The p-arsanilic acid (p-ASA) is widely used in agriculture as a food additive to control parasites. It leaves the body almost unchanged and is subsequently destroyed by environmental factors with the formation of toxic forms of inorganic arsenic. UVA irradiation of p-ASA with the addition of the Fe(III) oxalate complex leads to an effective degradation of the target compound. However, this method needs high concentrations of reagents and the keeping high [oxalate]:[Fe(III)] ratio to maintain proper efficiency of Fe(III) oxalate system. In this work, to overcome these problems, potassium persulfate (PS) was used as an additional oxidizer to improve Fe(III) oxalate system. It was found that the sulfate radical produced upon PS activation reacts readily with both neutral and monoanionic forms of p-ASA yielding corresponding organic cation radical, bimolecular rate constants are (7.3 ± 0.6)×109 and (2.4 ± 0.4)×109 M− 1s− 1, accordingly. Addition of PS allows one to reduce the working concentration of oxalate and achieve the complete degradation of both p-ASA and organic byproducts to the less toxic inorganic As(V) up to 3 ppm of initial p-ASA concentration. Also the proposed approach demonstrates efficiency even for low (< 0.5 ppm) concentrations of p-ASA.
Research Article
Synergetic Effect of Potassium Persulfate and Iron Oxalate on Photodegradation of Para-Arsanilic Acid
https://doi.org/10.21203/rs.3.rs-325135/v1
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ferrioxalate complex
p-ASA
AOPs
hydroxyl radical
sulfate radical
photodegradation
laser flash photolysis
The widespread use of p-ASA in agriculture has led to the water contamination with inorganic arsenic (As(III) and As(V)), as well as p-ASA itself (Jones 2007, Mangalgiri et al, 2015, Silbergeld and Nachman 2008, Yao et al., 2013). Since p-ASA leaves the body of livestock in almost unchanged form, it is destroyed in the water under the influence of external factors, like pH, temperature and light (Arroyo-Abad et al., 2012, Cortinas et al., 2006, Makris et al., 2008). As inorganic forms of arsenic are toxic to the body, the methods of water purification from p-ASA and similar compounds need to be addressed.
Fe(III) carboxylates are the natural photoactive compounds demonstrating efficient generation of reactive oxygen species (ROS, hydroxyl radicals, mainly) under the action of solar or artificial UV radiation (Faust and Zepp, 1993; Wu and Deng, 2000; Wan et al., 2019; Zuo and Hoigne, 1994). So these complexes are perspective for application in the advanced oxidation processes (AOPs). It has been demonstrated that the combined action of UVA light and Fe (III) - carboxylates could be used for effective degradation of many persistent pollutants, including bisphenols, herbicides, pharmaceuticals, etc. (Chen et al., 2013; Zhong and Yun, 2005; Pozdnyakov et al., 2018; Wan et al., 2018; Wan et al., 2019; Zhou et al., 2004).
Regeneration of active Fe(III) form during the photolysis of Fe(III)-oxalate complexes allows some researches to call such systems as “photocatalytical” though in fact their functioning demands the constant consumption of oxalate ligands as a “sacrificial agent”. This is first, but not the last drawback of Fe(III) oxalate system (Pozdnyakov et al., 2018). The second, photochemical reactions usually require relatively high concentrations of reagents ([Fe(III) > 10-4 M, [Ox] > 10-3 M, [pollutant] > 10-4 M]) (Batista et al., 2014; Nogueira and Guimaraes, 2000; Zhan et al., 2006; Zhou et al., 2004a), which are far beyond of concentrations available under natural conditions. Finally, the Fe(III)-oxalate systems function effectively only with a high ratio of [Ox]: [Fe(III)] (Balmer and Sulzberger, 1999; Graça et al., 2017; Lee et al., 2003; Zhan et al., 2006; Zhou et al., 2004b) needed for the stabilization and regeneration of the most photoactive Fe(III) form ([Fe(Ox)3]3-) during photolysis.
In our previous work (Tyutereva et al., 2020) the influence of iron oxalate complexes on p-ASA photodegradation was investigated. The photosystem showed a good degradation quantum yield of ·OH radical production (f308nm = 0.06) and allows to achieve full degradation of both p-ASA and its aromatic by-products at concentration of pollutant less than 7 µM. It is also demonstrated that during photooxidation significant part (~30%) of inorganic arsenic formed was absorbed by photogenerated Fe(III) complexes.
However, in the application of only the Fe(III) oxalate as photoactive agent the concentrations of oxalate and iron could not be below 0.5 mM and 20 µM, accordingly, to maintain stable degradation rate of p-ASA. To increase the efficiency of this process and to reduce the concentration of reagents, we suggest the use of PS as additional oxidizer.
The activation of PS yields the sulfate radical, a very strong oxidizing agent, which can quickly and more selectively react with aromatic compounds than the hydroxyl radical (Real et al. 2016). PS also reduces the concentration of oxalate and increases its efficiency in photodegradation processes, significantly accelerating the degradation of target compounds. The use of PS allows reducing the amount of "sacrificial" oxalate, while maintaining stabilization and regeneration of the most photoactive form of Fe(III).
Thus, the aim of this work is to study the synergetic effect of the iron oxalate complex and PS on the photodegradation of p-ASA. To shed the light on mechanisms of reactions we directly detected the primary intermediates, determined the rate constants of their reactions with p-ASA and optimized the photolysis conditions to achieve the best degradation of both the target compound and organic byproducts.
2.1. Chemicals
p-ASA (C6H8AsNO3, 98%) was from Aladdin Industrial Corporation (Shanghai, China). The arsenic standards used were arsenious oxide (M&B, Denmark) and sodium arsenate (Fluka, St. Louis, USA). Cetyl-trimethylammonium hydroxide (CTAH) solution (Sigma Aldrich), 2,6-pyridinedicarboxylic acid (Sigma Aldrich) and sodium hydroxide (chemically pure) were used for ion chromatography and for pH adjustment, accordingly. Deionized water used to prepare all experimental solutions was obtained with an Ultra Clear water treatment system (SG Wasser, Germany). The pH was measured by the Anion-4100 (Infraspak-Analit, Russia) ionometer.
2.2. Arsenic determination in solutions
Concentration of total arsenic was determined by atomic-emission spectrometry with inductively coupled plasma (ICP-AES). iCap 6000 Duo ICP-AES instrument (Thermo Scientific, USA) with concentric nebulizer and CID detector was applied. The data acquisition and processing was performed by iTEVA (Thermo Scientific, USA) software. The working parameters of ICP-AES were the following: power supply 1150 W, nebulizer flow rate 0.75 L∙min-1, cooling gas flow 12 L∙min-1. The interference of other elements was eliminated by adding scandium nitrate, Sc(NO3)2 with concentration of 0.5 ppm, as internal standard, to all tested solutions. ICP-AES in combination with hydride generation technique (HG-ICP-AES) was used to measure As(III) at the concentration level < 0.1 ppm according to the guide (Thermo Scientific, USA).
2.3. Arsenic speciation in solutions
The concentration of inorganic arsenic (As(III) and arsenate As(V) ions), p-ASA and oxalate ions were determined by capillary zone electrophoresis (CZE). Capel 3R CZE system (Lumex Company, Russia) equipped with photometric UV detector was used (Polyakova et al., 2020). Upon the registration of the analysts the following retention order was observed: oxalate, arsenate, p-ASA and arsenite. The correctness of peaks assignment was confirmed by using corresponding individual standards. Determination of the total arsenic content to assess the balance of the element was performed by ICP-AES.
2.4. Optical spectroscopy and photolysis setups
Agilent 8453 spectrophotometer (Agilent Technologies, USA) was used to record optical spectra. The stationary photolysis experiments were carried out in quartz cells with a total volume of 10 ml and an optical path length of 5 cm at normal oxygen content in solution. Excimer XeCl lamp (308 nm) manufactured by the Institute of High Current Electronics, SB RAS, Tomsk, Russia (Sosnin et al., 2006) was used as a source of a stationary radiation.
Time-resolved experiments were performed using a conventional laser flash photolysis setup (Pozdnyakov et al., 2006). Quartz cells with a total volume of 3 ml and an optical path length of 1 cm at normal oxygen content in solutions were applied. A tunable LS-2137U laser (Lotis, Belarus) with an excitation wavelength of 355 or 266 nm was used as excitation source with pulse duration of about 6 ns, and pulse energy from 1 to 15 mJ. The optical sensitivity of the setup is up to 5 × 10-4, the spectral range is 270–800 nm, and the temporal resolution is 50 ns. For the numerical calculations of kinetic curves, the differential equations were solved by means of thef fourth-order Runge-Kutta method using RUNGE (developer Yu.V. Ivanov) software (Pozdnyakov et al., 2008).
2.5. Photodegradation studies
The photodegradation of p-ASA was also studied by LC 1200 high performance liquid chromatography (HPLC) system (Agilent Technologies, USA) equipped with a diode array detector from the Center of Collective Use «Mass spectrometric investigations» SB RAS. Separations were performed using an Agilent Zorbax Eclipse RapidResolution XBD-C18 column (4.6 × 100 mm, 80 Å, 1.8 μm) in isocratic elution of mobile phase (5% of acetonitrile and 0.1% v/v of formic acid) with the flow rate 0.5 mL/min. The injection volume was 80 μL. Retention time of p-ASA was 2.75 min. Concentration of total aromatics (TAR) was estimated by integration of all peaks in a chromatogram at 250 nm corresponding to the initial compound and its photoproducts. Analysis of the obtained results was performed using Agilent ChemStation software.
3.1 Laser flash photolysis of potassium persulfate in the presence of p-ASA
To generate a sulfate radical and study its reaction with p-ASA monoanion (pKa values: 2.00, 4.02, and 8.92 (Jaafar, 2001)), the flash (266 nm) photolysis of potassium persulfate at pH 7 was performed. To minimize a contribution of the intrinsic pASA photochemistry (Tyutereva et al., 2019) we used laser pulses of low energy (typically, around 1 mJ/pulse). Figure 1 demonstrates the evolution of the absorption spectra of PS in aqueous solution in the presence of p-ASA at pH 7 and the kinetic curves for three characteristic wavelengths. Immediately after the laser pulse, a wide absorption band was observed with a maximum at 440 nm, corresponding to the spectrum of the SO4•- radical (Ivanov et al., 2000). Then this band narrows and grows in amplitude within first two microseconds with a subsequent decrease of the signal on the time scale of tens of microseconds (Fig. 1).
The secondary intermediate observed at time delays more than 3 μs after the laser pulse was assigned to the p-ASA radical-cation (RNH2+•) formed in the reaction of a sulfate radical with a p-ASA (2). A similar intermediate, with an absorption maximum at 430 nm, was registered earlier in the direct photolysis of p-ASA (Tyutereva et al., 2019).
The decay of RNH2+• radical at the initial times can be described by a first-order equation (3) with an effective rate constant k3:
Since the SO4-• radical (Ivanov et al., 2000) undergoes effective recombination in reaction (1), an analytical solution for the schemes (1-3) cannot be realized. However, kinetic curves of transient absorption decay were calculated numerically by the 4th order Runge-Kutta method using the kinetic schemes (1-3) (see Fig. S1 and additional information in ESI for details). The influence of p-ASA concentration on the sulfate radical decay is the most pronounced at 490 nm (Fig. 2). For calculations we used the absorption coefficient of the sulfate radical at 490 nm, ε(SO42-•) = 1000 M-1cm-1 (Ivanov et al., 2000) and the recombination constant of this radical, 2k2 = 2.9×109 M-1s-1, determined from the second-order fit of the kinetic curve at 490 nm in the absence of p-ASA (Fig. 2, curve 1); these parameters were fixed during the subsequent calculations. The remaining parameters were fitted to achieve the best agreement between the calculated and experimental curves in a wide range of p-ASA concentrations (Fig. 2).
The numerical calculations gave the value of the rate constant for the reaction of SO4-• radical with p-ASA monoanion as k2 = (2.4 ± 0.4)×109 M-1s-1; the observed decay rate constant of RNH2+• radical and its absorption coefficient at 490 nm as k3 = (5.7 ± 1.5)×103 M-1s-1 and e(RNH2+•) = (3.4 ± 0.5)×102 M-1cm-1, respectively. At pH 3 the reaction rate constant of the sulfate radical with the neutral form of p-ASA is k2 = (7.3 ± 0.6) ×109 M-1s-1 that is three times greater than that obtained for the anionic one (Fig. S2). This indicates the strong influence of Coulomb repulsion during the interaction of two negatively charged species at pH 7. However both rate constants are close to diffusion-controlled limit, so one can conclude that SO4-• radical could readily oxidize p-ASA at a wide pH range even at low concentration of the target molecule.
3.2 Stationary photolysis of the Fe(Ox)33- complex with PS in the presence of p-ASA.
In our previous article (Tyutereva et all., 2020) we demonstrated that at neutral pH the Fe(III) oxalate complexes effectively generate •OH radical under UV irradiation, which reacts with p-ASA with high rate constant, (8.6 ± 0.5) × 109 M-1s-1. Subsequent oxidation of organic radical formed leads to complete degradation of both p-ASA and basic aromatic photoproducts with formation of inorganic As(V) mainly, under optimal conditions. It is also worth to note, that p-ASA has own photochemistry under 308 nm irradiation but it is negligible in our experimental conditions due to high absorption and photoactivity of Fe(III) complexes.
The presence of both Fe(III) oxalate and PS in the solution accelerates the photodegradation of p-ASA and gives an opportunity to reduce the working concentration of oxalate ions in the solution (from 0.5 to 0.12 mM) without changes in the high degradation efficiency of not only the target compound but the aromatic photoproducts as well (Fig. 3). This can be explained by the catalytic decomposition of PS by photogenerated Fe(II) ions with the regeneration of the initial Fe(III) oxalate complex and the formation of additional oxidative sulfate radical:
Fe(II) + S2O82- = Fe(III) + SO42- + SO4-• (4)
Addition of PS also allows to oxidize p-ASA at high concentration of the pollutant that cannot be done in the presence of iron oxalate alone. Figure 4 shows the dependence of the yield of different arsenic species on the irradiation time of p-ASA – Fe(III) oxalate system in the presence and absence of PS at initial concentration of p-ASA about 3 ppm (4×10-5 M). Without PS only a partial (about 50%) degradation of p-ASA was observed after 40 min of irradiation (Fig. 4A). The total content of arsenic including p-ASA and inorganic species at the end of irradiation (~1.7 ppm) was significantly less than the total arcenic concentration (~2.8 ppm) before irradiation that indicates an accumulation of some organic arcenic byproducts with questionable toxicity.
In the presence of PS (Fig. 4B) the complete degradation of p-ASA was observed with generation of As(V) mainly, which can be removed by standard water treatment procedures (Vircikova et all., 1996; Lawrence and Higgs, 1999). It worth to note that the concentration of As(V) measured by CZE method shows evident decrease after 20 minutes of irradiation comparatively with the experiment withot PS (Fig. 4B). To explain such puzzling effect the test experiments with model As(V) - Fe(III) oxalate – PS system without and with irradiation were carried out under the same conditions.
Figure 5A demonstrates the difference between actual (by preparation) concentration of As(V) measured using methods of ICP-AES and CZE before and after irradiation. Prior to irradiation both methods give results consistent with each other and actual As(V) concentration (Fig. 5A). In the case of photolysed samples, ICP-AES shows no influence of irradiation on the measured concentrations of As(V), while CZE demonstrates a pronounced decrease of As(V) by 0.7 – 2 ppm, depending on the initial arsenic content. This discrepancy increases with exposure (Fig. 5B) that gives a right to assume that a portion of As(V) transforms into another form or product, which is out of registration by CZE, e.g., conjugate or complex, on account of Fe(III)-oxalate degradation. Analogous loss of As(V) measured by CZE was reported in our previous work (Tyutereva et al., 2020) where oxidation of p-ASA was studied in Fe(III)-oxalate system in the absence of PS. According to (Wang et all., 2020), this unknown product is most likely an As(V) – ferric oxyhydroxides colloid.
Another argument for the assumption that part of arsenic is sorbed on photogenerated iron oxyhydroxides are the results of total As determination using ICP-AES analysis of undisturbed samples few days after photolysis (Fig. S3). The total arsenic concentration is restored by a shaking of samples prior the measurements. Figure 5B also illustrates this effect (curves 1 and 1'). This important observation indicates that both processes including p-ASA oxidation and As(V) sorption occurs during a single photochemical process. The significance of this effect for further application of this photochemical approach for p-ASA removal from natural waters is in the reduction of the number and the cost of water purification steps. We intend to continue research in this prospective direction in the nearest future.
Assuming that p-ASA is completely converted to inorganic forms of arsenic we can estimate the loss of As(V) in CZE measurements after 20 and 40 min of irradiation, accordingly. Result was lower that one predicted by photolysis of the model system with As(V) (³1 ppm, Fig. 5B). However, it should be taken into account that in the model system ROS, generated by excitation of Fe(III) oxalate, react immediately with the complex itself. In the presence of p-ASA these active species firstly react with target pollutant and start the degradation of complex only at the final stage of photolysis. So, we can expect the stabilization of Fe(III)-oxalate system in presence of p-ASA as compared with As(V).
After correction of As(V) concentrations measured by CZE, we can conclude that in the presence of 1 mM PS the complete degradation of p-ASA to inorganic arsenic occurs at concentrations of p-ASA up to 3 ppm. The ratio of As(III)/As(V) formed during p-ASA oxidation depends on the initial concentration of the pollutant. Figure 6 demonstrates the arsenic species distribution in solution after 20 min of irradiation depending on the starting concentration of p-ASA. Almost complete conversion of p-ASA to inorganic arsenic could be seen (Fig. 6). Another important result is the decrease of As(III)/As(V) ratio from 0.21 to negligible value (Fig. 6, inset) with the lowering of p-ASA concentration from 3.3 to 0.44 ppm. Apparently, at low p-ASA concentrations all photogenerated As(III) is completely oxidized by PS. High concentrations of p-ASA lead to higher consumption of PS that, in turn, results in incomplete degradation of photogenerated As(III). At real concentration of the organic arsenicals in contaminated surface water (~ tens ppb, Mangalgiri et all., 2015) all p-ASA is expected to fully converts to As(V) in Fe(III)-oxalate – PS system, which significantly reduces the toxic effect of p-ASA on the environment.
Joint oxidation of p-ASA by •OH and SO4-• radical was studied by the combination of steady-state and time-resolved optical methods. The rate constant for the reaction of SO4-• radical with p-ASA monoanion was calculated as k1 = (2.4 ± 0.4)×109 M-1s-1. At pH 3 the neutral form of p-ASA reacts with SO4-• radical three times faster (k2 = (7.3 ± 0.6) ×109 M-1s-1). This indicates the strong influence of Coulomb repulsion on the interaction of two negatively charged partners at neutral pH. It was shown that addition of PS to Fe(III)-oxalate system has several advantages for p-ASA photooxidation. First, the presence of PS significantly reduces the concentration of oxalate required for complete p-ASA degradation. Second, higher degradation efficiency of the target compound leads to lower exposure doses and total time for photolysis. Third, the predominant formation of As(V) reduces the number and cost of subsequent purification cycles. The proposed approach demonstrates efficiency even for low concentrations of p-ASA. Also the observed effect of partial sorption of As(V) on photogenerated Fe(III) species formed upon p-ASA photooxidation opens a new opportunities of Fe(III)-oxalate – PS system application in water purification processes though further studies of the adsorption prosesses are necessary for their effective use in the real applications.
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Competing interests - "The authors declare that they have no competing interests"
Funding:Russian Science Foundation (Grant RSF-NSFC № 21-43-00004) and the National Natural Science Foundation of China (Grant NSFC-RSF 22061132001)
Authors' contributions:Y.E. Tyutereva: Investigation, P.S. Sherin: Writing - Original Draft, E.V. Polyakova: Investigation, V.P. Grivin: Methodology, Software, V.F. Plyusnin: Funding acquisition, O.V. Shuvaeva: Investigation, Methodology, J. Xu: Conceptualization, Funding acquisition, F. Wu: Conceptualization, I.P. Pozdnyakov: Conceptualization, Investigation, Writing - Original Draft, Writing - Review & Editing
Acknowledgements
The financial support of the Russian Science Foundation (Grant RSF-NSFC № 21-43-00004) and the National Natural Science Foundation of China (Grant NSFC-RSF 22061132001) is gratefully acknowledged.
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