Article pubs.acs.org/est
Cite This: Environ. Sci. Technol. XXXX, XXX, XXX−XXX
Photomineralization of Effluent Organic Phosphorus to Orthophosphate under Simulated Light Illumination Xin Zhang, Jing Li, Wen-Yuan Fan, and Guo-Ping Sheng* CAS Key Laboratory of Urban Pollutant Conversion, Department of Applied Chemistry, University of Science and Technology of China, Hefei 230026, P. R. China
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S Supporting Information *
ABSTRACT: Organic phosphorus (OP), one of the main forms of phosphorus in effluent from biological wastewater treatment plants, may contribute to the bioavailable phosphorus pool as well as water eutrophication. However, little is known about the photomineralization of OP or the possible impacts on the phosphorus cycle in water bodies. Herein, the photomineralization of effluent OP was investigated. An increase in orthophosphate concentration was observed under illumination. The 31P liquid nuclear magnetic resonance spectra demonstrated that the release of orthophosphate resulted from photomineralization of OP. Furthermore, the photoproduced hydroxyl radicals (·OH) were proved to play a dominant role in the OP photomineralization. Nitrate, effluent organic matter (EfOM), and Fe(III) presented in effluent were the main chromophores for ·OH photoproduction, and their contributions to ·OH production and photomineralization of OP followed the order: nitrate > EfOM > Fe(III). Additionally, the carbonate (or bicarbonate) in the effluent and high pH were unfavorable for OP photomineralization. The present study revealed the photomineralization behavior of OP in actual effluent, suggesting that photomineralization of OP might contribute to eutrophication and may play a non-negligible role in phosphorus turnover in water bodies.
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INTRODUCTION
OP can be biologically transformed to IP by heterotrophic bacteria, most notably in bottom sediments. However, some autotrophic bacteria, phytoplankton, and zooplankton are also able to convert OP to IP when orthophosphate is unavailable.6 Some free enzymes in water, such as alkaline phosphatase released by various organisms, are also capable of hydrolyzing OP to bioavailable IP.13 In addition to the biological mineralization of OP, abiotic mineralization, i.e., sunlightinduced photomineralization of OP, is hypothesized to be another important process for phosphorus turnover in water. Sunlight-induced photomineralization of dissolved organic matter releases inorganic carbon (e.g., CO2 and CO) and nitrogen (e.g., NH4+ and NO2−).14−16 However, little information is known about photomineralization or the possible environmental impacts of OP in the WWTP effluent. Previous studies have reported photorelease of IP from suspended sediment.17 Photomineralization of several organophosphorus pesticides in deionized water (e.g., glyphosate, methyl parathion, and diazinon) has also been reported.18−20 However, in actual bodies of water containing complex OP and
Phosphorus is a critical limiting nutrient for primary productivity in bodies of water.1 Extensive amounts of phosphorus usually cause eutrophication of aquatic environments and pose a threat to water quality.2,3 Controlling phosphorus is considered more important than controlling nitrogen to reduce eutrophication in many lakes and in some estuaries.4 Phosphorus discharged from wastewater treatment plants (WWTPs) is a major phosphorus source for various bodies of water,5 where the phosphorus exists in both inorganic and organic forms. Inorganic phosphorus (IP) is usually orthophosphate, with a small amount of pyrophosphate and/or polyphosphate, while organic phosphorus (OP) mainly includes phosphorus esters (e.g., nucleic acids) and phosphonates (e.g., glyphosate).6 The concentration of discharged OP in wastewater effluents is typically higher than those in natural waters.7−9 The organic forms of phosphorus in effluent from WWTPs using advanced nutrient-removal technologies can account for 26−81% of total phosphorus.9 Although OP is less bioavailable than IP, OP becomes bioavailable after microbial decomposition, enzyme hydrolysis, or remineralization of sedimentary phosphorus.10−12 Consequently, the contribution of effluent OP to the bioavailable phosphorus pool as well as eutrophication cannot be neglected. © XXXX American Chemical Society
Received: January 16, 2019 Revised: April 6, 2019 Accepted: April 9, 2019
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DOI: 10.1021/acs.est.9b00348 Environ. Sci. Technol. XXXX, XXX, XXX−XXX
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Environmental Science & Technology
experiments. Irradiated light (290−400 nm) intensity on the quartz tubes surface was measured as 40 W/m2 using a UV light photometer (UV-340A, Lutron, Beijing, China). During the experiments, 12 quartz tubes containing 50 mL effluent samples rotated at a constant speed around the lamp to ensure uniform illumination. The quartz tubes for the dark controls were covered with three layers of aluminum foil to shield the solution from light. The solution in each tube was stirred at 400 rpm to maintain a uniform illumination. The reaction temperature was maintained at 25 ± 1 °C with circulating water. At a regular time interval, 1 mL of solution was sampled from each tube for the phosphorus test. To investigate the participation of ROS in OP photomineralization, the effluent samples were spiked with 0.1% (v/v) isopropanol (IPA) to quench the hydroxyl radicals (·OH) or with 5 mM furfuryl alcohol (FFA) to scavenge singlet oxygen (1O2).23 The effluent was adjusted to different pH levels with 0.1 M HCl or NaOH to explore the effect of pH on photomineralization. In addition, different levels of nitrate, humic acids, Fe(III), and HCO3− were separately added to the effluent to investigate their effects on OP photomineralization, and the initial pH was adjusted to 6.5. Humic acids (Sigma, St. Louis, MO, U.S.A.) were purified as described previously before use.24 In addition, DNA, phospholipids, phytic acid, triphenyl phosphate (TPP), and glyphosate were selected as model OPs to explore the photomineralizing capabilities of OP with different structures. The model OP photomineralization experiments were performed in deionized water and the WWTP effluent. Other experimental conditions and phosphorus measurements were the same as those for the effluent OP. Phosphorus Extraction and 31P Liquid Nuclear Magnetic Resonance Analysis. The NaOH/EDTA method was used to extract phosphorus from effluent A.25 In detail, effluent A was lyophilized before and after illumination. Then, the freeze-dried samples were extracted with 0.25 M NaOH and 0.05 M EDTA in D2O solution at 200 rpm and 25 °C. After a 16 h extraction, the extract was centrifuged at 10 000 × g for 5 min, and the supernatant was removed for a phosphorus speciation analysis using liquid 31P NMR. The liquid 31P nuclear magnetic resonance (NMR) spectrum was obtained using a Bruker Avance spectrometer (Bruker Co., Mannheim, Germany) at 162 MHz and 25 °C. Each sample was acquired with 128 scans, using a 90° pulse width and a 7 s relaxation delay. All chemical shifts were calibrated using 85% H3PO4 as an external standard. Detection of Hydroxyl Radicals. Terephthalic acid (TA) was employed as a selective probe to trap ·OH and produce fluorescent 2-hydroxyterephthalic acid (2-HTA). The concentration of 2-HTA, with an excitation wavelength of 315 nm and an emission wavelength of 425 nm, was detected using a fluorescence spectrophotometer (LS55, PerkinElmer Co., U.S.A.). The steady-state concentration of ·OH ([·OH]ss) was calculated based on the formation rate of 2-HTA using the following equation,26
various concomitant compounds, such as wastewater effluent or effluent-impacted waters, photomineralization of OP and the mechanisms involved remain unknown. Considering the possible impacts of OP photomineralization and the complexity of WWTP effluent, it is necessary to investigate photomineralization of OP under actual conditions and to clarify the related mechanisms. In the present study, the photomineralization of OP was investigated in WWTP effluent containing complex OP and chromophores. Several common model OP compounds (e.g., DNA, triphenyl phosphate, and glyphosate) from natural and anthropogenic sources were selected to investigate the photomineralization capabilities of OP with different structures. To explore the photomineralization mechanisms, quenching experiments were performed to investigate the reactive oxygen species (ROS) involved during photomineralization. The key chromophores for ·OH photoproduction were investigated to reveal the contributions of different chromophores to total photomineralization. In addition, the effects of pH and the coexistence of nitrate, humic acids, Fe(III), and (bi)carbonate on ·OH production were separately investigated and were linked to their effects on OP photomineralization to clarify the mechanisms of the effects of these environmental factors. The results of this study will lead to a better understanding about the mineralization of OP in actual waters and its possible influence on eutrophication as well as the phosphorus cycle in the biosphere.
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MATERIALS AND METHODS Sample Collection and Characterization. Wastewater effluent samples were collected from a WWTP in Hefei, China. The effluent sample was passed through 30−50 μm filter paper to remove large suspended particles, which was denoted as effluent A. The OP in effluent A was defined as total OP (TOP). Then, the filtrate was filtered through a 0.45-μm cellulose acetate membrane to remove insoluble components, which was denoted effluent B. The OP in effluent B was dissolved OP (DOP) (Figure S1). The filtered effluent samples were stored in the dark at 4 °C prior to the experiments. The contents of total inorganic phosphorus (TIP) in effluent A and dissolved inorganic phosphorus (DIP) in effluent B were measured by the molybdate blue method.21 After digestion with 5% potassium persulfate, the contents of total phosphorus (TP) in effluent A and total dissolved phosphorus (TDP) in effluent B were determined.21 The contents of TOP and DOP were calculated by subtracting TIP and DIP from TP and TDP, respectively. The content of particulate OP (POP) was determined by subtracting DOP from TOP. A water-quality autoanalyzer (Aquakem 200, ThermoFisher Co., Finland) was utilized to measure nitrate content in the effluent samples. Total organic carbon (TOC) content was measured with a TOC analyzer (Multi N/C 2100, Analytik, Jena, Germany). The protein and humic substance contents in the effluents were determined by a modified Lowry method, and carbohydrate content were measured by the anthrone method.22 Total Fe content was measured using inductively coupled plasma mass spectrometry (X Series 2, ThermoFisher Co., U.S.A.). The water qualities of the effluent A and B are listed in Table S1. Photochemical Experiments. A merry-go-round photochemical reactor (XPA-7, Nanjing Xujiang Motor Factory, Nanjing, China), equipped with a 500 W mercury lamp and a 290 nm cutoff filter, was employed for all photochemical
d[2‐HTA] = 0.35 × k ·OH·TA × [TA][·OH]ss dt
where k·OH·TA is the rate constant, which is 3.3 × 109 M−1 s−1, and 0.35 is the reaction yield for 2-HTA.26 To investigate the role of iron present in the effluent in ·OH photoproduction, chelating resin (Chelex 100 sodium form, Sigma) was used to remove the iron in the effluent. Anionexchange resin (Dowex 1 × 2 chloride form, Sigma) was used B
DOI: 10.1021/acs.est.9b00348 Environ. Sci. Technol. XXXX, XXX, XXX−XXX
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Environmental Science & Technology to remove nitrate to investigate the role of nitrate in ·OH photoproduction.9 Catalase (40 U/mL) was added to remove hydrogen peroxide to investigate the contribution of Fe-based photo-Fenton on the production of ·OH.
ing that the effluent used in this study contained almost no polyphosphate or pyrophosphate, considering the fast hydrolysis of polyphosphate or pyrophosphate with boiled hydrochloric acid.27 Therefore, the concentration of OP in the effluent was calculated by subtracting the concentration of orthophosphate from the TP or TDP. The effluent contained 0.15 mg/L TOP and 0.1 mg/L DOP, which accounted for 19.7% of TP and 15.4% of TDP in effluents A and B, respectively. The POP in effluent A was calculated to be 0.05 mg/L. Figure 1b shows the increases in orthophosphate concentration under light illumination, while it remained almost constant under dark conditions, suggesting that OP can be photomineralized to orthophosphate. As shown in the 31P NMR spectrum, the peak intensity of orthophosphate increased, accompanied by decreased peak intensity of OP after 45 h of illumination (Figure 1a), further demonstrating that the orthophosphate release resulted from photomineralization of OP. The photomineralization of OP followed pseudo-first-order kinetics (Figure 1c), and the photomineralization rate constants for TOP, DOP, and POP were calculated to be 1.02 × 10−2, 6.4 × 10−3, and 2.02 × 10−2 h−1, respectively. The photomineralization rate constant of POP was 3.2 times that of DOP. Photomineralization of Model OP. Considering the complexity of OP in the effluent samples, several typical OPs were selected as model compounds to further explore the photomineralization abilities of OP with different molecular structures. DNA, phospholipid, and phytic acid were selected to represent biological OP, and triphenyl phosphate (TPP) and glyphosate were selected as common anthropogenic OPs. OP mainly consists of compounds with P−O−C (phosphorus ester) and P−C bonds (phosphonate).28 Most OP in wastewater or natural water is phosphorus esters containing P−O−C bonds.29 DNA, phospholipid, phytic acid, and TPP were the phosphorus esters, whereas glyphosate was the phosphonate. Figure 2 shows that the efficiencies of model OP
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RESULTS Photomineralization of Effluent OP. As shown in Table S1, the initial concentrations of TP and TIP in effluent A were 0.76 and 0.61 mg/L, respectively. The TDP and DIP concentrations in effluent B were 0.65 and 0.55 mg/L, respectively. The liquid 31P NMR spectrum of freeze-dried effluent A in Figure 1a shows that the phosphorus was mainly
Figure 1. (a) Liquid 31P nuclear magnetic resonance (NMR) spectra of freeze-dried effluent A before and after 45-h of irradiation; (b) changes in the orthophosphate concentrations in the effluent under irradiated or dark conditions; (c) plots of ln( [OP]t/[OP]0) versus irradiation time (Ortho-P, orthophosphate; OP, organic phosphorus).
Figure 2. Changes in the concentrations of various model OP compounds in (a) deionized water or (b) the effluent under illumination. Initial contents of all model OP were 2.0 mg P/L.
photomineralization followed the order: phytic acid > TPP > DNA > phospholipid > glyphosate, and the photomineralization fractions were 59.3%, 40.0%, 22.0%, 4.9%, and 0.2% in deionized water and 97.3%, 77.5%, 36.0%, 11.8%, and 9.7% in the effluent, after 45 h of illumination. Roles of Reactive Oxygen Species. Photodegradation of organic compounds in wastewater or natural water usually involves direct photolysis and indirect photodegradation mediated by ROS, such as ·OH and 1O2.30 As shown in Figure 3a and b, the photomineralization rates of OP decreased
orthophosphate, with a chemical shift of 5.1−5.3, and OP (∼3.5−4.5) accounted for a small percentage. No peaks of polyphosphate or pyrophosphate were observed in the chemical shift range of −30 to 0 ppm (Figure S2), indicating that all the IP in the effluent existed as orthophosphate. After boiling hydrolysis of the effluent sample with 25 mM hydrochloric acid for 1 h, the orthophosphate concentration was almost unchanged (data not shown), further demonstratC
DOI: 10.1021/acs.est.9b00348 Environ. Sci. Technol. XXXX, XXX, XXX−XXX
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tion rates of OP and photorelease rates of orthophosphate increased with increasing NO3− level (Figure 4a and b; Figure
Figure 3. Changes in the concentrations of effluent OP (a and b) under illumination with the addition of isopropanol or furfuryl alcohol; and (c) steady-state concentrations of photoproduced hydroxyl radicals of the original effluent, effluent with catalase (40 U/mL), chelating resin-treated effluent, or anion-exchange resintreated effluent.
significantly after dosing IPA. Nevertheless, the photomineralization rates were almost unchanged after dosing FFA (Figure 3a and b). Therefore, ·OH dominated photomineralization of OP in WWTP effluent, while 1O2 contributed little to this process. As shown in Figure 3c, the steady-state concentration of · OH ([·OH]ss) was calculated to be 7.2 × 10−16 M for effluent A and 5.3 × 10−16 M for effluent B under illumination. Dosing catalase was used to eliminate ·OH from the photo-Fenton process, and [·OH]ss decreased slightly. After removing the iron present in the effluent with a chelating resin, [·OH]ss decreased 11.3% and 11.9%, respectively, in effluents A and B. [·OH]ss decreased significantly by 66.7% for effluent A and 65.7% for effluent B after removal of nitrate with anionexchange resin. The contribution of effluent organic matter (EfOM) to ·OH production was calculated as the remainder of nitrate and iron contributions. Therefore, the contributions of nitrate, EfOM, and iron to the total ·OH production were estimated to be 66.7%, 22.0% and 11.3% for effluent A, respectively, and 65.7%, 22.4% and 11.9% for effluent B, respectively. Effects of NO3−, Humic Acids, Fe(III), HCO3−, and pH on Photomineralization of Effluent OP. Because of the relatively higher nitrate level in the effluent compared to that in natural water, nitrate (NO3−) is usually considered to be an important ·OH photosensitizer.31 Here, different amounts of NO3− were added to the effluent samples to examine the effect of nitrate on photomineralization of OP. The photomineraliza-
Figure 4. Changes in the concentrations of OP in the effluent dosed with various concentrations of (a, b) nitrate, (c, d) humic acids (HA), (e, f) Fe(III), and (g, h) HCO3− under illumination.
S3a and b). After spiking 1.0, 2.5, and 5 mM NO3−, the OP photomineralization rates increased 0.3-, 1.2-, and 2.2-fold for effluent A, respectively, and 0.3-, 1.5-, and 3.0-fold for effluent B, respectively. After removing nitrate, the photomineralization rate constants of OP decreased 59.3% and 57.3% for effluents A and B, respectively (Figure S4). Humic substances are the main photosensitizer in EfOM and are photoexcited to produce ·OH.32 To examine the effect of humic substances on OP photomineralization, the effluent was illuminated by dosing different concentrations of humic acids. The results showed that the added humic acids suppressed photomineralization of OP and photorelease of orthophosphate (Figure 4c and d; Figure S3c and d). Fe-based compounds are also a crucial photoactive species in water and produce ·OH under illumination.33 Therefore, the effect of Fe(III) on photomineralization of OP was also explored. After dosing 2, 5, and 10 μM Fe(III), the photomineralization rates of OP and photorelease rates of orthophosphate were almost unchanged for effluents A and B (Figure 4e and f; Figure S3e and f). To confirm the role of iron, the iron present in effluent was removed with a chelating resin, and then the photomineralization rate constants D
DOI: 10.1021/acs.est.9b00348 Environ. Sci. Technol. XXXX, XXX, XXX−XXX
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of ·OH in photomineralization of model OP, such as glyphosate and diazinon.18,20 In addition, ·OH also contributes to the photorelease of orthophosphate from suspended sediment OP.17 In some advanced oxidation processes (e.g., UV or UV/H2O2), OP is effectively oxidized by ·OH to release orthophosphate.35 Consequently, the photoproduced ·OH is crucial for OP photomineralization. Herein, the formation of · OH in effluent under light illumination was demonstrated. The steady-state concentration of photoproduced ·OH in the effluent was higher than that in some sunlit surface waters (4.3 × 10−18−2.0 × 10−16 M),36 possibly due to the higher content of nitrate and higher ·OH quantum yield of EfOM.31 Figure 3c revealed that photoproduced ·OH was mainly derived from photolysis of nitrate, and the contributions of various chromophores to ·OH photoproduction followed the order: NO3− > EfOM > Fe(III). Previous study demonstrated that nitrate photolysis dominates ·OH photoproduction in sunlit waters, when the nitrate/organic matter (OM) ratio is higher than 3.3 × 10−5 mol NO3−/mg C.37 In the present study, the nitrate/OM ratios were 11.5 × 10−5 and 12.6 × 10−5 mol NO3−/mg C for effluents A and B, respectively. Hence, it is reasonable to say that nitrate photolysis dominated ·OH production and effluent OP photomineralization. Moreover, under different nitrate, humic acids, Fe(III), or HCO3− levels or under different pH conditions, the [·OH]ss values in effluent A were always higher than those in effluent B (Figure S6), which might partly explain why OP photomineralization of effluent A was always faster than that of effluent B. In addition, photomineralization of POP was faster than that of DOP (Figure 1c), which might also contribute to the faster photomineralization of OP in effluent A compared to that in effluent B. The photoproduced ·OH showed different reactivities toward different types of OP, and the second-order rate constants for DOP, TOP, and POP with ·OH were 2.7 × 109, 3.0 × 109, and 5.9 × 109 M−1 s−1, respectively (Figure S7). These results indicate that POP in the effluent was more photosusceptible than DOP. Interestingly, the calculated second-order rate constants of ·OH and OP (Figure S7) were comparable to those of ·OH and dissolved organic matter (DOM) (k·OH·DOM = 5.6 × 109 M−1 s−1),38 suggesting that photomineralization of OP, as an inherent part of organic matter, might be related to photolysis of organic matter. Photomineralization of Effluent OP Depending on the Molecular Structure. The photomineralization capabilities of OP were closely related to their molecular structures (Figure 2). The phosphorus esters with aryl moieties (TPP and phytic acid) were more easily photomineralized in deionized water and effluent compared to phosphorus ester with alkyl moieties (DNA and phospholipids). This might be due to that P−O with alkyl moieties was more stable than P−O with aryl moieties.39 In addition, glyphosate with a P−C bond was refractory to photomineralization to orthophosphate. A previous work also reported that phosphonates with P−C bonds were relatively recalcitrant to photodecomposition compared to a phosphorus ester with a weaker P−O−C bond.40 Therefore, phosphorus esters were more readily photomineralized than phosphonates under light illumination, and the phosphorus esters with aryl moieties were more susceptible to be photomineralized than those with alkyl moieties. Environmental Factors Affecting Effluent OP Photomineralization. Photomineralization of effluent OP was dominated by photoproduced ·OH. [·OH]ss was significantly
decreased 16.2% and 15.7%, respectively, for effluents A and B (Figure S5). Furthermore, after adding catalase to exclude the contribution of the Fe-based photo-Fenton process, the photomineralization rate constants of effluent OP decreased 11.7% and 10.1% for effluents A and B, respectively (Figure S5). Carbonate (or bicarbonate) is a common anion in water that scavenges ·OH;34 therefore, the effect of carbonate (or bicarbonate) on effluent OP photomineralization was investigated. Adding bicarbonate resulted in a decrease in photomineralization rates of OP and photorelease rates of orthophosphate (Figure 4g and h; Figure S3g and h). After spiking 2.5, 5.0, and 10.0 mM HCO3−, the photomineralization rates decreased 11.9%, 18.0%, and 32.3%, respectively, for effluent A and 6.7%, 22.5%, and 26.8%, respectively, for effluent B. This result suggests that the carbonate (or bicarbonate) in the effluent inhibited photomineralization of OP. Increasing photomineralization rates of OP and photorelease rates of orthophosphate were observed with increasing acidity of the effluent samples (Figure 5; Figure S3i and j). The
Figure 5. Changes in the concentrations of OP in the effluent under illumination at various pH levels.
photomineralization rate constants of OP increased from 8.3 × 10−3 to 3.1 × 10−2 h−1 for effluent A and from 3.1 × 10−3 to 2.94 × 10−2 for effluent B, with a pH decrease from 9.5 to 5.0 (Figure 5).
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DISCUSSION Indirect Photolysis by ·OH Contributes to Photomineralization of Effluent OP. Low-wavelength UV light (e.g., UV254) is effective for OP photomineralization.35 However, the contributions to OP direct photolysis by simulated light were less than 23.5% and 20.6% for effluents A and B, respectively (Figure 3a and b). Quenching experiments indicated that the photoproduced ·OH dominated photomineralization (Figure 3a and b). The faster photomineralization rates of the model OPs in effluent containing various ·OH chromophores than those in deionized water supported the crucial role of ·OH in the photomineralization of OP (Figure 2). Previous studies also reported a critical role E
DOI: 10.1021/acs.est.9b00348 Environ. Sci. Technol. XXXX, XXX, XXX−XXX
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tion rates of OP in the effluents decreased slightly after removing iron or dosing catalase to eliminate the contribution of Fe-based photo-Fenton (Figure S5). These results suggest that Fe(III) present in the effluent may have contributed to OP photomineralization, but its contribution was less than that of nitrate (Figure 3c), considering the low contents of iron in effluents A and B (Table S1). Carbonate (or bicarbonate) in water can quench ·OH and be transformed to the carbonate radical (CO3•−) (eq 4):34
influenced by the levels of concomitant nitrate, humic acids, iron, and HCO3− as well as pH (Figure S6). Therefore, these environmental factors influenced the photomineralization of OP by affecting the level of photoproduced ·OH in the effluents. Nitrate in water can be effectively excited by sunlight reaching the earth’s surface (>290 nm) to produce ·OH (eq 1) and reactive nitrogen:41,42 hv
NO−3 → NO2 + ·OH
(1)
hv NO−3 →
(2)
NO−2
3
+ O( P)
HCO−3 + ·OH → CO•− 3 + H 2O
The decrease of [·OH]ss and concomitant decrease in the photomineralization rate after dosing HCO3− (Figure S6d; Figure 4g and h) indicated that HCO3− suppressed photomineralization by quenching ·OH. Although CO3•− has a longer lifetime and was produced during this quenching process, CO3•− only selectively oxidizes electron-rich compounds, such as nitrogen- or sulfur-containing compounds.47 Therefore, HCO3− exerted an inhibitory effect on OP photomineralization. A low pH seemed to be favorable for photomineralization of effluent OP (Figure 5). The quantum yield of ·OH from nitrate photolysis improved as pH decreased, probably because it produced peroxynitrous acid (HOONO) (eq 3), which decomposes to ·OH, but the deprotonated form ONOO− cannot.48 The low pH also favored photoproduction of ·OH from humic acids.49 − These results suggest that the pH of receiving waters could affect photomineralization rates of OP by shifting the [·OH]ss levels indirectly. Environmental Implications. Orthophosphate production from photomineralization of OP in wastewater effluent was demonstrated in the present study. In addition to inorganic nitrogen, inorganic phosphorus is another crucial factor that causes water eutrophication. Its control is considered to be more important than inorganic nitrogen to weaken eutrophication in some lakes and estuaries.4 Hence, photomineralization of OP caused by sunlight may facilitate eutrophication in actual waters, which should be considered during water pollution control. Currently, some iron- or aluminum-based coagulants have been adopted to further reduce the phosphorus level in effluent to meet the phosphorus-discharge standard. However, these common coagulants fail to effectively remove OP, which may result in considerable residual OP in WWTP effluent and may limit the further decrease of effluent phosphorus levels. Considering the release of orthophosphate due to OP photomineralization, the residual OP needs to be further removed to reduce its impact on some environmentally sensitive areas. Tertiary treatment should be applied to reduce the low level of OP. Given the important role of ·OH in OP mineralization, some advanced oxidation processes involving · OH (e.g., UV or UV/H2O2) may be favorable for OP mineralization; then, the orthophosphate produced can be further removed by coagulating sedimentation. IP is usually assimilated to OP by phytoplankton in natural waters. The OP can be released into the water by cell secretion or lysis. Subsequently, OP is biologically remineralized and utilized by organisms. OP is the main form of dissolved phosphorus in large surface water regions.6 Therefore, mineralization of OP becomes the critical factor that controls the productivity of organisms in these waters. The present study suggested that, with the exception of microbiological mineralization, photochemically mediated OP mineralization
hv
NO−3 → ONOO−
(4)
(3)
Increasing the nitrate level resulted in a higher concentration of ·OH (Figure S6a), thereby rendering faster photomineralization of effluent OP (Figure 4a and b). After removing nitrate, the significant decrease of [·OH]ss (Figure 3c) and the resulting great loss of photomineralization rate (Figure S4) demonstrated that nitrate was the dominant factor for OP photomineralization in the effluents. The concentrations of nitrate in both effluents decreased slightly with illumination. However, the [·OH]ss did not change significantly with increasing illumination time (Figure S8). These findings suggest that the loss of nitrate caused by photolysis was negligible and did not significantly change [·OH]ss. Humic acids, as the predominant photosensitizer in EfOM, show dual roles in ·OH photoproduction in waters containing nitrate.43 On the one hand, humic acids serve as photosensitizers to produce ·OH, but on the other hand humic acids also compete with nitrate for photons (screening effect) to decrease the ·OH production from nitrate photolysis.43 To exclude the screening effect and investigate other possible mechanisms, [·OH]ss was corrected by dividing the screening factors (Figure S9) after dosing with humic acids according to a previous study.44 After correcting by the screening factors, the [·OH]ss spiked with humic acids was still lower than that of the control without dosed humic acids (Figure S6b), suggesting that humic acids also reduced [·OH]ss by scavenging ·OH.38 In addition to inhibiting ·OH production, humic acids may also compete with OP for photons to suppress their direct photolysis.20 These results indicate an inhibiting effect of humic acids on ·OH production, which led to inhibition of OP photomineralization in the WWTP effluent.18−20 Iron present in water is considered to be another important photoactive species for photodegradation or phototransformation of compounds.33 The photomineralizations of some model OPs in deionized water were reported to be significantly promoted after spiking with Fe(III).18,20 However, dosing Fe(III) to the effluents failed to promote photomineralization of OP. Fe(III) dosed to water containing organic matter forms amorphous ferric oxide and organic-Fe(III) complexes.45 Furthermore, the dosed Fe(III) in the effluent was also likely to be complexed by excess orthophosphate, which would decrease the free Fe(III) concentration and suppress the photochemical activity of Fe(III).46 Therefore, adding Fe(III) to the original effluents did not enhance photomineralization of OP (Figure 4e and f). When phosphate was removed from the effluents, the OP photomineralization rates increased after Fe(III) dosing (Figure S10), indicating that Fe(III) promotes the photomineralization. Furthermore, the photomineralizaF
DOI: 10.1021/acs.est.9b00348 Environ. Sci. Technol. XXXX, XXX, XXX−XXX
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Environmental Science & Technology
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also plays an important role regenerating phosphate. Therefore, the role of OP photomineralization in the phosphorus cycle and in biological productivity cannot be neglected in sunlit surface water. In addition, the relationship between microbial and photochemical mineralization of OP deserves investigation to better understand the phosphorus cycle in aquatic environments.
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ASSOCIATED CONTENT
S Supporting Information *
The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.9b00348.
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Protocol for phosphorus analysis, NMR spectra, changes in concentrations of orthophosphate, photomineralization rate constants, steady-state concentrations of photoproduced ·OH, second-order rate constants of · OH, screening factor calculation, and water qualities of effluents (PDF)
AUTHOR INFORMATION
Corresponding Author
*Fax: +86-551-63601592; e-mail:
[email protected]. ORCID
Guo-Ping Sheng: 0000-0003-4579-1654 Notes
The authors declare no competing financial interest.
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ACKNOWLEDGMENTS The authors wish to thank the National Key Research and Development Program of China (2016YFC0401101), the National Natural Science Foundation of China (51738012, 51825804, and 51821006), and the Key Research Program of Frontier Sciences, CAS (QYZDB-SSW-DQC020), for the partial support of this study.
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DOI: 10.1021/acs.est.9b00348 Environ. Sci. Technol. XXXX, XXX, XXX−XXX
