Chlorine Advanced Oxidation Process to Oil

Article pubs.acs.org/est

Application of a Solar UV/Chlorine Advanced Oxidation Process to Oil Sands Process-Affected Water Remediation Zengquan Shu,† Chao Li,†,‡ Miodrag Belosevic,‡ James R. Bolton,† and Mohamed Gamal El-Din*,† †

Department of Civil and Environmental Engineering, University of Alberta, 9105 116th Street, Edmonton, Alberta, Canada T6G 2W2 ‡ Department of Biological Sciences, University of Alberta, 11455 Saskatchewan Drive, Edmonton, Alberta, Canada T6G 2W2 S Supporting Information *

ABSTRACT: The solar UV/chlorine process has emerged as a novel advanced oxidation process for industrial and municipal wastewaters. Currently, its practical application to oil sands process-affected water (OSPW) remediation has been studied to treat fresh OSPW retained in large tailings ponds, which can cause significant adverse environmental impacts on ground and surface waters in Northern Alberta, Canada. Degradation of naphthenic acids (NAs) and fluorophore organic compounds in OSPW was investigated. In a laboratory-scale UV/chlorine treatment, the NAs degradation was clearly structure-dependent and hydroxyl radical-based. In terms of the NAs degradation rate, the raw OSPW (pH ∼ 8.3) rates were higher than those at an alkaline condition (pH = 10). Under actual sunlight, direct solar photolysis partially degraded fluorophore organic compounds, as indicated by the qualitative synchronous fluorescence spectra (SFS) of the OSPW, but did not impact NAs degradation. The solar/chlorine process effectively removed NAs (75−84% removal) and fluorophore organic compounds in OSPW in the presence of 200 or 300 mg L−1 OCl−. The acute toxicity of OSPW toward Vibrio fischeri was reduced after the solar/chlorine treatment. However, the OSPW toxicity toward goldfish primary kidney macrophages after solar/chlorine treatment showed no obvious toxicity reduction versus that of untreated OSPW, which warrants further study for process optimization.

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ification.2,5,6 Generally, advanced oxidation processes (AOPs), such as ozone and UV/H2O2, can be employed either as a tertiary treatment for secondary effluent remediation or as a pretreatment step for recalcitrant wastewater to produce readily biodegradable feed wastewater for biological treatment. However, the principal drawbacks of traditional AOPs for full-scale applications are the relatively high capital and operating costs for implementation and the high use of energy needed for each process.7 Given these drawbacks, there is considerable interest in the utilization of novel AOPs that are partially driven by renewable solar energy (i.e., sunlight photolysis) as the primary energy source for wastewater remediation practices including both industrial (e.g., OSPW) and municipal sources.

INTRODUCTION As a valuable energy source for North America, the oil sands industry in northern Alberta, Canada is currently undergoing significant development and expansion.1 During the surface mining operations, a caustic hot water extraction method is used to extract bitumen from the oil sands, which results in large volumes of oil sands process-affected water (OSPW) retained on site in tailings ponds. The OSPW is a complex mixture of suspended solids, salts, and other dissolved organic compounds [e.g., benzene, humic and fulvic acids, phenols, naphthenic acids (NAs), and polycyclic aromatic hydrocarbons (PAHs), among other fluorophore organic compounds].2 Recently, studies have been conducted for the assessment of OSPW remediation using a variety of treatment processes. Slow and partial microbial degradation of the commercial and OSPW NAs has been observed for NAs having higher ring numbers and/or alkyl branching being more resistant to biodegradation.3,4 Ozonation and UV/H2O2 oxidation were found to have a positive impact on the OSPW degradability and detox© 2014 American Chemical Society

Received: Revised: Accepted: Published: 9692

April 9, 2014 July 17, 2014 July 22, 2014 July 22, 2014 dx.doi.org/10.1021/es5017558 | Environ. Sci. Technol. 2014, 48, 9692−9701

Environmental Science & Technology

Article

Figure 1. Representative absorption of raw OSPW and OCl−, and the solar spectral irradiance spectrum, as measured at noon, August 22, 2013.

Table 1. Characteristics of OSPW and the Solar Experiment Information parameters

experiment

pH turbidity (NTU) TS (mg L−1) TDS (mg L−1) UV transmittance@254 nm (%) TOC (mg L−1) COD (mg L−1) BOD (mg L−1) oxidized NAs (NAs + O; mg L−1) NAs (mg L−1) bicarbonate (mg L−1) chloride (mg L−1) nitrate (mg L−1) sulfate (mg L−1) date

a

1 2a 3 4 a

raw OSPW

OC1− dose (mg L−1)

July 8 August 1 August 22 September 4

200 200 200 300

8.3−8.6 277 2083 2050 30 ± 1 47 ± 2 206 ± 8 13.6 10.7 ± 0.1 21.6 ± 0.5 683 ± 32 451 0.45 ± 0.01 134 ± 1 weather condition sunny from 10 a.m. to 1 p.m. mainly sunny with 3−4 h cloudy period sunny from 10 a.m. to 5 p.m. sunny from 10 a.m. to 5 p.m.

Indicates preliminary experiments for process optimization. Results of this optimization are included in the Supporting Information.

indicate that the molar absorption coefficient spectrum of OCl− overlaps the UV region (UVB and UVA from 280 to 400 nm, Figure 1) of the solar spectrum, which suggests that sunlight may be very efficient in photolyzing OCl−. This OCl− photolysis leads to the generation of the photooxidant hydroxyl radicals (·OH) and chlorine radicals (·Cl), which could further react with organic contaminants in wastewaters leading to their degradation. Of primary concern, as a class of constituents in OSPW, are the high concentrations of NAs that are believed to be the principal contributor for acute, subchronic, and chronic toxicity to aquatic organisms20−24 and the mammalian immune system.2,25,26 Recent advancement in high-resolution mass spectrometry has allowed for the identification of various NAs in OSPW, including classical, oxidized, aromatic, and sulfur/ nitrogenated naphthenic acids.27,28 Classical NAs are a complex mixture of acyclic, monocyclic, and polycyclic carboxylic acids that have the general formula CnH2n+ZO2, where n represents the carbon number and Z indicates the number of rings in the molecule.1 Although low in concentration in the OSPW,29

As one of the emerging UV-based AOPs, the UV/chlorine degradation of organic contaminants in wastewater is currently being evaluated.8−11 For solar UV, the sunlight photolysis of chlorine has been observed in waters (e.g., swimming pools, ponds or reservoirs and seawater) and shown to follow firstorder decay kinetics.12,13 Chlorine photolysis leads to production of highly reactive photooxidants, such as hydroxyl radicals (·OH) and chlorine radicals (·Cl).14 The photochemical reactions for chlorine in water include:15,16 HOCl ↔ OCl− + H+

pK a = 7.5 at 25°C

(1)

HOCl + hv (UV photons) → ·OH + ·Cl

(2)

OCl− + hv → ·O− + ·Cl

(3)

·O− + H 2O → ·OH + OH−

(4)

Previously, fundamental studies of the photolysis process, including photodegradation products and the photodegradation quantum yields of free chlorine by UV light at various pH values and wavelengths have been investigated.9,13,17−19 Results 9693

dx.doi.org/10.1021/es5017558 | Environ. Sci. Technol. 2014, 48, 9692−9701


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PAHs, among other fluorophore organic compounds, and their photodegradation byproducts present in the aquatic environment are also known to be toxic to many organisms.30−32 Despite their potential for toxicity, limited research has been undertaken previously to discover the impacts of these OSPW hydrocarbons on aquatic organisms. The principal objective of this work was to investigate the feasibility of a sunlight-induced photochemical process (i.e., the solar UV/chlorine process) as an application for OSPW remediation. The specific objectives were the following: (1) to examine the impact of pH (raw pH ∼ 8.3 and elevated pH 10) on OSPW NAs degradation kinetics using the bench-scale UV/chlorine process; (2) to evaluate the influence of various OCl− concentrations on degradation of OSPW NAs and fluorophore organic compounds using the solar/chlorine process; and (3) to assess the acute toxicity of OSPW before and after treatment toward Vibrio fischeri and goldfish primary kidney macrophages (PKMs) using the Microtox and in vitro bioassays, respectively.

Edmonton, Alberta, Canada (Latitude: 53° 31′ 34.51″, Longitude: −113°31′ 40.18″), during the summer of 2013. Experiments 1 and 2 were used as preliminary tests for process optimization and are not further discussed. However, results for these experiments are included in the Supporting Information (SI). After process optimization, the results of Experiments 3 and 4 are currently presented herein. For each experiment, a sample of approximately 50 mL of OSPW was placed in a 90 mm diameter Pyrex Petri dish under solar irradiance as the sunlight-exposed sample. Dark control samples (OSPW mixed with NaOCl as per each treatment concentration) were covered by aluminum foil under the same environment as that of the sunlight-exposed samples. The solar/chlorine degradation experiments were conducted in a round solid tray (i.e., bathing pool) with internal diameter of 1.2 m. A 45 L sample of sediment-free OSPW (approximately 7 cm in depth in the tray) was used for each experiment. After covering the tray with aluminum foil, NaOCl was added into the OSPW to achieve concentrations of 200 and 300 mg L−1 OCl− for Experiments 3 and 4, respectively. The OSPW was thoroughly mixed for approximately 20 min before solar exposure and recirculated (∼2 L min−1) during the treatment using a peristaltic pump (Masterflex L/S, Illinois, U.S.A.). The absolute solar irradiance (irradiance reading after calibration) was measured by a spectroradiometer (JAZ-A, Ocean Optics, Inc., Florida, U.S.A.) with the software program SpectraSuite. The approximate UV fluence in the solar system was determined using detailed calculations included in SI (Figure S1; Tables S1−S4). The OCl− concentration was determined by dividing the absorbance at 292 nm by the OCl− molar absorption coefficient (365 M−1 cm−1) as described by Feng et al.17 Analytical Equipment and Methods. Absorption spectra measurements of the OCl− solution and the OSPW were performed using a UV visible spectrophotometer (Varian Cary 50 Bio) with 10.0 mm path length quartz cells (Fisher Scientific, CA). For the analysis of NAs, a Waters Acquity UPLC-MS System (Milford, MA) was used for rapid separation and accurate quantification. A detailed description of the detection method is provided in the SI. For monitoring OSPW fluorophore organic compounds using a qualitative approach, a synchronous fluorescence spectra (SFS) of the OSPW were recorded with a Varian Cary Eclipse fluorescence spectrometer (Ontario, Canada) using the Cary Eclipse system. The optimized wavelength offset (Δλ) between the excitation and emission monochromators for synchronous fluorescence was set at 18 nm, and spectra were recorded between 200 to 500 nm excitation wavelengths with excitation and emission slits set at 5 nm and scan speed 10 nm s−1. The photomultiplier tube (PMT) voltage of the detector was set at 780 V. Toxicity Measurements for the OSPW Degradation. Residual OCl− after treatment in all OSPW samples was neutralized by Na 2 S 2O 3 at the molar ratio ([OCl −]/ 2[Na2S2O3]) of 1:2 prior to toxicity tests to prevent OCl− from impacting toxicity results. Acute toxicity tests for untreated and treated OSPW were carried out using a Model 500 Microtox toxicity analyzer (Azur Environmental, Delaware, U.S.A.), using Vibrio fischeri strains. The 81.9% screening test protocol was used for the sample toxicity assessment, and the results were reported as luminescence inhibition (%). A toxicity >50% is considered the threshold for acute toxicity to V. fischeri. For the in vitro assay of OSPW toxicity toward goldfish primary kidney macrophages (PKMs), a detailed protocol can be found

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MATERIALS AND METHODS Chemicals and Reagents. The internal standard (myristic acid-13C) used in UPLC-high-resolution mass spectrometry analysis and nitrobenzene (NB) were obtained from SigmaAldrich, Canada. Fresh sodium hypochlorite solution (NaOCl, available chlorine 10−15%, Sigma-Aldrich, Canada) was used to prepare the OCl− solutions. Solutions (50%, w/w) of sodium hydroxide (NaOH) and hydrochloric acid (HCl) were used to adjust the pH. Sodium thiosulfate (Na2S2O3, Fisher Scientific, Japan) was used at the end of experiments to quench the residual OCl−. The OSPW was provided by one of the oil sands companies in Fort McMurray, Alberta and was characterized according to the standard methods33 prior to use (Table 1). Laboratory-Scale UV/Chlorine Experiments. The raw OSPW was centrifuged (5 min at 10 000 rpm) to remove suspended particles allowing for increased UV transmittance, whereas NAs removal via centrifugation was found to be negligible (data not shown). Two treatments including raw OSPW at natural pH ∼ 8.3 and adjusted pH = 10 were used to examine the OSPW degradation. A 200 mg L−1 OCl− was used because of the presence of a high concentration of hydroxyl radical-scavenging species, such as carbonate and bicarbonate species. Samples were taken periodically for analysis, and each treatment had two replicates. A quasi-collimated beam UV apparatus (model PSI-I-120, Calgon Carbon Corporation, U.S.A.) equipped with a 1 kW medium pressure (MP) Hg lamp (Calgon Carbon, Pittsburgh, PA, U.S.A.) was used to generate polychromatic UV light. To produce a narrow wavelength band centered at 303 nm (±8 nm), an interference filter (Andover Corporation, New Hampshire, U.S.A.) was installed in the collimated tube. The irradiance was measured by a calibrated UV detector (International Light, model SED 240) connected to a radiometer (International Light, Model IL 1400A). Samples were held in a 5.4 cm diameter beaker (60 mL), and the water path length was 2.62 cm. The distance from the lamp to the top of the water surface was 26 cm. An average UV irradiance (Eavg) of 0.50 mW cm−2 was maintained through multiplying incident irradiance meter reading by a correction factor 3.6 derived from S254 nm/ S303 nm (where S represents sensor sensitivity) throughout the experiments. Solar/Chlorine Experiments. Four experiments were performed (Table 1) on the campus of University of Alberta, 9694



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Figure 2. Overall degradation of OSPW NAs using the UV/chlorine process at pH 8.3 and 10 in lab-scale experiments (a). Average irradiance at 303 nm = 0.50 mW cm−2; OSPW NAs degradation under sunlight in the presence of 200 and 300 mg L−1 OCl−, respectively (b). The solar irradiance over treatment time is provided in Table S1, and detailed calculation method of fluence in the solar system is described in the SI (Tables S2−S4).

CO32− with its higher second order ·OH rate constant (3.9 × 108 M−1 s−1) compared to that of HCO3− (8.5 × 106 M−1 s−1).34 This increased scavenging effect will reduce the availability of photooxidants to degrade NAs at higher pHs, thus corroborating the degradation results of the current study. Given these lab-scale results, no pH adjustment was considered necessary for the solar/chlorine treatment experiments discussed below. Interestingly, 29 ± 2% and 10 ± 3% of the OSPW NAs were oxidized at pH ∼ 8.3 and pH = 10, respectively, for control treatments with the addition of OCl− without UV treatment. Potentially, these results may be attributed to the HOCl/OCl− and HCO3−/CO32− equilibria that play a crucial role determining the degradation (and the rate constants) as the pH shifts from ∼8.3 to 10. Active forms of free chlorine, such as the strongly pH-dependent species HOCl/OCl− (pKa = 7.5) can react with most organic compounds.35 According to Deborde and von Gunten,36 hypochlorous acid is the dominant reactive chlorine species for the reaction with the majority of aliphatic and aromatic organic compounds because of its higher oxidation potential (1.49 V) and its Cl−O bond polarization characterization. Therefore, the fact that higher NAs removal in the dark occurred at pH ∼ 8.3 prior to UV treatment can be explained by the presence of HOCl based on the pHdependent distribution of the principal aqueous chlorine species. A pseudo first-order degradation of OSPW NAs is indicated by a linear plot of ln(C0/C) versus the UV dose (F = Eavgt, mJ cm−2), where C0 and C are the initial and final NAs concentration in OSPW (Figure 2a), Eavg is the average fluence rate in the solution, and t is the exposure time (s). The fluencebased pseudo-first-order decay rate constants (k′, cm2 mJ−1) can be determined from the slope of each plot, and the rate equation can be expressed as

in the SI. Briefly, 7 day PKMs were harvested from goldfish, enumerated, and used for a nitric oxide assay involving production of reactive nitrogen intermediates (RNI). Toxicity is determined by the amount of nitrite in supernatants as calculated by a nitrite standard curve and compared to control treatments (lower RNI indicates higher toxicity).

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RESULTS AND DISCUSSION

OSPW NAs Degradation Using the UV/Chlorine Process. Overall, the NAs removal in the OSPW after the UV/chlorine process were 80 ± 2% and 61 ± 4% for pH ∼ 8.3 and pH = 10 treatments, respectively. These total NAs degradations could be broken down to 29 ± 2% during initial chlorination and 51 ± 3% over 150 min of UV irradiance for pH ∼ 8.3 and 10 ± 3% during initial chlorination and 51 ± 3% over 180 min of UV irradiance for pH = 10. Given the role of OCl− in the degradation of organic contaminants, and its relatively strong absorption at the lower end of the solar spectrum, degradation would be expected to be higher at pH = 10 than at pH ∼ 8.3, as indicated in our previous study.10 At pH = 10, OCl− is dominant with a percent >99%, whereas for pH ∼ 8.3, the distribution of HOCl and OCl− is approximately 17% and 83%, respectively. Given the current results, the presence of HOCl appears to be enhancing the OSPW degradation. This can potentially be explained by reported quantum yields of active chlorine photolysis (defined as moles of active chlorine decomposed per einstein of UV photons absorbed) for HOCl (1.0−1.5) being higher than that for OCl− (0.87−1.3) when subjected to UV irradiation.9,17,18 The quantum yields of HOCl or OCl− may be varying because of various lamps used for investigations (e.g., low or medium pressure lamps). In addition, the observed quantum yields of · OH formation by the photolysis (defined as moles of ·OH formed per einstein of UV photons absorbed by active chlorine) of HOCl (1.40) are higher than that of OCl− (0.28).18 Additionally, according to the distribution of carbonate species in relation to solution pH, HCO3− is dominant with [CO32−] < 10 mg L−1 (calculated from pKa1 = 6. 4, pKa2 = 10.3 for carbonate species) at pH ∼ 8.3, and as the pH increases to pH = 10, the proportion of HCO3− and CO32− becomes 0.68/0.32 (with [HCO3−] and [CO32−] approximately 480 and 220 mg L−1, respectively). As a result, a more significant scavenging effect from carbonate species can be expected at pH 10 because of the increased proportion of

−d[C ] = k′[C ] dF

(5)

The overall k′ values (derived from eq 5) for NAs degradation in OSPW were 1.5 × 10−4 cm2 mJ−1 at pH = 10 and 2.9 × 10−4 cm2 mJ−1 at pH ∼ 8.3, respectively. The overall NAs degradation rate constant at the raw pH ∼ 8.3 was found to be approximately two times higher than that at pH = 10, indicating that the oxidation process proceeded more efficiently in the OSPW, without pH adjustment as discussed previously. 9695



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Figure 3. Synchronous fluorescence spectra (SFS) of OSPW before and after solar photolysis (a); SFS of OSPW before and after the solar/chlorine degradation at 200 mg L−1 OCl− (b) and 300 mg L−1 OCl− (c) in Experiments 3 and 4, respectively. Dark control represents OSPW + NaOCl samples covered by aluminum foil under sunlight.

The results indicate that ·OH radicals contributed predominantly to OSPW NAs degradation, whereas the roles of ·Cl, chlorine, and other active species are insignificant during the UV/chlorine process (detailed discussion in SI). It is believed that the high concentration of Cl− in OSPW can react with ·Cl rapidly (reaction rate constant k = 6.5 × 109 M−1 s−1) to form the relatively less active species ·Cl2−, whereas the reaction between ·OH and Cl− is negligible at the raw OSPW pH.41 Direct Solar Photolysis of OSPW. The removal of NAs via solar photolysis in the absence of chlorine was negligible (data not shown). The absorption of photons is the first step for any photoreaction to occur, and because of the very low molar absorption coefficient of NAs within the solar wavelength region, direct solar photolysis cannot be effective in their degradation.5 As a rapid and sensitive qualitative approach for analyzing fluorophore organic compounds in complex mixtures, SFS has been used extensively,42−46 and its application for the detection of hydrocarbons having fluorescent signature in OSPW has also been reported.29,47 The removal of fluorophore organic compounds in OSPW via solar photolysis was characterized from the SFS, as shown in Figure 3a. For OSPW, peaks at 280 nm (I), 310 nm (II), and 330 nm (III) may be associated with monoaromatic hydrocarbons and PAHs (grouped as fluorophore organic compounds in general), as suggested in the literature.27,48 Consistent with well-documented phenomena,49,50 the removal of fluorophore organic compounds in OSPW arose because of the absorption of photons in the UVA (315−400 nm) range as indicated by a marginal reduction in the excitation wavelength intensity within the range of 290−410 nm after up to 7 h of solar direct photolysis (Figure 3a).51 This reduction in intensity can be attributed to the change in the electronic state of the molecule from a singlet ground state to one or more excited singlet states, which elicits a chemical reaction that breaks or rearranges the chemical bonds within the molecule.52 Additionally, fluorophore organic compounds with higher molecular weight and increasing alkyl substitution tend to have higher quantum yields and, thus, are more sensitive to photochemical oxidation.50,53 Solar/Chlorine Degradation of the OSPW. The addition of the OCl− in conjunction with sunlight resulted in the overall

The comparison of individual NAs degradation rate constants is presented in Figure S3. As suggested by the overall rates, all the NA species (i.e., various Z and carbon numbers) at pH ∼ 8.3 were more reactive than those at pH = 10 in terms of time-based pseudo-first-order rate constant (kt). The overall percent distribution of NAs based on Z and carbon numbers in OSPW is illustrated in Figure S2. The Z number reflects the number of fused rings in the NAs compounds,1 with the distributions including: 4% [1 ring (Z = −2)], 32% [2 rings (Z = −4)], 29% [3 rings (Z = −6)], 9% [4 rings (Z = −8)], 10% [5 rings (Z = −10)] and 16% [6 rings (Z = −12)]. Generally, the NAs with Z = −4 and −6 (61%) and carbon number ranging from 13−18 (82%) constituted the largest proportion of NAs in OSPW with similar compositions from OSPW and extracts described previously.2,37−39 It is also worth noting that the rate constants for NAs with 2 and 3 cyclic rings were relatively low compared to other NAs at both pH levels, although they were the most abundant species (51% of the overall NAs) in OSPW (Figure S3a and S3b). Based on the same Z number, the pseudo-first-order rate constants generally increased with increasing the number of carbons. A similar trend has been observed by Afzal et al.5 for Milli-Q water containing model NAs and OSPW matrix using the lowpressure UV/H2O2 oxidation process. This enhanced reactivity could be related to the increment of hydrogen atoms and/or alkyl groups as more carbons are introduced, resulting in higher reactivity toward ·OH. Within each carbon group, the pseudofirst-order rate constants also increased as the number of cyclic rings increased. According to Perez-Estrada et al.,1 increasing the number of rings resulted in larger numbers of tertiary carbon atoms having H atoms that are more reactive than those on secondary or primary carbons. In addition, the role of the photooxidants ·OH and ·Cl, along with other reactive species formed during the UV/chlorine degradation of OSPW NAs (pH ∼ 8.3) were verified in the presence of NB (a well-known ·OH scavenger with negligible reaction to ·Cl18,40). Briefly, the presence of NB caused a significant decrease of the NAs first-order degradation rate during the UV/chlorine process. With addition of 500 mg L−1 NB, no NAs removal was achieved after 170 min treatment, under the condition where ·OH radicals were fully scavenged. 9696



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Figure 4. Profiles of OSPW NAs for raw OSPW (a); after 7 h of solar/chlorine treatment in Experiment 3 (b) and Experiment 4 (c); overall % NAs removal over time during the solar/chlorine degradation (d). Note that in (d), the removal of NAs was divided into two phases denoted by the dashed line (--). Phase (1) was a chlorination period from approximately 0 to 20 min when OCl− was added to OSPW for rapid mixing under aluminum cover, whereas phase (2) started at 20 min when solar exposure commenced.

increasing OCl− concentration, which may be attributed to the higher solar photon absorption via increased OCl− concentration that results in a higher hydroxyl radical concentration impacting the NAs degradation rates. After 7 h of degradation, the total NAs concentration decreased from 21.6 to 5.2 mg L−1 (Figure 4b and Table S5) for Experiment 3 at 200 mg L−1 OCl−. For Experiment 4 at 300 mg L−1 OCl−, the total NAs concentration decreased further to 3.4 mg L−1 (Figure 4c and Table S5). Overall, the NAs removal from OSPW after solar/ chlorine treatment was 76% and 84%, including 9% and 15% during mixing (chlorination period from −20 to 0 min) and 67% and 69% over the 7 h degradation period in Experiments 3 and 4, respectively (Figure 4d). In accordance with the results obtained in the UV/chlorine lab-scale treatment section, NAs with higher ring and carbon numbers (i.e., Z from −8 to −12 and n from 18 to 20) were removed more readily confirming the structure-dependent degradation of OSPW NAs (Figure 4). Interpreting the results for both the qualitative OSPW fluorophore organic compounds (Figure 3b,c) and the quantitative OSPW NAs (Figure 4d), it can be seen that the oxidation process slowed down or ceased after 200 and 240 min for Experiments 3 and 4, respectively, indicating the rapid depletion of OCl− (OSPW UV spectra provided in Figure S5) and subsequently the end of the oxidation process. Other characteristics of OSPW determined after solar/chlorine treatment indicated an overall improvement of OSPW quality (Table S5). The COD decreased from 206 to 142 mg L−1 after 7 h of sunlight exposure and the UV transmittance at 254 nm rose from 30% to 45% after treatment at 300 mg L−1 OCl−. However, the chloride concentration increased after each treatment, which signals the conversion of OCl− during the degradation process to Cl− as the principal final product. In

reduction of all fluorophore organic compounds over time as observed by marked decreases in peak intensities for both Experiments 3 (Figure 3b) and 4 (Figure 3c). The dark control and 0 min treatments for both experiments indicated that the chlorination alone could appreciably degrade fluorophore organic compounds without sunlight UV input (Figure 3b,c). Compared to the solar photolysis or chlorination alone, the solar/chlorine process led to almost a complete oxidation of fluorophore organic compounds in the OSPW. After 7 h of the solar/chlorine degradation, the lowest peak intensity occurred in Experiment 4 (Figure 3c) using 300 mg L−1 OCl−, implying a higher OCl− dose has higher oxidization capacity toward fluorophore organic compounds. Given this high degradation, the solar/chlorine process should be considered as a viable approach in degradation of the OSPW fluorophore organic compounds. In contrast to the surrogate measurement of OSPW fluorophore organic compounds using SFS, the NAs degradation was measured directly by monitoring NAs concentrations using UPLC-MS allowing for a more thorough investigation of their degradation patterns. Because OSPW NAs degradation kinetics will vary depending on solar irradiance, which changes daily and seasonally, it is more appropriate to describe their degradation as a function of UV fluence rather than treatment time. Indeed, the time-based NAs decay did not show a linear trend for either Experiment 3 or 4 (Figure S4) because of the nonconstant solar energy input as indicated in Figure S1. After converting to a solar UV fluence, a linear plot of ln(C0/C) versus solar UV fluence was obtained, and thus the slope of the plot represents the fluence-based rate constant (Figure 2b). As indicated by the increased degradation rate of Experiment 4 versus Experiment 3, the first-order rate constant increased with 9697



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Figure 5. Toxic effects of treated OSPW at pH ∼ 8.3 and 10 using the UV/chlorine process (a); OSPW samples before/after the solar/chlorine treatment in Experiments 3 and 4 using 200 and 300 mg L−1 OCl−, respectively (b), and OSPW control samples (c) on Vibrio fischeri using 81.9% screen test Microtox protocol with 15 min incubation. Inhibition levels of raw OSPW in solar/chorine experiments were recorded as 40−42%. Note that in (b), the solar exposure started at 0 min.

Figure 6. Toxic effects of treated OSPW for Experiments 3 and 4 on A. salmonicida: (a) RNI production of control, raw OSPW, and Experiment 3 and Experiment 4 OSPW after solar/chlorine process. (b) PKMs were stimulated, or not (control), for 72 h with heat-killed A. salmonicida. PKMs were exposed in vitro for 18 h to the raw and solar/chlorine treated OSPW samples. RNI production in macrophages was induced by heat-killed A. salmonicida and analyzed 72 h later by the Griess reaction. Results are expressed as nitrite concentration in the medium. Results are mean ± SEM of macrophage cultures established from three fish, processed individually. Asterisks (*) indicate statistically significant differences at P < 0.05.

view of byproduct formation, the increase of Cl− seems beneficial to the process because a high concentration of Cl− in OSPW acts as a natural ·Cl sink, thereby minimizing the generation of the chlorinated byproducts during the degradation process. OSPW Toxicity Study. Byproducts formed during the degradation process could potentially be more toxic to aquatic organisms than their parent compounds. Therefore, Microtox and in vitro bioassays using goldfish microphages were employed to measure the acute toxicity of OSPW samples after the UV/chlorine and solar/chlorine treatment processes. For the lab-scale batch experiments, both pH ∼ 8.3 and pH = 10 treatments were considered nontoxic to V. fischeri after UV/ chlorine degradation with average inhibition rates of 42 ± 3% and 31 ± 3%, respectively (Figure 5a). Overall, the toxicity of OSPW at pH = 10 remained consistent over time and underwent a very slight decrease after 150 min UV degradation. In contrast, the toxicity of OSPW at pH ∼ 8.3 was above the threshold for toxicity at the start of the experiment (t = 0 min) and exhibited a slight decreasing trend over the entire experimental duration. Interestingly, the toxicity level at pH ∼ 8.3 was higher than that at pH 10. This may arise from the fact that more significant degradation at pH ∼ 8.3, created

more oxidation byproducts, leading to a higher inhibition treatment level. Nevertheless, the overall toxicity for both pH levels during the UV/chlorine treatment remained below the 50% inhibition threshold. For Experiments 3 and 4, the solar/chlorine treatment of the OSPW led to increased detoxification to V. fischeri over time (Figure 5b). Specifically, for Experiment 3 (200 mg L−1 OCl−), the inhibition levels of OSPW toward V. fischeri gradually decreased from 37% at 0 min (after 20 min mixing or chlorination) to 19% at 420 min. For Experiment 4 (300 mg L−1 OCl−), the inhibition at 0 min spiked to 67% which could be attributed to the presence of the high OCl− concentration in OSPW. This spike then decreased markedly to a final inhibition level of 23% at 420 min (Figure 5b). The results for V. fischeri toxicity tests of the raw OSPW and sunlight/dark controls for Experiments 3 and 4, respectively, are shown in Figure 5c. For both experiments the toxicity of the control samples (sunlight and dark) was significantly higher than the raw OSPW. Enhanced toxicity to aquatic organisms could be attributed to solar photolysis of fluorophore organic compounds, possibly petroleum PAHs,31 leading to the formation of byproducts, such as alcohols, aldehydes, ketones, and acids, considered to contribute principally to the toxicity increase.30,50,54,55 For the 9698



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dark control samples, both NAs and fluorophore organic compounds in the OSPW could undergo reactions with the HOCl and/or OCl−. As the only significant reaction pathway between an organic compound and chlorine,36 electrophilic substitution could result in various chlorinated NAs and fluorophore organic compounds, which may be associated with the enhanced toxicity of the chlorination byproducts.56−58 The effects of the OSPW treated in Experiments 3 and 4 on the antimicrobial responses of goldfish primary kidney macrophages (PKMs) were examined by measuring the production of reactive nitrogen intermediates (RNI) as shown in Figure 6. The exposure of PKMs to raw OSPW and for Experiment 3 and 4 OSPW samples caused a reduction in RNI production compared to positive control cells with Experiment 3 having a statistically significant reduction (Figure 6a). As a toxicity bioassay control, the PKMs stimulated with heat-killed A. salmonicida showed significantly elevated RNI when compared with that produced by nonstimulated cells (Figure 6b). The significant impact found in Experiment 3 may be the result of byproducts generated through the solar/ chlorine treatment process which introduces new toxicity products within the OSPW. However, the increase in RNI in Experiment 4 at the higher 300 mg L−1 OCl− (Figure 6a) indicates that the higher chlorine concentration may result in further degradation of the organic byproducts, thereby reducing overall OSPW toxicity. Further study at higher OCl − concentrations may be needed to confirm this assumption. In previous studies, significant inhibitory effects on RNI production were observed in mouse bone-marrow-derived macrophages (BMDM) after exposure to OSPW organic fractions,25,59 which suggests that the organics could be the principal contributor of the OSPW toxicity. However, there are inherent deviations in toxicity bioassays because of the speciesspecific differences. These deviations, in addition to the lower NAs concentration in the current study, result in the inability to extrapolate the current results to the potential depression of the BMDM nitric oxide response.

ACKNOWLEDGMENTS The authors acknowledge the financial support provided by the research grants from Trojan Technologies and a NSERC Collaborative Research and Development (CRD) grant (M.G.E.D. and M.B.), Alberta Innovates−Energy and Environment Solutions, Helmholtz-Alberta Initiative (HAI), a NSERC research grant for the research tools and instruments, a NSERC Discovery Grant (to J.R.B. and M.B.), and the NSERC Industrial Research Chair in Oil Sands Tailings Water Treatment (Dr. Gamal El-Din). We thank Dr. Arvinder Singh from the Department of Biological Sciences, University of Alberta for the in vitro analyses. In addition, thanks go to Ms. Maria Demeter and Ms. Nian Sun for their technical support in Dr. Gamal El-Din’s research laboratories in the Department of Civil and Environmental Engineering at the University of Alberta.

ENVIRONMENTAL SIGNIFICANCE The structure-dependent reactivity of NAs in OSPW subjected to the lab-scale UV/chlorine treatment in this study agrees well with previous findings for other ·OH driven AOPs.1,2,5 Significant degradation of fluorophore organic compounds and NAs was observed under sunlight in the presence of OCl−, indicating the potential for this process to be used in OSPW remediation. Chlorination alone could further induce acute toxicity to OSPW, and thus chlorination or mixing period should be significantly reduced or shortened in our future work. In this way, more chorine species are available for photolysis, highlighting the ·OH-based degradation pathway. Overall, the findings indicate that the solar UV/chlorine treatment process is a promising “green” treatment approach leading to the decontamination and detoxification of OSPW. However, the dependence of the chlorine photochemical decay potential is dependent on the light field and flux that is related to daytime, season, and location. For higher latitude sites (e.g., Fort McMurray, Canada), the days are much longer in summer; however, in winter, there may not be sufficient sunlight to drive the degradation process. This study addresses the variable solar input issue by describing the OSPW degradation as a function of UV dose (i.e., solar energy input), thus enabling the normalization of the solar irradiance variation which will allow for the evaluation and comparison of solar/chlorine treatment

(1) Perez-Estrada, L. A.; Han, X. M.; Drzewicz, P.; Gamal El-Din, M.; Fedorak, P. M.; Martin, J. W. Structure-reactivity of naphthenic acids in the ozonation process. Environ. Sci. Technol. 2011, 45 (17), 7431− 7437. (2) Wang, N.; Chelme-Ayala, P.; Perez-Estrada, L.; Garcia-Garcia, E.; Pun, J.; Martin, J. W.; Belosevic, M.; Gamal El-Din, M. Impact of ozonation on naphthenic acids speciation and toxicity of oil sands process-affected water to Vibrio fischeri and mammalian immune system. Environ. Sci. Technol. 2013, 47 (12), 6518−6526. (3) Han, X. M.; Scott, A. C.; Fedorak, P. M.; Bataineh, M.; Martin, J. W. Influence of molecular structure on the biodegradability of naphthenic acids. Environ. Sci. Technol. 2008, 42 (4), 1290−1295. (4) Hwang, G.; Dong, T.; Islam, M. S.; Sheng, Z.; Perez-Estrada, L. A.; Liu, Y.; Gamal El-Din, M. The impacts of ozonation on oil sands process-affected water biodegradability and biofilm formation characteristics in bioreactors. Bioresour. Technol. 2013, 130, 269−277. (5) Afzal, A.; Drzewicz, P.; Perez-Estrada, L. A.; Chen, Y.; Martin, J. W.; Gamal El-Din, M. Effect of molecular structure on the relative reactivity of naphthenic acids in the UV/H2O2 advanced oxidation process. Environ. Sci. Technol. 2012, 46 (19), 10727−10734. (6) Martin, J. W.; Barri, T.; Han, X.; Fedorak, P. M.; Gamal El-Din, M.; Perez, L.; Scott, A. C.; Jiang, J. T. Ozonation of oil sands processaffected water accelerates microbial bioremediation. Environ. Sci. Technol. 2010, 44 (21), 8350−8356. (7) Oller, I.; Malato, S.; Sanchez-Perez, J. A. Combination of advanced oxidation processes and biological treatments for wastewater

processes at locations worldwide. This process can not only provide a novel alternative to addressing the urgent need for oil sands industry tailings pond management but also may seem appealing for municipal wastewaters due to its simplicity and lower cost in comparison to other solar-driven processes.60−62

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ASSOCIATED CONTENT

S Supporting Information *

Detailed methodology for NAs analysis, solar UV dose calculation, and in vitro toxicity and related graphs and tables are provided in the Supporting Information. This material is available free of charge via the Internet at http://pubs.acs.org.

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AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. Fax: +1 780 492 0249. Tel.: +1 780 492 5124. Notes

The authors declare no competing financial interest.

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REFERENCES



Article

decontaminationA review. Sci. Total Environ. 2011, 409 (20), 4141−4166. (8) Sichel, C.; Garcia, C.; Andre, K. Feasibility studies: UV/chlorine advanced oxidation treatment for the removal of emerging contaminants. Water Res. 2011, 45 (19), 6371−6380. (9) Wang, D.; Bolton, J. R.; Hofmann, R. Medium pressure UV combined with chlorine advanced oxidation for trichloroethylene destruction in a model water. Water Res. 2012, 46 (15), 4677−4686. (10) Chan, P. Y.; Gamal El-Din, M.; Bolton, J. R. A solar-driven UV/ Chlorine advanced oxidation process. Water Res. 2012, 46 (17), 5672− 5682. (11) Jin, J.; Gamal El-Din, M.; Bolton, J. R. Assessment of the UV/ chlorine process as an advanced oxidation process. Water Res. 2011, 45 (4), 1890−1896. (12) Nowell, L. H.; Hoigne, J. Photolysis of aqueous chlorine at sunlight and ultraviolet wavelengths 0.1. Degradation rates. Water Res. 1992, 26 (5), 593−598. (13) Cooper, W. J.; Jones, A. C.; Whitehead, R. F.; Zika, R. G. Sunlight-induced photochemical decay of oxidants in natural waters: Implications in ballast water treatment. Environ. Sci. Technol. 2007, 41 (10), 3728−3733. (14) Nowell, L. H.; Hoigne, J. Photolysis of aqueous chlorine at sunlight and ultraviolet wavelengths 0.2. Hydroxyl radical production. Water Res. 1992, 26 (5), 599−605. (15) Oliver, B. G.; Carey, J. H. Photochemical production of chlorinated organics in aqueous solutions containing chlorine. Environ. Sci. Technol. 1977, 11 (9), 893−895. (16) Buxton, G. V.; Subhani, M. S. Radiation-chemistry and photochemistry of oxychlorine ions. 1. Radiolysis of aqueous solutions of hypochlorite and chlorite ions. J. Chem. Soc., Faraday Trans. 1 1972, 68, 947−957. (17) Feng, Y. G.; Smith, D. W.; Bolton, J. R. Photolysis of aqueous free chlorine species (HOCl and OCl−) with 254 nm ultraviolet light. J. Environ. Eng. Sci. 2007, 6 (3), 277−284. (18) Watts, M. J.; Linden, K. G. Chlorine photolysis and subsequent OH radical production during UV treatment of chlorinated water. Water Res. 2007, 41 (13), 2871−2878. (19) Feng, Y. G.; Smith, D. W.; Bolton, J. R. A potential new method for determination of the fluence (UV dose) delivered in UV reactors involving the photodegradation of free chlorine. Water Environ. Res. 2010, 82 (4), 328−334. (20) Clemente, J. S.; Fedorak, P. M. A review of the occurrence, analyses, toxicity, and biodegradation of naphthenic acids. Chemosphere 2005, 60 (5), 585−600. (21) Headley, J. V.; McMartin, D. W. A review of the occurrence and fate of naphthenic acids in aquatic environments. J. Environ. Sci. Health, Part A: Toxic/Hazard. Subst. Environ. Eng. 2004, 39 (8), 1989−2010. (22) Hagen, M. O.; Garcia-Garcia, E.; Oladiran, A.; Karpman, M.; Mitchell, S.; Gamal El-Din, M.; Martin, J. W.; Belosevic, M. The acute and sub-chronic exposures of goldfish to naphthenic acids induce different host defense responses. Aquat. Toxicol. 2012, 109, 143−149. (23) He, Y.; Patterson, S.; Wang, N.; Hecker, M.; Martin, J. W.; Gamal El-Din, M.; Giesy, J. P.; Wiseman, S. B. Toxicity of untreated and ozone-treated oil sands process-affected water (OSPW) to early life stages of the fathead minnow (Pimephales promelas). Water Res. 2012, 46 (19), 6359−6368. (24) He, Y. H.; Wiseman, S. B.; Wang, N.; Perez-Estrada, L. A.; Gamal El-Din, M.; Martin, J. W.; Giesy, J. P. Transcriptional responses of the brain-gonad-liver axis of fathead minnows exposed to untreated and ozone-treated oil sands process-affected water. Environ. Sci. Technol. 2012, 46 (17), 9701−9708. (25) Garcia-Garcia, E.; Ge, J. Q.; Oladiran, A.; Montgomery, B.; Gamal El-Din, M.; Perez-Estrada, L. C.; Stafford, J. L.; Martin, J. W.; Belosevic, M. Ozone treatment ameliorates oil sands process water toxicity to the mammalian immune system. Water Res. 2011, 45 (18), 5849−5857. (26) Garcia-Garcia, E.; Pun, J.; Hodgkinson, J.; Perez-Estrada, L. A.; Gamal El-Din, M.; Smith, D. W.; Martin, J. W.; Belosevic, M. Commercial naphthenic acids and the organic fraction of oil sands

process water induce different effects on pro-inflammatory gene expression and macrophage phagocytosis in mice. J. Appl. Toxicol. 2012, 32 (12), 968−979. (27) Rowland, S. J.; West, C. E.; Jones, D.; Scarlett, A. G.; Frank, R. A.; Hewitt, L. M. Steroidal aromatic “naphthenic acids” in oil sands process-affected water: structural comparisons with environmental estrogens. Environ. Sci. Technol. 2011, 45 (22), 9806−9815. (28) Grewer, D. M.; Young, R. F.; Whittal, R. M.; Fedorak, P. M. Naphthenic acids and other acid-extractables in water samples from Alberta: What is being measured? Sci. Total Environ. 2010, 408 (23), 5997−6010. (29) Kavanagh, R. J.; Burnison, B. K.; Frank, R. A.; Solomon, K. R.; Van Der Kraak, G. Detecting oil sands process-affected waters in the Alberta oil sands region using synchronous fluorescence spectroscopy. Chemosphere 2009, 76 (1), 120−126. (30) Bobinger, S.; Andersson, J. T. Photooxidation products of polycyclic aromatic compounds containing sulfur. Environ. Sci. Technol. 2009, 43 (21), 8119−8125. (31) Duesterloh, S.; Short, J. W.; Barron, M. G. Photoenhanced toxicity of weathered Alaska north slope crude oil to the calanoid copepods Calanus marshallae and Metridia okhotensis. Environ. Sci. Technol. 2002, 36 (18), 3953−3959. (32) Barron, M. G.; Little, E. E.; Calfee, R.; Diamond, S. Quantifying solar spectral irradiance in aquatic habitats for the assessment of photoenhanced toxicity. Environ. Toxicol. Chem. 2000, 19 (4), 920− 925. (33) American Public Health Association. Standard Methods for the Examination of Water and Wastewater, 21st ed.; American Water Works Association and Water Environment Federation: Washington, DC, 2005. (34) Buxton, G. V.; Greenstock, C. L.; Helman, W. P.; Ross, A. B. Critical-review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (·OH/·O−) in aqueoussolution. J. Phys. Chem. Ref. Data. 1988, 17 (2), 513−886. (35) Pinkston, K. E.; Sedlak, D. L. Transformation of aromatic etherand amine-containing pharmaceuticals during chlorine disinfection. Environ. Sci. Technol. 2004, 38 (14), 4019−4025. (36) Deborde, M.; von Gunten, U. Reactions of chlorine with inorganic and organic compounds during water treatmentKinetics and mechanisms: A critical review. Water Res. 2008, 42 (1−2), 13−51. (37) Rogers, V. V.; Liber, K.; MacKinnon, M. D. Isolation and characterization of naphthenic acids from Athabasca oil sands tailings pond water. Chemosphere 2002, 48 (5), 519−527. (38) Holowenko, F. M.; MacKinnon, M. D.; Fedorak, P. M. Characterization of naphthenic acids in oil sands wastewaters by gas chromatography-mass spectrometry. Water Res. 2002, 36 (11), 2843− 2855. (39) Jones, D.; Scarlett, A. G.; West, C. E.; Rowland, S. J. Toxicity of individual naphthenic acids to Vibrio f ischeri. Environ. Sci. Technol. 2011, 45 (22), 9776−9782. (40) Fang, J. Y.; Fu, Y.; Shang, C. The roles of reactive species in micropollutant degradation in the UV/free chlorine system. Environ. Sci. Technol. 2014, 48 (3), 1859−1868. (41) Grebel, J. E.; Pignatello, J. J.; Mitch, W. A. Effect of halide ions and carbonates on organic contaminant degradation by hydroxyl radical-based advanced oxidation processes in saline waters. Environ. Sci. Technol. 2010, 44 (17), 6822−6828. (42) Baker, A. Fluorescence properties of some farm wastes: implications for water quality monitoring. Water Res. 2002, 36 (1), 189−195. (43) Carvalho, S. I. M.; Otero, M.; Duarte, A. C.; Santos, E. B. H. Spectroscopic changes on fulvic acids from a kraft pulp mill effluent caused by sun irradiation. Chemosphere 2008, 73 (11), 1845−1852. (44) Eickhoff, C. V.; Gobas, F.; Law, F. C. P. Screening pyrene metabolites in the hemolymph of dungeness crabs (Cancer Magister) with synchronous fluorescence spectrometry: Method development and application. Environ. Toxicol. Chem. 2003, 22 (1), 59−66. 9700



Article

(45) Lu, F.; Chang, C. H.; Lee, D. J.; He, P. J.; Shao, L. M.; Su, A. Dissolved organic matter with multi-peak fluorophores in landfill leachate. Chemosphere 2009, 74 (4), 575−582. (46) Lehto, K. M.; Vuorimaa, E.; Lemmetyinen, H. Photolysis of polycyclic aromatic hydrocarbons (PAHs) in dilute aqueous solutions detected by fluorescence. J. Photochem. Photobiol. A 2000, 136 (1−2), 53−60. (47) Frank, R. A.; Roy, J. W.; Bickerton, G.; Rowland, S. J.; Headley, J. V.; Scarlett, A. G.; West, C. E.; Peru, K. M.; Parrott, J. L.; Conly, F. M.; Hewitt, L. M. Profiling oil sands mixtures from industrial developments and natural groundwaters for source identification. Environ. Sci. Technol. 2014, 48 (5), 2660−2670. (48) Gordon, D. C.; Keizer, P. D.; Hardstaff, W. R.; Aldous, D. G. Fate of crude-oil spilled on seawater contained in outdoor tanks. Environ. Sci. Technol. 1976, 10 (6), 580−585. (49) Plata, D. L.; Sharpless, C. M.; Reddy, C. M. Photochemical degradation of polycyclic aromatic hydrocarbons in oil films. Environ. Sci. Technol. 2008, 42 (7), 2432−2438. (50) Garrett, R. M.; Pickering, I. J.; Haith, C. E.; Prince, R. C. Photooxidation of crude oils. Environ. Sci. Technol. 1998, 32 (23), 3719−3723. (51) Yu, H. T. Environmental carcinogenic polycyclic aromatic hydrocarbons: photochemistry and phototoxicity. J. Environ. Sci. Health, Part C: Environ. Carcinog. Ecotoxicol. Rev. 2002, 20 (2), 149−183. (52) Bolton, J. R. Ultraviolet Applications Handbook, 3rd ed.; ICC Lifelong Learn Inc.: Edmonton, Alberta, Canada, 2010. (53) Jacobs, L. E.; Weavers, L. K.; Chin, Y. P. Direct and indirect photolysis of polycyclic aromatic hydrocarbons in nitrate-rich surface waters. Environ. Toxicol. Chem. 2008, 27 (8), 1643−1648. (54) Ankley, G. T.; Erickson, R. J.; Phipps, G. L.; Mattson, V. R.; Kosian, P. A.; Sheedy, B. R.; Cox, J. S. Effects of light-Intensity on the phototoxicity of fluoranthene to a benthic macroinvertebrate. Environ. Sci. Technol. 1995, 29 (11), 2828−2833. (55) Payne, J. R.; Phillips, C. R. Photochemistry of petroleum in water. Environ. Sci. Technol. 1985, 19 (7), 569−579. (56) Hu, J. Y.; Aizawa, T.; Ookubo, S. Products of aqueous chlorination of bisphenol A and their estrogenic activity. Environ. Sci. Technol. 2002, 36 (9), 1980−1987. (57) Bedner, M.; Maccrehan, W. A. Transformation of acetaminophen by chlorination produces the toxicants 1,4-benzoquinone and Nacetyl-p-benzoquinone imine. Environ. Sci. Technol. 2006, 40 (2), 516− 522. (58) Oliveira, D. P.; Carneiro, P. A.; Rech, C. M.; Zanoni, M. V. B.; Claxton, L. D.; Umbuzeiro, G. A. Mutagenic compounds generated from the chlorination of disperse azo-dyes and their presence in drinking water. Environ. Sci. Technol. 2006, 40 (21), 6682−6689. (59) Garcia-Garcia, E.; Pun, J.; Perez-Estrada, L. A.; Gamal El-Din, M.; Smith, D. W.; Martin, J. W.; Belosevic, M. Commercial naphthenic acids and the organic fraction of oil sands process water downregulate pro-inflammatory gene expression and macrophage antimicrobial responses. Toxicol. Lett. 2011, 203 (1), 62−73. (60) Klamerth, N.; Malato, S.; Maldonado, M. I.; Agura, A.; Fernandez-Alba, A. R. Application of photo-Fenton as a tertiary treatment of emerging contaminants in municipal wastewater. Environ. Sci. Technol. 2010, 44 (5), 1792−1798. (61) Klamerth, N.; Malato, S.; Aguera, A.; Fernandez-Alba, A.; Mailhot, G. Treatment of municipal wastewater treatment plant effluents with modified photo-Fenton as a tertiary treatment for the degradation of micro pollutants and disinfection. Environ. Sci. Technol. 2012, 46 (5), 2885−2892. (62) Chong, M. N.; Jin, B.; Chow, C. W. K.; Saint, C. Recent developments in photocatalytic water treatment technology: A review. Water Res. 2010, 44 (10), 2997−3027.

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Chlorine Advanced Oxidation Process to Oil

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