BPA (Fig. 3). Apparently, DO was a necessary factor for the significant photochemical oxidation of BPA in this study. Had Fe(OH)2þ been the main photoactive Fe(III) species for the oxidation of BPA (eqs. (1)e(4)), O2 would not be needed, in that ~0.39 mmol L1 OH could have been produced even in the absence of DO according (eq. (1)), and accordingly, 0.013e0.39 mmol/L BPA could have been transformed or mineralized. However, only 0.0066 mmol/L BPA (15%) was reduced in the absence of DO (Fig. 3A), indicating that the processes delineated in eqs. (1)e(4) are not the predominant mechanisms involved in this study. On the basis of the findings in this study, in particular, the observed important roles of colloidal Fe(III) species and O 2 , we propose that the high photochemical activity of Fe(III) is mainly exerted via the following mechanisms (Eqs. (5)e(9)), [Fe(OH)3]m þ 2O2 þ hv / Fe(II) þ 2O2-þ H2O
(5)
O 2 þ Fe(III) / Fe(II) þ O2
(6)
þ O 2 þ Fe(II) þ 2 H / Fe(III) þ H2O2
(7)
M. Pan et al. / Chemosphere 167 (2017) 353e359
357
Fig. 3. The Roles of Dissolved Oxygen and the Related Radicals in Photochemical Oxidation of BPA. A) Photochemical oxidation of BPA in absence or presence of DO, B) ESR signals of the DMPO-O-2 in methanol media and DMPO-OH in aqueous solution, respectively, in which filled squares and stars are the signs of DMPO and DMPO-O-2, respectively, and C) steady state O 2 concentrations in-situ.
Fe(II) þ Hþ þ H2O2 / Fe(III) þ OH
(8)
Fe(III) þ H2O2 þ OH / Fe(II) þ HO2
(9)
It has been reported that iron oxides or minerals, such as nano Fe3O4, can activate molecular oxygen and produce O 2 (Ardo et al., 2015) via the similar mechanisms mentioned above. Note that in the real water condition, radical transformation can occur among O 2 , OH, HO2 etc. following reactions shown in SI eqs. S(7)e(17). It is possible that these radicals collectively resulted in the photodegradation of BPA, as described in Scheme 1. 3.4. Photochemical activity of Fe(III) in natural sunlight We also tested the degradation of BPA by Fe(III) in natural sunlight (Fig. 4A). The UV-vis spectra (200e700 nm) of natural sunlight (Nankai University campus in Tianjin, 117:15 E 39:13 N, China, on June 6, 2014, recorded using a spectrometer (USBe4000, Ocean Optics Inc.)) and the simulated solar light used in this study are compared in Fig. 4B. Interestingly, 90% BPA was degraded in 10 h and as high as 95% removal of BPA was observed in 14 h, and the degradation rate even exceeded that observed under simulated solar light due to that natural sunlight showed higher light intensity than simulated sunlight, moreover, the former contains few UV light from 300 to 400 nm as indicated in Fig. 4B. Note that there was no noticeable removal during the night (from 8:00 p.m. to 8:00 a.m. the next day), further proving that significant transformation of BPA by Fe(III) can occur in natural environments. The fact that natural sunlight can activate naturally occurring Fe(III) species and produce oxidants such as O 2 further proves that Fe(III) species may significantly affect the fate, transport
Scheme 1. Schematic of Photochemical Mechanism of Fe3þ in Photochemical Oxidation of BPA Irradiated by Simulated Solar Light.
358
M. Pan et al. / Chemosphere 167 (2017) 353e359
spin resonance (ESR)das the key active radical responsible for the degradation of bisphenol A (BPA). This was because the colloidal [Fe(OH)3]m prefers to produce O 2 þ H O) whereas Fe(OH)2þ ([Fe(OH)3]m þ 2O2 þ hv / Fe(II) þ 2O 2 2 tends to yield OH primarily (FeOH2þ þ hv / Fe2þ þ OH). Briefly, iron ion in water should attract both environmental researchers and mangers more attention. Our researches maybe also introduce the novel process for oxidation and removal of organic pollutants from the point source iron-containing via irradiated by solar light with iron ion as green photocatalysts. Acknowledgments This work was supported by the National Natural Science Foundation of China (Grant 21577069), Key Project of Chinese National Programs for Fundamental Research and Development (973 Program) (Grant 2014CB932001), and Tianjin Municipal Science and Technology Commission (15JCZDJC40000). We thank Liping Yang for her assistance in analysis of UPLC-MS, Yanyan Gong for her assistance with MINTEQ analysis, and Yao Li for his assistant with O 2 measurement. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.chemosphere.2016.09.144. References
Fig. 4. The Roles of Solar light in Photochemical Oxidation of BPA. A) Function of (simulated) solar light, and B) the spectrum of the simulated solar light (500 W Xe lamp) and natural solar light (Tianjin, Jun-06, 2014) and strong UV light (500 W Hg light).
and effects of environmental contaminants. For example, transformation of organic contaminants by Fe(III) species may alter contaminant toxicity; in this study BPA was transformed to less toxic intermediates (SI Fig. S7). Findings of our study may also provide insights for the design of novel photochemical processes based on Fe(III) species. 4. Conclusions The natural sunlight can activate iron ion-related species and produce strong oxidants such as oxygen-involved radicals even without UV and ligands adopt by previous researchers, furthermore, these oxidants can oxidize organic pollutants and transform them. The important role of colloidal [Fe(OH)3]m as a core photocatalytic species from Fe(III) hydrolysis were explored and verified. Interestingly, O 2 , rather than OH, was identifieddvia electron
lieu, S., Ona-Nguema, G., Delarue, G., Brest, J., Pironin, E., Morin, G., 2015. Ardo, S., Ne Oxidative degradation of nalidixic acid by nano-magnetite via Fe2þ/O2Mediated reactions. Environ. Sci. Technol. 49, 4506e4514. Barry, D., Prommer, H., Miller, C., Engesgaard, P., Brun, A., Zheng, C., 2002. Modelling the fate of oxidisable organic contaminants in groundwater. Adv. Wat. Resour. 25, 945e983. Bates, H., Uri, N., 1953. Oxidation of aromatic compounds in aqueous solution by free radicals produced by photo-excited electron transfer in iron Complexes1. J. Am. Chem. Soc. 75, 2754e2759. Bigham, J., Nordstrom, D.K., 2000. Iron and aluminum hydroxysulfates from acid sulfate waters. Rev. Mineral. Geochem 40, 351e403. Chen, Y., Zhang, K., Zuo, Y.G., 2013. Direct and indirect photodegradation of estriol in the presence of humic acid, nitrate and iron complexes in water solutions. Sci. Total Environ. 463, 802e809. Conte, L.O., Querini, P., Albizzati, E.D., Alfano, O.M., 2014. Photonic and quantum efficiencies for the homogeneous photo-Fenton degradation of herbicide 2, 4-D using different iron complexes. J. Chem. Technol. Biotechnol. 89, 1967e1974. Cundy, A.B., Hopkinson, L., Whitby, R.L., 2008. Use of iron-based technologies in contaminated land and groundwater remediation: a review. Sci. Total Environ. 400, 42e51. Edwards, K.J., Bond, P.L., Druschel, G.K., McGuire, M.M., Hamers, R.J., Banfield, J.F., 2000. Geochemical and biological aspects of sulfide mineral dissolution: lessons from Iron Mountain, California. Chem. Geol. 169, 383e397. Evans, M., Uri, N., 1949. Photochemical polymerization in aqueous solution. Nature 164, 404, 404. Feng, W., Nansheng, D., 2000. Photochemistry of hydrolytic iron (III) species and photoinduced degradation of organic compounds. A minireview. Chemosphere 41, 1137e1147. Flynn Jr., C.M., 1984. Hydrolysis of inorganic iron (III) salts. Chem. Rev. 84, 31e41. Guo, H., Barnard, A.S., 2013. Naturally occurring iron oxide nanoparticles: morphology, surface chemistry and environmental stability. J. Mater. Chem. A 1, 27e42. He, Y., Goldsmith, C.R., 2012. Observation of a ferric hydroperoxide complex during the non-heme iron catalysed oxidation of alkenes and alkanes by O 2. Chem. Commun. 48, 10532e10534. Heijman, C.G., Grieder, E., Holliger, C., Schwarzenbach, R.P., 1995. Reduction of nitroaromatic compounds coupled to microbial iron reduction in laboratory aquifer columns. Environ. Sci. Technol. 29, 775e783. Kappler, A., Haderlein, S.B., 2003. Natural organic matter as reductant for chlorinated aliphatic pollutants. Environ. Sci. Technol. 37, 2714e2719. McKnight, D., Kimball, B., Bencala, K., 1988. Iron photoreduction and oxidation in an acidic mountain stream. Science 240, 637e640. Mitsunobu, S., Takahashi, Y., Terada, Y., Sakata, M., 2010. Antimony (V) incorporation into synthetic ferrihydrite, goethite, and natural iron oxyhydroxides. Environ. Sci. Technol. 44, 3712e3718. Moberly, J.G., Staven, A., Sani, R.K., Peyton, B.M., 2010. Influence of pH and inorganic phosphate on toxicity of zinc to Arthrobacter sp. isolated from heavy-metal-
M. Pan et al. / Chemosphere 167 (2017) 353e359 contaminated sediments. Environ. Sci. Technol. 44, 7302e7308. Nansheng, D., Feng, W., Fan, L., Mei, X., 1998. Ferric citrate-induced photodegradation of dyes in aqueous solutions. Chemosphere 36, 3101e3112. Nordstrom, D.K., Alpers, C.N., Ptacek, C.J., Blowes, D.W., 2000. Negative pH and extremely acidic mine waters from Iron Mountain, California. Environ. Sci. Technol. 34, 254e258. Ren, Z., Dong, Y., Liu, Y., 2014. Enhanced glyphosate removal by montmorillonite in the presence of Fe (III). Ind. Eng. Chem. Res. 53, 14485e14492. Rose, A.L., Waite, T.D., 2005. Reduction of organically complexed ferric iron by superoxide in a simulated natural water. Environ. Sci. Technol. 39, 2645e2650. Rose, S., Ghazi, A.M., 1997. Release of sorbed sulfate from iron oxyhydroxides precipitated from acid mine drainage associated with coal mining. Environ. Sci. Technol. 31, 2136e2140. Schnitzer, M., Hansen, E., 1970. Organo-metallic interactions in soils: 8. An evaluation of methods for the determination of stability constants of metal-fulvic acid complexes. Soil Sci. 109, 333e340. Sharp, E.L., Jarvis, P., Parsons, S.A., Jefferson, B., 2006. The impact of zeta potential on the physical properties of ferric-NOM flocs. Environ. Sci. Technol. 40, 3934e3940. Stumm, W., Morgan, J.J., 2012. Aquatic Chemistry: Chemical Equilibria and Rates in Natural Waters. John Wiley & Sons. Stumm, W., Sulzberger, B., 1992. The cycling of iron in natural environments: considerations based on laboratory studies of heterogeneous redox processes. Geochim. Cosmochim. Acta 56, 3233e3257. Van der Zee, F.P., Cervantes, F.J., 2009. Impact and application of electron shuttles on the redox (bio) transformation of contaminants: a review. Biotechnol. Adv. 27, 256e277. Watanabe, K., Manefield, M., Lee, M., Kouzuma, A., 2009. Electron shuttles in biotechnology. Curr. Opin. Biotechnol. 20, 633e641. Xu, J., Li, J., Wu, F., Zhang, Y., 2013. Rapid photooxidation of as (III) through surface
359
complexation with nascent colloidal ferric hydroxide. Environ. Sci. Technol. 48, 272e278. Zhang, W.x., Elliott, D.W., 2006. Applications of iron nanoparticles for groundwater remediation. Rem. J. 16, 7e21. Zhou, D., Wu, F., Deng, N., Xiang, W., 2004. Photooxidation of bisphenol A (BPA) in water in the presence of ferric and carboxylate salts. Water Res. 38, 4107e4116. Zuo, Y.G., 1995. Kinetics of photochemical chemical cycling of iron coupled with organic-substances in-cloud and fog droplets. Geochim. Cosmochim. Acta 59, 3123e3130. Zuo, Y.G., Deng, Y.W., 1997. Iron(II)-catalyzed photochemical decomposition of oxalic acid and generation of H2O2 in atmospheric liquid phases. Chemosphere 35, 2051e2058. Zuo, Y.G., Hoigne, J., 1992. Formation of hydrogen-peroxide and depletion of oxalicacid in atmospheric water by photolysis of iron(III) oxalato complexes. Environ. Sci. Technol. 26, 1014e1022. Zuo, Y.G., Hoigne, J., 1994. Photochemical decomposition of oxalic, glyoxalic and pyruvic-acid catalyzed by iron in atmospheric waters. Atmos. Environ. 28, 1231e1239. Zuo, Y.G., Zhan, J., 2005. Effects of oxalate on Fe-catalyzed photoxidation of dissolved sulfur dioxide in atmospheric water. Atmos. Environ. 39, 27e37. Zuo, Y.G., Zhan, J., Wu, T.X., 2005. Effects of monochromatic UV-Visible light and sunlight on Fe(III)-catalyzed oxidation of dissolved sulfur dioxide. JAtC 50, 195e210. Zuo, Y.G., Zhang, K., Zhou, S., 2013. Determination of estrogenic steroids and microbial and photochemical degradation of 17 alpha-ethinylestradiol (EE2) in lake surface water, a case study. Environ. Sci.-Process Impacts 15, 1529e1535. Zuo, Y.G., Zhu, Z., 2014. Simultaneous identification and quantification of 4cumylphenol, 2,4-bis-(dimethylbenzyl)phenol and bisphenol A in prawn Macrobrachium rosenbergii. Chemosphere 107, 447e453.