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co-precipitation process
Water Research 153 (2019) 21e28
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Highly efficient removal of phosphonates from water by a combined Fe(III)/UV/co-precipitation process Shuhui Sun a, Shu Wang a, Yuxuan Ye a, Bingcai Pan a, b, * a b
State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, China Research Center for Environmental Nanotechnology (ReCENT), Nanjing University, Nanjing 210023, China
a r t i c l e i n f o
a b s t r a c t
Article history: Received 30 September 2018 Received in revised form 7 January 2019 Accepted 11 January 2019 Available online 16 January 2019
Considerable amount of phosphorous is present as organic phosphonates (usually in the form of metal complexes, e.g., Ca(II)-phosphonate) in domestic and industrial effluents, which cannot be effectively removed by traditional processes for phosphate. Herein, we employed a proprietary process, i.e., Fe(III) displacement/UV irradiation/co-precipitation (denoted Fe(III)/UV/NaOH), to enable an efficient removal of Ca(II)-phosphonate complexes from water. The combined process includes three basic steps, i.e., Fe(III) replacement with the complexed Ca(II) to form Fe(III)-phosphonate of high photo-reactivity, UV-mediated degradation of Fe(III)-phosphonate to form phosphate and other intermediates, and the final phosphorous removal via co-precipitation at pH ¼ 6. The operational conditions for the combined process to remove a typical phosphonate Ca(II)-NTMP (nitrilotrismethylenephosphonate) are optimized, where ~60% NTMP is transformed to phosphate with the total phosphorous reduction from 1.81 mg/L to 0.17 mg/L. Under UV irradiation, the cleavage of NTMP is identified at the C-N and C-P bonds to form the intermediate products and phosphate in sequence. Also, the combined process is employed for treatment of two authentic effluents before and after activated sludge treatment, resulting in the phosphorous drop from 4.3 mg/L to 0.23 mg/L and from 0.90 mg/L to 0.14 mg/L respectively, which is much superior to other processes including Fenton/co-precipitation. In general, the combined process exhibits great potential for efficient removal of phosphonates from contaminated waters. © 2019 Elsevier Ltd. All rights reserved.
Keywords: Phosphonates Photolysis Fe(III) replacement Advanced treatment
1. Introduction Phosphonates are anthropogenic chelating agents widely used as scale inhibitors, corrosion inhibitors, and laundry detergents (Rott et al., 2018a), featuring one or more phosphonic acid groups [C-PO(OH)2] (Jaworska et al., 2002; Nowack, 2003). The global phosphonate consumption was 56,000 t/a in 1998 (Davenport et al., 2000) and increased to 94,000 t/a in 2012 (EPA, 2013). Fig. 1 gives an overview of the formula structure of six typical phosphonates, including 2-phosphonobutane-1,2,4-tricarboxylic acid (PBTC), 1hydroxyethane 1,1-diphosphonic acid (HEDP), nitrilotris(methylene phosphonic acid) (NTMP), ethylenediamine tetra(methylene phosphonic acid) (EDTMP), diethylenetriamine penta(methylene phosphonic acid) (DTPMP) and hexamethylenediamine tetra(methylene phosphonic acid) (HDTMP).
* Corresponding author. Research Center for Environmental Nanotechnology (ReCENT), Nanjing University, Nanjing 210023, China. E-mail address: [email protected] (B. Pan). https://doi.org/10.1016/j.watres.2019.01.007 0043-1354/© 2019 Elsevier Ltd. All rights reserved.
Except for inorganic phosphate, phosphonate is one of the key species of elemental phosphorous in industrial and municipal effluents as well as in natural waters (Nowack and Stone, 2006; Gu et al., 2011; Majed et al., 2012; Qin et al., 2015). Though few open data are available on the phosphate concentration in industrial effluents, we sampled several biotreated textile effluents and discriminated the six phosphonates (in Fig. 1) with their total amount reaching 0.1e0.5 mg P/L. In addition, phosphonates have been identified in the influents of German and Swiss WWTPs (Nowack, 1998, 2002). For instance, NTMP of 0.2e1.1 mM was detected in the influent of a WWTP as influenced by textile industry (Nowack, 2002), with an average DTPMP concentration of 0.12 mM in the effluent (Nowack, 1998). It is assumed that in Europe, 9000e18,600 t/a of phosphonate is finally discharged into receiving waters (Rott et al., 2018b). Phosphonates can hardly be biomineralized during activated sludge treatment, either aerobically (Horstman and Grohmann, 1988) or anaerobically (Nowack, 1998). Traditional processes such as adsorption and precipitation/flocculation could enable highly efficient removal of inorganic phosphate. However, they usually work ineffectively for organic phosphonates
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Fig. 1. Molecular structure of several typical phosphonates (based on ACS, 2016).
(Neft et al., 2010; Gu et al., 2011; Rott et al., 2018a). The complexation of phosphonates with flocculants results in an excessive dosage of metal salts and adversely affect the final removal of total phosphorous (TP) (Nowack, 2006; Rott et al., 2017b). Consequently, an increasing fraction of dissolved organic phosphorus (DOP) detected in the effluent of wastewater treatment plants (WWTPs) (Rott et al., 2018a). On the other side, phosphonates usually remain in the adsorbed form in activated sludge and metal precipitates (Nowack, 2002) after traditional wastewater treatment processes (Rott et al., 2017b). During the disposal of the solid residues (for instance, used in agriculture), phosphonates are brought into the environment. Also, considerable amounts of phosphonates may enter the receiving bodies via the discharge of treated effluent. Phosphonates are subject to natural elimination in environment (Nowack and Stone, 2000; Matthijs et al., 1989; Schowanek and Verstraete, 1990), accompanying a long-term release of bioavailable phosphate in water. Moreover, there still exists substantial risk for phosphonate transformation to more toxic aminomethylphosphonic acid (AMPA) particularly in the presence of sunlight (Grandcoin and Baures, 2017). Thus, the contribution of phosphonates to water eutrophication and environmental safety should be particularly concerned (Studnik et al., 2015; Qin et al., 2015; Drzyzga and Lipok, 2017), and development of efficient and cost effective approaches to realize phosphonate removal from contaminated waters is urgently desired. As well known, decarboxylation of Fe(III)-carboxyl complexes could proceed via ligands to metals charge transfer (LMCT) under UV irradiation. Based on the principle, a combined process, i.e., Fe(III) displacement/UV irradiation/precipitation has been recently developed to effectively remove various toxic metal-carboxyl complexes from water, including Cu(II)-citrate (Pan et al., 2014; Xu et al., 2015) and Cu(II)-EDTA (Shan et al., 2018). In brief, most metal-carboxyl complexes are inert to UV irradiation, and the free
Fe(III) ions was deliberately added to the metal-carboxyl complex solution to form Fe(III)-carboxyl complexes of high photo-reactivity because Fe(III) displays much higher complexation constants with carboxyl ligands, where free toxic metal ions are simultaneously released for subsequent removal by chemical precipitation. Later, it was successfully extended to remove Cr(III)-citrate from synthetic solution and authentic tannery effluent despite of its extremely slow decomplexation rate (Ye et al., 2017), where the hydroxyl radicals formed during Fe(III)/UV process tended to oxide the complexed Cr(III) to free Cr(VI), greatly speeding up the decomplexation and driving a fast formation of Fe(III)-citrate for photolysis. Moreover, the toxic Cr(VI) was effectively reduced back to free Cr(III) for co-precipitation in the presence of Fe(II) formed during photolysis. As for phosphonates, limited study has been available on the photo-reactivity of their Fe(III) complexes in the past decades. Matthijs et al. (1989) reported Fe(III)-EDTMP could be photodegraded through a photoinduced ligand-to-metal charge transfer procedure under sunlight irradiation. Lesueur et al. (2005) further confirmed the Fe(III) mediated photolysis of four common aminophosphonates during UV irradiation. Kuhn et al. (2018) evidenced that during photo-degradation of DTPMP in the presence of Fe(II), the initial cleavage of DTPMP was initiated at the C-N bond, resulting in the gradual drop of C-PO(OH)2 groups in the phosphonate. The above study inspired us to consider the potential of photo-assisted process for water decontamination from phosphonates. In this study, we demonstrated the applicability of the Fe(III)/ UV/NaOH process to eliminate various Ca(II)-phosphonate complexes from synthetic solution and authentic effluents. NTMP was employed as the model phosphonate to optimize the operational conditions and elucidate the underlying mechanism. It is a typical aminophosphonate characterized by three C-PO(OH)2 groups and
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usually present as metal complex, e.g., Ca(II)-NTMP complex, in natural waters or biotreated effluents (Nowack, 2003). Afterward, the process was extended to removal of other Ca(II)-phosphonates, and some typical processes including Fenton/precipitation were employed for comparison. Two authentic effluents were sampled to further demonstrate the applicability of the Fe(III)/UV/NaOH process. 2. Materials and methods 2.1. Materials and instruments All the reagents used in the experiments are of analytical reagent grade without further purification. PBTC (50% aqueous solution) was obtained from Matrix Scientific Co., USA. Solid HEDP$H2O (95%), NTMP (97%) and iminodi (methylenephosphonic) acid (IDMP, 97%) were purchased from SigmaAldrich (St. Louis, MO, USA). EDTMP (98%), HDTMP (97%) and DTPMP (50% Aqueous solution) were provided by Aladdin Biochemical Technology Co., Ltd, China. AMPA (99.0%) and 9fluorenylmethylchloroformate FMOC-Cl (98.0%) were purchased from J & K Scientific Co., Ltd., China. HPLC-grade methanol (MeOH) and acetonitrile (ACN) were obtained from Merck Co., Germany. Other chemicals, including FeCl3$6H2O, FeSO4$7H2O, CaCl2, CuCl2, NiSO4, H2O2, NaOH and HCl, were purchased from Sinopharm Chemical Reagent Co., China. The sample of natural organic matter (NOM) was purchased from the International Humic Substances Society (Catalog No. 2R101N, Suwannee River NOM). The stock solution of Ca-NTMP, Ca-PBTC, Ca-HEDP, Ca-EDTMP, Ca-DTPMP and Ca-HDTMP were prepared by mixing CaCl2 with the corresponding ligand acids in the molar ratio of 1:1. All the solutions were prepared with ultrapure water, and pH was adjusted by HCl or NaOH. The authentic water samples (influent and effluent) were collected from a WWTP (located in Jiangyin City, China) based on A/O process, as designed to treat the mixture of printing and dyeing wastewater (~40%) and domestic wastewater (~60%) with the total capacity of 10000 m3 per day. UV irradiation experiments were carried out in a rotating disk photoreactor (Nanjing Stone Tech Electric Equipment Co., China) with a 300-W medium pressure mercury lamp (Shanghai Hongguang Tungsten & Molybdenum Technology Co., Ltd.) as the UV source. The light intensity was about 2.0 mW/cm2 measured at 365 nm by a radiometer (Photoelectric Instrument Factory of Beijing Normal University, China). 2.2. Experimental procedures The combined process, i.e., Fe(III)/UV/NaOH, included two main steps in sequence: (1) FeCl3 of preset amount was added into solutions of Ca-NTMP and other Ca-phosphonates, then the mixture was diluted to 50 mL by using ultra-pure water to keep the initial P at 1.81 mg/L (0.02 mM) for NTMP. After the addition of Fe(III), pH was adjusted to a predetermined value. Then, each solution was subject to UV irradiation. Methanol (100 mM) was added as the scavenger of $OH radicals to examine its possible role during UV irradiation. (2) After UV irradiation, the pH value increased by the maximum of 0.3 unit. NaOH was further added to the solution to achieve a preset pH, then the mixture was kept stirring at 60 rpm for 30 min and stood for 60 min. Afterward, the supernatant was filtered through a 0.22-mm membrane, and the concentration of concerned species in solution, including phosphonates and some of their intermediate products, TP, was determined.
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For comparison, Fenton/NaOH, Fenton/UV/NaOH, and UV/H2O2, were performed following similar procedures described below. Whether the chemicals and UV irradiation were employed depended on the nature of the processes. Generally, they involved two steps in sequence: (1) FeSO4 was added to 50-mL phosphonate solution (0.02 mM) and the initial pH was adjusted to a predetermined value, where H2O2 was further added to initiate the reaction (FeSO4 free for UV/H2O2). After stirring for 1 h and under the UV irradiation (not for Fenton), the formed PO3 4 was determined. (2) NaOH was added to the pretreated solution to achieve a preset pH, then the mixture kept stirring at 60 rpm for 30 min and stood for 60 min. Afterward, the supernatant was filtered through a 0.22-mm membrane, and the TP concentration was determined. Detailed conditions of the above processes are available in the related figures and tables. 2.3. Analytical methods TP and PO3 4 were determined by molybdenum blue spectrophotometric method (Su, 2001) with a UV-vis spectrometer (T6, PGENERAL, China) at the wavelength of 700 nm. For TP determination, sample digestion using peroxodisulfate (K2S2O8) was required prior to analysis. The concentration of Fe(III)-NTMP was measured by HPLC (Waters 1525) equipped with a UV detector (Waters 2489) (Nowack, 1997). A C18 column (150 mm 4.6 mm, 5 mm particle size, Sunfire, Waters) was used for chromatographic separation, with experimental conditions detailed in Table S1 in Supplementary Materials. AMPA and iminodi(methylenephosphonic acid) (IDMP) is determined by HPLC-fluorescence detector (Waters 2745) with pre-column FMOC-Cl derivatization according to the previous method (Wang et al., 2016). 3. Results and discussion 3.1. Validation of the Fe(III)/UV/NaOH process 3.1.1. Effect of Fe(III) dosage The feasibility of the combined process, i.e., Fe(III)/UV/NaOH, on the removal of Ca(II)-NTMP was initially probed. For comparison, another combined process of UV free, i.e., Fe(III) addition followed by precipitation (namely Fe(III)/NaOH), was employed, and the results are depicted in Fig. 2a. In the set of the experiments, the concentration of Fe(III) varied from 0.040 to 0.20 mM, while the initial pH was set as 3.0, the UV irradiation time was 60 min, and the precipitation pH was optimized at 6.0 (Fig. 2d). As observed, Fe(III)/NaOH could only result in limited TP removal (49%) from the Ca(II)-NTMP solution even at the Fe(III) dosage of 0.20 mM. Comparatively, UV/Fe(III)/NaOH could result in nearly 100% TP removal at the Fe(III) dosage of 0.12 mM or higher. Such relatively low removal by the Fe(III)/NaOH process (<50%) is possibly due to the poor precipitation of ferric hydroxide in the presence of organic ligands (Zhou et al., 2008; Henneberry et al., 2011; Hu et al., 2015), which compete with hydroxy anions to bind Fe(III). Obviously, the UV-irradiation unit is indispensable to the combined process for enhanced TP removal. Also, the formation of phosphate from the Ca(II)-NTMP solution was detected after Fe(III) addition and UV irradiation. As seen in Fig. 2a, the production rate of phosphate increased with the increasing Fe(III) addition, which is consistent with the final TP removal. As the Fe(III) dosage increased to 0.12 mM, the phosphate formation reached 61%, and the residual TP after precipitation dropped down to <0.3 mg/L. Further Fe(III) addition to 0.2 mM offered limited promotion to TP removal. In the subsequent
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Fig. 2. (a) Effect of Fe(III) dosage on Fe(III)-mediated processes for removal of Ca-NTMP complexes. Fe(III)/NaOH: Fe(III) addition followed by precipitation; Fe(III)/UV/NaOH: Fe(III)/ UV process followed by co-precipitation; (b) Effect of UV irradiation duration on the formation of PO3 4 and TP removal by the Fe(III)/UV/NaOH process; (c) Effect of initial pH on the formation of PO3 4 by the Fe(III)/UV process; (d) Effect of precipitation pH on the residual TP by the Fe(III)/UV/NaOH process ([Ca-NTMP]0 ¼ 0.020 mM (1.81 mg TP/L), Fe(III) dosage ¼ 0.12 mM for b, c and d, initial pH ¼ 3.0 for a, b, and d, UV irradiation ¼ 1 h for a and d, precipitation pH ¼ 6.0 for a, b, and c).
experiments, we set the optimal Fe(III) dosage as 0.12 mM for the Fe(III)/UV/NaOH process except for otherwise specified. 3.1.2. Effect of UV irradiation time The influences of UV irradiation duration on the formation of PO3 4 as well as the final removal of TP are illustrated in Fig. 2b. The PO3 4 formation reached ~40% after 10-min UV irradiation, indicating the cleavage of the C-P bond of NTMP. Similar to the effect of Fe(III) dosage, increasing the UV irradiation time is favorable for the final TP removal. In the initial 10 min, about 40% PO3 4 was formed and nearly 80% TP was removed after precipitation, indicating that NTMP was partially transformed to low-molecular-weight phosphonates, which was readily removed during Fe(III) coagulation. After 60-min irradiation, the phosphate formation reached 60% with the residual TP below 0.1 mg/L after precipitation. The above results implied that the formation of phosphate could be an indicator for the final TP removal. In the forthcoming experiments, we set 60 min as the optimal UV irradiation time for the Fe(III)/UV/ NaOH process except for otherwise specified. 3.1.3. Effect of initial pH The effect of initial pH (1.5e10.0) on the phosphate formation during UV irradiation was evaluated, and the results are shown in Fig. 2c. As seen, pH ¼ 3.0 is the optimal value for phosphate formation (65% at pH 3.0), and higher pH values resulted in lower efficiency of NTMP photolysis and phosphate formation. At
pH ¼ 2.5e3.0, Fe(III) is mainly present in the form of FeOH2þ, displaying the highest photo-reactivity among all the Fe(III) species to generate $OH radicals based on eq. (1) (Hug et al., 2001). At pH<2 and pH>3, the predominant Fe(III) species turned to be Fe3þ and ferric hydroxide colloids/sediments, respectively, both suppressing the generation of $OH radicals intensively (Bajt et al., 2001) and resulting in a low PO3 4 formation. At alkaline pHs, phosphonates preferably bind Fe(III), and the Fe(III)-phosphonate complexes are attacked by hydroxyl radicals more easily than the counterpart phosphonates (Rott et al., 2017a). Note that phosphate formation still reached 40e50% at neutral or alkaline pHs. Also, the initial pHs were set in the range of 2.5e3.0 for the subsequent Fe(III)/UV/NaOH processes except for otherwise specified. FeOH2þ þ hv / $Fe2þ þ $OH
(1)
3.1.4. Effect of the precipitation pH Effect of the precipitation pH on TP removal is described in Fig. 2d. As seen, the pH values from 4.5 to 6.0 resulted in the lowest residual TP, whereas at pH from 8.0 to 12.0, the residual TP remained at a relatively high level. Such pH-dependent precipitation correlated well with the zeta potential of the resultant flocs (Fig. S1), which exhibited a dramatic drop from 1.3 mV to ~ 40 mV at the pH range from 5.0 to 8.0. In theoretical viewpoint, the higher the pH, the stronger the electrostatic charge repulsion between phosphate/phosphonates and the formed iron hydroxide
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surface. In the following section, the optimal pH was set as 6.0 for precipitation. 3.2. Mechanism of the Fe(III)/UV/NaOH process 3.2.1. Fe(III) displacement The stability constant of Ca(II)-NTMP (logK ¼ 7.6) is much lower than that of Fe(III)-NTMP (logK ¼ 21.2) (Popov et al., 2001), and Fe(III) ions are capable of displacing Ca(II) from Ca(II)-NTMP in thermodynamic viewpoint. The effectiveness of Fe(III) displacement was reflected by the formed Fe(III)-NTMP complexes, as determined by ion-pair high-performance liquid chromatography (Nowack, 1997). Expectedly, the addition of Fe(III) resulted in an equimolar formation of Fe(III)-NTMP complexes (Fig. S2), indicating that Fe(III) could displace equimolar Ca(II) from Ca(II)-NTMP complexes to release the free Ca(II) ions (eq. (2)). Ca(II)-NTMP þ Fe(III) /Fe(III)-NTMP þ Ca(II)
(2)
3.2.2. Role of ·OH radicals As previously reported (Bajt et al., 2001; Hug et al., 2001), photoexcitation of FeOH2þ would result in the formation of $OH radicals (eq. (1)), which was demonstrated by EPR spectra of Fe(III)/ UV system recorded in Fig. 3a. Clearly, the signals of the $OH radicals without UV irradiation were ignorable, while strong signals were detected in the Fe(III)/UV system even in the first 2 min, indicating the generation of massive $OH radicals. As the irradiation proceeded after 20 min, the intensity of $OH signals turned much weaker, possibly due to the consumption of both the radicals and the FeOH2þ species. To further explore the role of $OH radicals in NTMP degradation, we examined the effect of methanol, a typical $OH radical scavenger, on the formation of PO3 4 . As shown in Fig. 3b and 20-min irradiation could result in the efficiency of phosphate transformation higher than 50% for the Fe(III)/UV process. The addition of 100 mM methanol significantly inhibited the phosphate formation, and 60 min was required to achieve similar efficiency. Thus, we believe that $OH radicals play a significant role in the formation of PO3 4 . Note that such adverse effect from the added methanol gradually turned weaker as the irradiation proceeded. This is mainly because another important process, i.e., direct photolysis of Fe(III)-phosphonate, contributed significantly to the formation of phosphate, as discussed below.
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3.2.3. Photolysis of Fe(III) complexes It is well known that ferrous carboxylates could be photodegraded under UV irradiation via LMCT process (eq. (3)), such as Fe(III)-citrate (Xu et al., 2015; Ye et al., 2017), Fe(III)-EDTA (Shan et al., 2018), and Fe(III)-oxalate (Safarzadeh-Amiri et al., 1997; Chen et al., 2007; Doumic et al., 2015). [Fe3þL] þ hv / Fe2þ þ L
(3)
In fact, similar LMCT process occurred for Fe(III)-phosphonate complexes. Generally, phosphonate in natural water or bioeffluents is mainly present in the form of metal complexes (Ca2þ, Mg2þ, etc.), most of which are of poor photo-reactivity and insensitive to UV irradiation. In theory, Fe(III) could displace the complexed metals (such as Ca(II)) to form Fe(III)-phosphonates due to their stability constants several orders in magnitude higher than the former ones, and the formed Fe(III)-phosphonates would be effectively photodegraded under UV irradiation. For instance, no photodegradation of HEDP was observed in distilled water or in the background of Ca(II), however, the addition of Fe(III) or Cu(II) resulted in a rapid photodegradation of HEDP (Nowack, 2003). Matthijs (1989) reported that Fe(III)-EDTMP could be photo-degraded via stepwise processes, i.e., from the parent compound to ethylenediaminetrimethylenephosphonate, ethylenediaminedimethylene phosphonate, and ethylenediaminemonomethylenephosphonate in sequence, the latter of which is stable in the presence of Fe(III) and light irradiation. Kuhn et al. (2017) identified IDMP as the major breakdown product of EDTMP during UV irradiation of non-complex EDTMP. Kuhn et al. (2018) found that in the presence of Fe(III) and UV irradiation, the C-N bond of DTPMP was cleaved, resulting in the formation of IDMP, ethylamino(bismethylenephosphonic acid) (EABMP) and AMPA. Based on the above discussion, the degradation scheme of Ca(II)-NTMP is proposed in Fig. 4. First, Fe(III) replaced Ca(II) from Ca(II)-NTMP to form Fe(III)-NTMP, where the free Ca(II) is released stoichiometrically. Afterward, the UV irradiation activates the photolysis of Fe(III)-phosphonate complexes based on LMCT process, resulting in the generation of Fe(II) and a phosphonyl radical (Kononova and Nesmeyanova, 2002). Then, PO(OH)O$ donates one electron, leading to the formation of the carbon-centered radical (PO(OH)2-(CH2)2-N-$CH2) and the release of phosphate (Barrett and McBride, 2005; Jaisi et al., 2016). The methylene radical is rapidly intercepted by O2 to form a peroxyl radical, which will be
Fig. 3. (a) EPR spectra of the Fe(III)/UV process under different irradiation duration (initial pH ¼ 3.0, [Ca-NTMP]0 ¼ 0.020 mM (1.81 mg TP/L), Fe(III) dosage ¼ 0.120 mM); (b) Effect of methanol on the formation of PO3 4 ([Ca-NTMP]0 ¼ 0.020 mM, initial pH ¼ 3.0, Fe(III) dosage ¼ 0.120 mM).
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Fig. 4. Proposed reaction pathway for transformation of Ca-NTMP for Fe(III)/UV unite of during the Fe(III)/UV/NaOH process.
insignificantly during the whole process, further validating the reliability of the above mechanism. A slight drop in the TP of the four P species during the initial reaction indicated the formation of other intermediate phosphonates, which were transformed to IDMP, AMPA and phosphate eventually.
Fig. 5. Profiles of the degradation products of Ca-NTMP during Fe(III)/UV process. [TP*] ¼ [PO3 4 ]þ[AMPA]þ[IDMP]þ[NTMP] (Fe(III) dosage ¼ 0.12 mM, initial pH ¼ 3.0, UV irradiation ¼ 1 h).
decomposed to superoxide and the iminium cation, and the latter one further reacts with H2O to form IDMP (Nowack and Stone, 2003). Simultaneously, another pathway, i.e., $OH-mediated degradation, occurred on the substrate. The initial cleavage of NTMP was identified at the C-N bond, leading to the immediate formation of IDMP and AMPA (evidence by HPLC analyses, as shown in Fig. 5). To further examine the reasonability of the proposed process, the mass balance of elemental P was conducted for photodegradation of Fe(III)-NTMP. As observed in Fig. 5, the TP of four P species, i.e., NTMP and IDMP, AMPA, and phosphate, varied
3.2.4. Co-precipitation for TP removal According to Fig. 2a, Fe(III)/UV/NaOH process could result in 90% TP removal at the Fe(III) dosage of 0.12 mM, while Fe(III)/NaOH could only result in limited TP removal (24%) from the Ca(II)-NTMP solution under similar conditions. Fe(III) precipitation and flocculation was effective for removal of phosphate, however, phosphonates tend to form complexes with typical Fe(III) and Al(III) flocculants, thereby compromising the final TP removal (Rott et al., 2017b). The above results further demonstrated that the UVirradiation unit was essential to enhance the final TP removal during precipitation. We also characterized the solid samples after co-precipitation in terms of SEM, XRD and elemental analysis (Fig. S3-S5). Apparently, the free metal cations, including Ca(II), Fe(II) and Fe(III), NTMP and its intermediate phosphonates, were co-precipitated with the formed phosphate. 3.3. Extension to other phosphonates To further examine the feasibility of the combined Fe(III)/UV/ NaOH process for other phosphonates, we employed Ca-PBTC, CaHEDP, Ca-EDTMP, Ca-DTPMP and Ca-HDTMP as the pollutants of concern. Their degradation properties by Fe(III) were examined in terms of phosphate formation and compared to the Fenton process (Fe(II)/H2O2). The results in Fig. 6 indicated that, for the Fenton process (pH ¼ 3), the concerned phosphonates could be converted
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classical Fenton process on the final TP removal during coprecipitation (pH ¼ 6), and direct Fe(III) coagulation was employed for reference. For all the tested six phosphonates (Fig. 7), pre-oxidation by both processes is essential to improve the TP removal of co-precipitation. Comparatively, the Fenton/NaOH process enhanced the final P removal for phosphonates (>50%), whereas the Fe(III)/UV/NaOH exhibited the most attractive performance, resulting in the final P removal higher than 90% with the residual total P below 0.5 mg/L (Level 1-A standard for Wastewater Treatment Plant in China, GB18918-2002) in all the test cases. 3.4. Validation of the combined process in treatment of real effluent
Fig. 6. The phosphate formation during the decomposition of phosphonates by Fe(III)/ UV and other processes. ([Ca]0 ¼ [phosphonate] ¼ 0.020 mM, initial pH ¼ 3.0, UV irradiation ¼ 1 h, H2O2 dosage ¼ 2 mM, other conditions are listed in Table S2).
Prior to the validation experiments, we examined the effect of natural organic matters (NOMs) and other metal ions (Cu2þ and Ni2þ) on the formation of phosphate and the final TP removal from the Ca(II)-NTMP solution. Results in Fig. S6 suggested that the presence of NOM and both metals at the test levels (5 mg TOC/L for NOM and 0.1 mM for metal ions) did not exert significant effect on the efficiency of the combined process. To validate the feasibility of the combined process in practical applicability, two authentic samples from a WWTP in Jiangyin City (China) was collected for test, the influent and the effluent. Direct Fe(III) coagulation and traditional Fenton followed by precipitation (namely Fenton/NaOH) were also employed for comparison. The initial TP of the test influent was 4.3 mg/L (1.82 mg/L in phosphate and 2.48 mg/L in DOP), while that for the effluent was 0.90 mg/L (0.24 mg/L in phosphate and 0.66 mg/L in DOP). Results in Table 1 indicated that Fe(III)/UV/NaOH exhibited more efficient phosphorous removal than the other two processes. Such satisfactory performance renders us to believe that the combined Fe(III)/UV/NaOH process is a promising option for water decontamination from phosphonates. 4. Conclusions
Fig. 7. Removal of total phosphorous from calcium-phosphonate solutions based on various processes ([Ca]0 ¼ [ligand] ¼ 0.020 mM, initial pH ¼ 3.0, UV irradiation ¼ 1 h, precipitation pH ¼ 6.0, H2O2 dosage ¼ 2 mM, other conditions were listed in Table S2).
to phosphate by the maximum of 20%, whereas for the Fe(III)/UV process, the same phosphonates exhibited much higher degradation efficiency (>60%), and the degradation of PBTC and HEDP even reached up to 90% and 80%, respectively. It seems that the elemental N present in phosphonate adversely affected the formation of phosphate, possibly because part of the hydroxyl radicals has to be consumed on the leakage of the C-N bond. Generally, Fe(III)/UV is more efficient than Fenton process to degrade phosphonates. Also, we further compared the effectiveness of Fe(III)/UV with
In the present study, the application of a proprietary combined process, i.e., Fe(III)/UV/NaOH, was extended to enhanced removal of phosphonate (in the form of Ca(II) complexes). The basic steps for the process included Fe(III) replacement with the complexed Ca(II) to form Fe(III)-phosphonate, UV-mediated degradation of phosphonate to cleave its C-N and C-P bonds and form phosphate and other breakdown products, as well as the final removal of TP through co-precipitation. By utilizing the process, the residual TP of the NTMP solution can be reduced from 1.81 mg/L to below <0.2 mg/L. Test on the treatment of real wastewater further validated the potential of the combined process in enhanced phosphorous removal from water. This study provides a highly efficient option for water decontamination from phosphonate, and further study is required on the pilot-scale operation of the combined process.
Table 1 Treatment of wastewater influent and secondary settling tank effluent by Fe(III)/UV/NaOHa and other processes. Sample
Fe(III)/NaOH
Influentb
c
Effluent
a
Initial TP (mg/L) Residual P (mg/L) TP removal (%) Initial TP (mg/L) Residual P (mg/L) TP removal (%)
Fenton/NaOH
4.30 (1.82 in phosphate and 2.48 in DOP) 2.31 0.93 46.3 78.4 0.90 (0.24 in phosphate and 0.66 in DOP) 0.63 0.38 30.0 57.7
Fe(III)/UV/NaOH 0.23 94.7 0.14 84.4
Initial pH ¼ 3.0, UV irradiation ¼ 1 h, precipitation pH ¼ 6.0. Sampled from a WWTP influent in Jiangyin City (China), the mixture of printing and dyeing wastewater (~40%) and domestic wastewater (~60%). (Fe(III) dosage ¼ 0.25 mM, Fe(II) dosage ¼ 0.25 mM, H2O2 dosage ¼ 2 mM). c Sampled from the secondary settling tank effluent of the identical WWTP in Jiangyin City (China). (Fe(III) dosage ¼ 0.05 mM, Fe(II) dosage ¼ 0.05 mM, H2O2 dosage ¼ 1 mM). b
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Acknowledgements We greatly appreciate the financial support from Natural Science Foundation of China (Grant No. 51761165011) and National Key Research and Development Program of China (Grant No. 2016YFA0203104). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.watres.2019.01.007. References ACS, 2016. Database of. American Chemical Society. https://scifinder.cas.org. Bajt, O., Mailhot, G., Bolte, M., 2001. Degradation of dibutyl phthalate by homogeneous photocatalysis with Fe(III) in aqueous solution. Appl. Catal. B Environ. 33 (3), 239e248. Barrett, K.A., McBride, M.B., 2005. Oxidative degradation of glyphosate and aminomethylphosphonate by manganese oxide. Environ. Sci. Technol. 39 (23), 9223e9228. Chen, Y., Wu, F., Lin, Y., Deng, N., Bazhin, N., Glebov, E., 2007. Photodegradation of glyphosate in the ferrioxalate system. J. Hazard Mater. 148 (1e2), 360e365. Doumic, L.I., Soares, P.A., Ayude, M.A., Cassanello, M., Boaventura, R.A.R., Vilar, V.J.P., 2015. Enhancement of a solar photo-Fenton reaction by using ferrioxalate complexes for the treatment of a synthetic cotton-textile dyeing wastewater. Chem. Eng. J. 277, 86e96. Drzyzga, D., Lipok, J., 2017. Analytical insight into degradation processes of aminopolyphosphonates as potential factors that induce cyanobacterial blooms. Environ. Sci. Pollut. Res. Int. 24 (31), 24364e24375. Davenport, B., DeBoo, A., Dubois, F., Kishi, A., 2000. CEH Report: Chelating Agents. SRI Consulting, Menlo Park, CA, USA. Cited by Nowack (2003). EPA, 2013. Phosphonates in Detergents. EPA Detergent Phosphonates Dossier. European Phosphonate Association. Grandcoin, A., Piel, S., Baures, E., 2017. Aminomethylphosphonic acid (AMPA) in natural waters: its sources, behavior and environmental fate. Water Res. 117, 187e197. Gu, A.Z., Liu, L., Neethling, J.B., Stensel, H.D., Murthy, S., 2011. Treatability and fate of various phosphorus fractions in different wastewater treatment processes. Water Sci. Technol. 63 (4), 804e810. Horstmann, B., Grohmann, A., 1988. Investigations about the biological degradation of phosphonates (Untersuchungen zur biologischen Abbaubarkeit von Phosphonaten). Vom Wasser 70, 163e178. Henneberry, Y.K., Kraus, T.E.C., Fleck, J.A., Krabbenhoft, D.P., Bachand, P.M., Horwath, W.R., 2011. Removal of inorganic mercury and methylmercury from surface waters following coagulation of dissolved organic matter with metalbased salts. Sci. Total Environ. 409 (3), 631e637. Hu, C., Chen, Q., Liu, H., Qu, J., 2015. Coagulation of methylated arsenic from drinking water: influence of methyl substitution. J. Hazard Mater. 293, 97e104. Hug, S.J., Canonica, L., Wegelin, M., Gechter, D., Gunten, U.V., 2001. Solar oxidation and removal of arsenic at circumneutral pH in iron containing waters. Environ. Sci. Technol. 35 (10), 2114e2121. Jaisi, D.P., Li, H., Wallace, A.F., Paudel, P., Sun, M., Balakrishna, A., Lerch, R.N., 2016. Mechanisms of bond cleavage during manganese oxide and UV degradation of glyphosate: results from phosphate oxygen isotopes and molecular simulations. J. Agric. Food Chem. 64 (45), 8474e8482. Jaworska, J., Van Genderen-Takken, H., Hanstveit, A., Van de Plassche, E., Feijtel, T., 2002. Environmental risk assessment of phosphonates used in domestic laundry and cleaning agents in The Netherlands. Chemosphere 47 (6), 655e665. Kononova, S.V., Nesmeyanova, M.A., 2002. Phosphonates and their degradation by microorganisms. Biochemistry 67 (2), 184e195. th, E., Geppert, H., Fischer, T., Liebsch, S., Martienssen, M., 2017. IdenKuhn, R., To tification of the complete photodegradation pathway of ethylenediaminetetra(methylenephosphonic acid) in aquatic solution. Clean. - Soil, Air, Water 45 (5), 1e8. Kuhn, R., Jensch, R., Bryant, I.M., Fischer, T., Liebsch, S., Martienssen, M., 2018. The influence of selected bivalent metal ions on the photolysis of diethylenetriamine penta(methylenephosphonic acid). Chemosphere 210, 726e733. Lesueur, C., Pfeffer, M., Fuerhacker, M., 2005. Photodegradation of phosphonates in water. Chemosphere 59 (5), 685e691.
Majed, N., Li, Y., Gu, A.Z., 2012. Advances in techniques for phosphorus analysis in biological sources. Curr. Opin. Biotechnol. 23 (6), 852e859. Matthijs, E., de Oude, N.T., Bolte, M., Lemaire, J., 1989. Photodegradation of ferric ethylenediaminetetra(methylenephosphonic acid) (EDTMP) in aqueoussolution. Water Res. 23 (7), 845e851. Neft, A., Claus, M.B., Schmidt, S., Minke, R., Steinmetz, H., 2010. Looking closer on phosphorus fractions in WWTP effluents: implications on the optimisation of phosphorus removal. In: Proceedings of the 1st IWA Austrian Young Water Professionals Conference, Vienna. IWA-4306. Nowack, B., 1997. Determination of phosphonates in natural waters by ion-pair high-performance liquid chromatography. J. Chromatogr. A 773 (1e2), 139e146. Nowack, B., 1998. The behavior of phosphonates in wastewater treatment plants of Switzerland. Water Res. 32 (4), 1271e1279. Nowack, B., Stone, A.T., 2000. Degradation of nitrilotris(methylenephosphonic acid) and related (amino)phosphonate chelating agents in the presence of manganese and molecular oxygen. Environ. Sci. Technol. 34 (22), 4759e4765. Nowack, B., 2002. Aminopolyphosphonate removal during wastewater treatment. Water Res. 36 (18), 4636e4642. Nowack, B., 2003. Environmental chemistry of phosphonates. Water Res. 37 (11), 2533e2546. Nowack, B., Stone, A.T., 2003. Manganese-catalyzed degradation of phosphonic acids. Environ. Chem. Lett. 1 (1), 24e31. Nowack, B., Stone, A.T., 2006. Competitive adsorption of phosphate and phosphonates onto goethite. Water Res. 40 (11), 2201e2209. Pan, B.C., Xu, Z., Ren, J., Lv, L., Zhang, W.M., Zhang, S.J., 2014. A combined process for treatment of wastewater containing heavy metal complexes. Chinese Patent, No. ZL 201410264651.1. €nkko € m€ Popov, K., Ro aki, H.L., Lajunen, L.H.J., 2001. Critical evaluation of stability constants of phosphonic acids. Pure Appl. Chem. 73 (10), 1641e1677. Qin, C., Liu, H., Liu, L., Smith, S., Sedlak, D.L., Gu, A.Z., 2015. Bioavailability and characterization of dissolved organic nitrogen and dissolved organic phosphorus in wastewater effluents. Sci. Total Environ. 511, 47e53. Rott, E., Minke, R., Bali, U., Steinmetz, H., 2017a. Removal of phosphonates from industrial wastewater with UV/FeII, Fenton and UV/Fenton treatment. Water Res. 122, 345e354. Rott, E., Minke, R., Steinmetz, H., 2017b. Removal of phosphorus from phosphonateloaded industrial wastewaters via precipitation/flocculation. J. Wat. Proc. Eng. 17, 188e196. Rott, E., Steinmetz, H., Metzger, J.W., 2018a. Organophosphonates: a review on environmental relevance, biodegradability and removal in wastewater treatment plants. Sci. Total Environ. 615, 1176e1191. Rott, E., Nouri, M., Meyer, C., Minke, R., Schneider, M., Mandel, K., DrenkovaTuhtan, A., 2018b. Removal of phosphonates from synthetic and industrial wastewater with reusable magnetic adsorbent particles. Water Res. 145, 608e617. Safarzadeh-Amiri, A., Bolton, J.R., Cater, S.R., 1997. Ferrioxalate-mediated photodegradation of organic pollutants in contaminated water. Water Res. 31 (4), 787e798. Schowanek, D., Verstraete, W., 1990. Phosphonate utilization by bacterial cultures and enrichments from environmental samples. Appl. Environ. Microbiol. 56 (4), 895e903. Shan, C., Xu, Z., Zhang, X., Xu, Y., Gao, G., Pan, B., 2018. Efficient removal of EDTAcomplexed Cu(II) by a combined Fe(III)/UV/alkaline precipitation process: performance and role of Fe(II). Chemosphere 193, 1235e1242. Studnik, H., Liebsch, S., Forlani, G., Wieczorek, D., Kafarski, P., Lipok, J., 2015. Amino polyphosphonates-chemical features and practical uses, environmental durability and biodegradation. N. Biotech. 32 (1), 1e6. Su, C.M., Puls, R.W., 2001. Arsenate and arsenite removal by zerovalent iron: effects of phosphate, silicate, carbonate, borate, sulfate, chromate, molybdate, and nitrate, relative to chloride. Environ. Sci. Technol. 35 (22), 4562e4568. Wang, S., Liu, B., Yuan, D., Ma, J., 2016. A simple method for the determination of glyphosate and aminomethylphosphonic acid in seawater matrix with high performance liquid chromatography and fluorescence detection. Talanta 161, 700e706. Xu, Z., Gao, G., Pan, B., Zhang, W., Lv, L., 2015. A new combined process for efficient removal of Cu(II) organic complexes from wastewater: Fe(III) displacement/UV degradation/alkaline precipitation. Water Res. 87, 378e384. Ye, Y., Jiang, Z., Xu, Z., Zhang, X., Wang, D., Lv, L., Pan, B., 2017. Efficient removal of Cr(III)-organic complexes from water using UV/Fe(III) system: negligible Cr(VI) accumulation and mechanism. Water Res. 126, 172e178. Zhou, Y., Xing, X.H., Liu, Z., Cui, L., Yu, A., Feng, Q., Yang, H., 2008. Enhanced coagulation of ferric chloride aided by tannic acid for phosphorus removal from wastewater. Chemosphere 72 (2), 290e298.
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Highly efficient removal of phosphonates from water by a combined Fe(III)/UV/co-precipitation process Shuhui Sun a, Shu Wang a, Yuxuan Ye a, Bingcai Pan a, b, * a b
State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, China Research Center for Environmental Nanotechnology (ReCENT), Nanjing University, Nanjing 210023, China
a r t i c l e i n f o
a b s t r a c t
Article history: Received 30 September 2018 Received in revised form 7 January 2019 Accepted 11 January 2019 Available online 16 January 2019
Considerable amount of phosphorous is present as organic phosphonates (usually in the form of metal complexes, e.g., Ca(II)-phosphonate) in domestic and industrial effluents, which cannot be effectively removed by traditional processes for phosphate. Herein, we employed a proprietary process, i.e., Fe(III) displacement/UV irradiation/co-precipitation (denoted Fe(III)/UV/NaOH), to enable an efficient removal of Ca(II)-phosphonate complexes from water. The combined process includes three basic steps, i.e., Fe(III) replacement with the complexed Ca(II) to form Fe(III)-phosphonate of high photo-reactivity, UV-mediated degradation of Fe(III)-phosphonate to form phosphate and other intermediates, and the final phosphorous removal via co-precipitation at pH ¼ 6. The operational conditions for the combined process to remove a typical phosphonate Ca(II)-NTMP (nitrilotrismethylenephosphonate) are optimized, where ~60% NTMP is transformed to phosphate with the total phosphorous reduction from 1.81 mg/L to 0.17 mg/L. Under UV irradiation, the cleavage of NTMP is identified at the C-N and C-P bonds to form the intermediate products and phosphate in sequence. Also, the combined process is employed for treatment of two authentic effluents before and after activated sludge treatment, resulting in the phosphorous drop from 4.3 mg/L to 0.23 mg/L and from 0.90 mg/L to 0.14 mg/L respectively, which is much superior to other processes including Fenton/co-precipitation. In general, the combined process exhibits great potential for efficient removal of phosphonates from contaminated waters. © 2019 Elsevier Ltd. All rights reserved.
Keywords: Phosphonates Photolysis Fe(III) replacement Advanced treatment
1. Introduction Phosphonates are anthropogenic chelating agents widely used as scale inhibitors, corrosion inhibitors, and laundry detergents (Rott et al., 2018a), featuring one or more phosphonic acid groups [C-PO(OH)2] (Jaworska et al., 2002; Nowack, 2003). The global phosphonate consumption was 56,000 t/a in 1998 (Davenport et al., 2000) and increased to 94,000 t/a in 2012 (EPA, 2013). Fig. 1 gives an overview of the formula structure of six typical phosphonates, including 2-phosphonobutane-1,2,4-tricarboxylic acid (PBTC), 1hydroxyethane 1,1-diphosphonic acid (HEDP), nitrilotris(methylene phosphonic acid) (NTMP), ethylenediamine tetra(methylene phosphonic acid) (EDTMP), diethylenetriamine penta(methylene phosphonic acid) (DTPMP) and hexamethylenediamine tetra(methylene phosphonic acid) (HDTMP).
* Corresponding author. Research Center for Environmental Nanotechnology (ReCENT), Nanjing University, Nanjing 210023, China. E-mail address: [email protected] (B. Pan). https://doi.org/10.1016/j.watres.2019.01.007 0043-1354/© 2019 Elsevier Ltd. All rights reserved.
Except for inorganic phosphate, phosphonate is one of the key species of elemental phosphorous in industrial and municipal effluents as well as in natural waters (Nowack and Stone, 2006; Gu et al., 2011; Majed et al., 2012; Qin et al., 2015). Though few open data are available on the phosphate concentration in industrial effluents, we sampled several biotreated textile effluents and discriminated the six phosphonates (in Fig. 1) with their total amount reaching 0.1e0.5 mg P/L. In addition, phosphonates have been identified in the influents of German and Swiss WWTPs (Nowack, 1998, 2002). For instance, NTMP of 0.2e1.1 mM was detected in the influent of a WWTP as influenced by textile industry (Nowack, 2002), with an average DTPMP concentration of 0.12 mM in the effluent (Nowack, 1998). It is assumed that in Europe, 9000e18,600 t/a of phosphonate is finally discharged into receiving waters (Rott et al., 2018b). Phosphonates can hardly be biomineralized during activated sludge treatment, either aerobically (Horstman and Grohmann, 1988) or anaerobically (Nowack, 1998). Traditional processes such as adsorption and precipitation/flocculation could enable highly efficient removal of inorganic phosphate. However, they usually work ineffectively for organic phosphonates
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Fig. 1. Molecular structure of several typical phosphonates (based on ACS, 2016).
(Neft et al., 2010; Gu et al., 2011; Rott et al., 2018a). The complexation of phosphonates with flocculants results in an excessive dosage of metal salts and adversely affect the final removal of total phosphorous (TP) (Nowack, 2006; Rott et al., 2017b). Consequently, an increasing fraction of dissolved organic phosphorus (DOP) detected in the effluent of wastewater treatment plants (WWTPs) (Rott et al., 2018a). On the other side, phosphonates usually remain in the adsorbed form in activated sludge and metal precipitates (Nowack, 2002) after traditional wastewater treatment processes (Rott et al., 2017b). During the disposal of the solid residues (for instance, used in agriculture), phosphonates are brought into the environment. Also, considerable amounts of phosphonates may enter the receiving bodies via the discharge of treated effluent. Phosphonates are subject to natural elimination in environment (Nowack and Stone, 2000; Matthijs et al., 1989; Schowanek and Verstraete, 1990), accompanying a long-term release of bioavailable phosphate in water. Moreover, there still exists substantial risk for phosphonate transformation to more toxic aminomethylphosphonic acid (AMPA) particularly in the presence of sunlight (Grandcoin and Baures, 2017). Thus, the contribution of phosphonates to water eutrophication and environmental safety should be particularly concerned (Studnik et al., 2015; Qin et al., 2015; Drzyzga and Lipok, 2017), and development of efficient and cost effective approaches to realize phosphonate removal from contaminated waters is urgently desired. As well known, decarboxylation of Fe(III)-carboxyl complexes could proceed via ligands to metals charge transfer (LMCT) under UV irradiation. Based on the principle, a combined process, i.e., Fe(III) displacement/UV irradiation/precipitation has been recently developed to effectively remove various toxic metal-carboxyl complexes from water, including Cu(II)-citrate (Pan et al., 2014; Xu et al., 2015) and Cu(II)-EDTA (Shan et al., 2018). In brief, most metal-carboxyl complexes are inert to UV irradiation, and the free
Fe(III) ions was deliberately added to the metal-carboxyl complex solution to form Fe(III)-carboxyl complexes of high photo-reactivity because Fe(III) displays much higher complexation constants with carboxyl ligands, where free toxic metal ions are simultaneously released for subsequent removal by chemical precipitation. Later, it was successfully extended to remove Cr(III)-citrate from synthetic solution and authentic tannery effluent despite of its extremely slow decomplexation rate (Ye et al., 2017), where the hydroxyl radicals formed during Fe(III)/UV process tended to oxide the complexed Cr(III) to free Cr(VI), greatly speeding up the decomplexation and driving a fast formation of Fe(III)-citrate for photolysis. Moreover, the toxic Cr(VI) was effectively reduced back to free Cr(III) for co-precipitation in the presence of Fe(II) formed during photolysis. As for phosphonates, limited study has been available on the photo-reactivity of their Fe(III) complexes in the past decades. Matthijs et al. (1989) reported Fe(III)-EDTMP could be photodegraded through a photoinduced ligand-to-metal charge transfer procedure under sunlight irradiation. Lesueur et al. (2005) further confirmed the Fe(III) mediated photolysis of four common aminophosphonates during UV irradiation. Kuhn et al. (2018) evidenced that during photo-degradation of DTPMP in the presence of Fe(II), the initial cleavage of DTPMP was initiated at the C-N bond, resulting in the gradual drop of C-PO(OH)2 groups in the phosphonate. The above study inspired us to consider the potential of photo-assisted process for water decontamination from phosphonates. In this study, we demonstrated the applicability of the Fe(III)/ UV/NaOH process to eliminate various Ca(II)-phosphonate complexes from synthetic solution and authentic effluents. NTMP was employed as the model phosphonate to optimize the operational conditions and elucidate the underlying mechanism. It is a typical aminophosphonate characterized by three C-PO(OH)2 groups and
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usually present as metal complex, e.g., Ca(II)-NTMP complex, in natural waters or biotreated effluents (Nowack, 2003). Afterward, the process was extended to removal of other Ca(II)-phosphonates, and some typical processes including Fenton/precipitation were employed for comparison. Two authentic effluents were sampled to further demonstrate the applicability of the Fe(III)/UV/NaOH process. 2. Materials and methods 2.1. Materials and instruments All the reagents used in the experiments are of analytical reagent grade without further purification. PBTC (50% aqueous solution) was obtained from Matrix Scientific Co., USA. Solid HEDP$H2O (95%), NTMP (97%) and iminodi (methylenephosphonic) acid (IDMP, 97%) were purchased from SigmaAldrich (St. Louis, MO, USA). EDTMP (98%), HDTMP (97%) and DTPMP (50% Aqueous solution) were provided by Aladdin Biochemical Technology Co., Ltd, China. AMPA (99.0%) and 9fluorenylmethylchloroformate FMOC-Cl (98.0%) were purchased from J & K Scientific Co., Ltd., China. HPLC-grade methanol (MeOH) and acetonitrile (ACN) were obtained from Merck Co., Germany. Other chemicals, including FeCl3$6H2O, FeSO4$7H2O, CaCl2, CuCl2, NiSO4, H2O2, NaOH and HCl, were purchased from Sinopharm Chemical Reagent Co., China. The sample of natural organic matter (NOM) was purchased from the International Humic Substances Society (Catalog No. 2R101N, Suwannee River NOM). The stock solution of Ca-NTMP, Ca-PBTC, Ca-HEDP, Ca-EDTMP, Ca-DTPMP and Ca-HDTMP were prepared by mixing CaCl2 with the corresponding ligand acids in the molar ratio of 1:1. All the solutions were prepared with ultrapure water, and pH was adjusted by HCl or NaOH. The authentic water samples (influent and effluent) were collected from a WWTP (located in Jiangyin City, China) based on A/O process, as designed to treat the mixture of printing and dyeing wastewater (~40%) and domestic wastewater (~60%) with the total capacity of 10000 m3 per day. UV irradiation experiments were carried out in a rotating disk photoreactor (Nanjing Stone Tech Electric Equipment Co., China) with a 300-W medium pressure mercury lamp (Shanghai Hongguang Tungsten & Molybdenum Technology Co., Ltd.) as the UV source. The light intensity was about 2.0 mW/cm2 measured at 365 nm by a radiometer (Photoelectric Instrument Factory of Beijing Normal University, China). 2.2. Experimental procedures The combined process, i.e., Fe(III)/UV/NaOH, included two main steps in sequence: (1) FeCl3 of preset amount was added into solutions of Ca-NTMP and other Ca-phosphonates, then the mixture was diluted to 50 mL by using ultra-pure water to keep the initial P at 1.81 mg/L (0.02 mM) for NTMP. After the addition of Fe(III), pH was adjusted to a predetermined value. Then, each solution was subject to UV irradiation. Methanol (100 mM) was added as the scavenger of $OH radicals to examine its possible role during UV irradiation. (2) After UV irradiation, the pH value increased by the maximum of 0.3 unit. NaOH was further added to the solution to achieve a preset pH, then the mixture was kept stirring at 60 rpm for 30 min and stood for 60 min. Afterward, the supernatant was filtered through a 0.22-mm membrane, and the concentration of concerned species in solution, including phosphonates and some of their intermediate products, TP, was determined.
23
For comparison, Fenton/NaOH, Fenton/UV/NaOH, and UV/H2O2, were performed following similar procedures described below. Whether the chemicals and UV irradiation were employed depended on the nature of the processes. Generally, they involved two steps in sequence: (1) FeSO4 was added to 50-mL phosphonate solution (0.02 mM) and the initial pH was adjusted to a predetermined value, where H2O2 was further added to initiate the reaction (FeSO4 free for UV/H2O2). After stirring for 1 h and under the UV irradiation (not for Fenton), the formed PO3 4 was determined. (2) NaOH was added to the pretreated solution to achieve a preset pH, then the mixture kept stirring at 60 rpm for 30 min and stood for 60 min. Afterward, the supernatant was filtered through a 0.22-mm membrane, and the TP concentration was determined. Detailed conditions of the above processes are available in the related figures and tables. 2.3. Analytical methods TP and PO3 4 were determined by molybdenum blue spectrophotometric method (Su, 2001) with a UV-vis spectrometer (T6, PGENERAL, China) at the wavelength of 700 nm. For TP determination, sample digestion using peroxodisulfate (K2S2O8) was required prior to analysis. The concentration of Fe(III)-NTMP was measured by HPLC (Waters 1525) equipped with a UV detector (Waters 2489) (Nowack, 1997). A C18 column (150 mm 4.6 mm, 5 mm particle size, Sunfire, Waters) was used for chromatographic separation, with experimental conditions detailed in Table S1 in Supplementary Materials. AMPA and iminodi(methylenephosphonic acid) (IDMP) is determined by HPLC-fluorescence detector (Waters 2745) with pre-column FMOC-Cl derivatization according to the previous method (Wang et al., 2016). 3. Results and discussion 3.1. Validation of the Fe(III)/UV/NaOH process 3.1.1. Effect of Fe(III) dosage The feasibility of the combined process, i.e., Fe(III)/UV/NaOH, on the removal of Ca(II)-NTMP was initially probed. For comparison, another combined process of UV free, i.e., Fe(III) addition followed by precipitation (namely Fe(III)/NaOH), was employed, and the results are depicted in Fig. 2a. In the set of the experiments, the concentration of Fe(III) varied from 0.040 to 0.20 mM, while the initial pH was set as 3.0, the UV irradiation time was 60 min, and the precipitation pH was optimized at 6.0 (Fig. 2d). As observed, Fe(III)/NaOH could only result in limited TP removal (49%) from the Ca(II)-NTMP solution even at the Fe(III) dosage of 0.20 mM. Comparatively, UV/Fe(III)/NaOH could result in nearly 100% TP removal at the Fe(III) dosage of 0.12 mM or higher. Such relatively low removal by the Fe(III)/NaOH process (<50%) is possibly due to the poor precipitation of ferric hydroxide in the presence of organic ligands (Zhou et al., 2008; Henneberry et al., 2011; Hu et al., 2015), which compete with hydroxy anions to bind Fe(III). Obviously, the UV-irradiation unit is indispensable to the combined process for enhanced TP removal. Also, the formation of phosphate from the Ca(II)-NTMP solution was detected after Fe(III) addition and UV irradiation. As seen in Fig. 2a, the production rate of phosphate increased with the increasing Fe(III) addition, which is consistent with the final TP removal. As the Fe(III) dosage increased to 0.12 mM, the phosphate formation reached 61%, and the residual TP after precipitation dropped down to <0.3 mg/L. Further Fe(III) addition to 0.2 mM offered limited promotion to TP removal. In the subsequent
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Fig. 2. (a) Effect of Fe(III) dosage on Fe(III)-mediated processes for removal of Ca-NTMP complexes. Fe(III)/NaOH: Fe(III) addition followed by precipitation; Fe(III)/UV/NaOH: Fe(III)/ UV process followed by co-precipitation; (b) Effect of UV irradiation duration on the formation of PO3 4 and TP removal by the Fe(III)/UV/NaOH process; (c) Effect of initial pH on the formation of PO3 4 by the Fe(III)/UV process; (d) Effect of precipitation pH on the residual TP by the Fe(III)/UV/NaOH process ([Ca-NTMP]0 ¼ 0.020 mM (1.81 mg TP/L), Fe(III) dosage ¼ 0.12 mM for b, c and d, initial pH ¼ 3.0 for a, b, and d, UV irradiation ¼ 1 h for a and d, precipitation pH ¼ 6.0 for a, b, and c).
experiments, we set the optimal Fe(III) dosage as 0.12 mM for the Fe(III)/UV/NaOH process except for otherwise specified. 3.1.2. Effect of UV irradiation time The influences of UV irradiation duration on the formation of PO3 4 as well as the final removal of TP are illustrated in Fig. 2b. The PO3 4 formation reached ~40% after 10-min UV irradiation, indicating the cleavage of the C-P bond of NTMP. Similar to the effect of Fe(III) dosage, increasing the UV irradiation time is favorable for the final TP removal. In the initial 10 min, about 40% PO3 4 was formed and nearly 80% TP was removed after precipitation, indicating that NTMP was partially transformed to low-molecular-weight phosphonates, which was readily removed during Fe(III) coagulation. After 60-min irradiation, the phosphate formation reached 60% with the residual TP below 0.1 mg/L after precipitation. The above results implied that the formation of phosphate could be an indicator for the final TP removal. In the forthcoming experiments, we set 60 min as the optimal UV irradiation time for the Fe(III)/UV/ NaOH process except for otherwise specified. 3.1.3. Effect of initial pH The effect of initial pH (1.5e10.0) on the phosphate formation during UV irradiation was evaluated, and the results are shown in Fig. 2c. As seen, pH ¼ 3.0 is the optimal value for phosphate formation (65% at pH 3.0), and higher pH values resulted in lower efficiency of NTMP photolysis and phosphate formation. At
pH ¼ 2.5e3.0, Fe(III) is mainly present in the form of FeOH2þ, displaying the highest photo-reactivity among all the Fe(III) species to generate $OH radicals based on eq. (1) (Hug et al., 2001). At pH<2 and pH>3, the predominant Fe(III) species turned to be Fe3þ and ferric hydroxide colloids/sediments, respectively, both suppressing the generation of $OH radicals intensively (Bajt et al., 2001) and resulting in a low PO3 4 formation. At alkaline pHs, phosphonates preferably bind Fe(III), and the Fe(III)-phosphonate complexes are attacked by hydroxyl radicals more easily than the counterpart phosphonates (Rott et al., 2017a). Note that phosphate formation still reached 40e50% at neutral or alkaline pHs. Also, the initial pHs were set in the range of 2.5e3.0 for the subsequent Fe(III)/UV/NaOH processes except for otherwise specified. FeOH2þ þ hv / $Fe2þ þ $OH
(1)
3.1.4. Effect of the precipitation pH Effect of the precipitation pH on TP removal is described in Fig. 2d. As seen, the pH values from 4.5 to 6.0 resulted in the lowest residual TP, whereas at pH from 8.0 to 12.0, the residual TP remained at a relatively high level. Such pH-dependent precipitation correlated well with the zeta potential of the resultant flocs (Fig. S1), which exhibited a dramatic drop from 1.3 mV to ~ 40 mV at the pH range from 5.0 to 8.0. In theoretical viewpoint, the higher the pH, the stronger the electrostatic charge repulsion between phosphate/phosphonates and the formed iron hydroxide
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surface. In the following section, the optimal pH was set as 6.0 for precipitation. 3.2. Mechanism of the Fe(III)/UV/NaOH process 3.2.1. Fe(III) displacement The stability constant of Ca(II)-NTMP (logK ¼ 7.6) is much lower than that of Fe(III)-NTMP (logK ¼ 21.2) (Popov et al., 2001), and Fe(III) ions are capable of displacing Ca(II) from Ca(II)-NTMP in thermodynamic viewpoint. The effectiveness of Fe(III) displacement was reflected by the formed Fe(III)-NTMP complexes, as determined by ion-pair high-performance liquid chromatography (Nowack, 1997). Expectedly, the addition of Fe(III) resulted in an equimolar formation of Fe(III)-NTMP complexes (Fig. S2), indicating that Fe(III) could displace equimolar Ca(II) from Ca(II)-NTMP complexes to release the free Ca(II) ions (eq. (2)). Ca(II)-NTMP þ Fe(III) /Fe(III)-NTMP þ Ca(II)
(2)
3.2.2. Role of ·OH radicals As previously reported (Bajt et al., 2001; Hug et al., 2001), photoexcitation of FeOH2þ would result in the formation of $OH radicals (eq. (1)), which was demonstrated by EPR spectra of Fe(III)/ UV system recorded in Fig. 3a. Clearly, the signals of the $OH radicals without UV irradiation were ignorable, while strong signals were detected in the Fe(III)/UV system even in the first 2 min, indicating the generation of massive $OH radicals. As the irradiation proceeded after 20 min, the intensity of $OH signals turned much weaker, possibly due to the consumption of both the radicals and the FeOH2þ species. To further explore the role of $OH radicals in NTMP degradation, we examined the effect of methanol, a typical $OH radical scavenger, on the formation of PO3 4 . As shown in Fig. 3b and 20-min irradiation could result in the efficiency of phosphate transformation higher than 50% for the Fe(III)/UV process. The addition of 100 mM methanol significantly inhibited the phosphate formation, and 60 min was required to achieve similar efficiency. Thus, we believe that $OH radicals play a significant role in the formation of PO3 4 . Note that such adverse effect from the added methanol gradually turned weaker as the irradiation proceeded. This is mainly because another important process, i.e., direct photolysis of Fe(III)-phosphonate, contributed significantly to the formation of phosphate, as discussed below.
25
3.2.3. Photolysis of Fe(III) complexes It is well known that ferrous carboxylates could be photodegraded under UV irradiation via LMCT process (eq. (3)), such as Fe(III)-citrate (Xu et al., 2015; Ye et al., 2017), Fe(III)-EDTA (Shan et al., 2018), and Fe(III)-oxalate (Safarzadeh-Amiri et al., 1997; Chen et al., 2007; Doumic et al., 2015). [Fe3þL] þ hv / Fe2þ þ L
(3)
In fact, similar LMCT process occurred for Fe(III)-phosphonate complexes. Generally, phosphonate in natural water or bioeffluents is mainly present in the form of metal complexes (Ca2þ, Mg2þ, etc.), most of which are of poor photo-reactivity and insensitive to UV irradiation. In theory, Fe(III) could displace the complexed metals (such as Ca(II)) to form Fe(III)-phosphonates due to their stability constants several orders in magnitude higher than the former ones, and the formed Fe(III)-phosphonates would be effectively photodegraded under UV irradiation. For instance, no photodegradation of HEDP was observed in distilled water or in the background of Ca(II), however, the addition of Fe(III) or Cu(II) resulted in a rapid photodegradation of HEDP (Nowack, 2003). Matthijs (1989) reported that Fe(III)-EDTMP could be photo-degraded via stepwise processes, i.e., from the parent compound to ethylenediaminetrimethylenephosphonate, ethylenediaminedimethylene phosphonate, and ethylenediaminemonomethylenephosphonate in sequence, the latter of which is stable in the presence of Fe(III) and light irradiation. Kuhn et al. (2017) identified IDMP as the major breakdown product of EDTMP during UV irradiation of non-complex EDTMP. Kuhn et al. (2018) found that in the presence of Fe(III) and UV irradiation, the C-N bond of DTPMP was cleaved, resulting in the formation of IDMP, ethylamino(bismethylenephosphonic acid) (EABMP) and AMPA. Based on the above discussion, the degradation scheme of Ca(II)-NTMP is proposed in Fig. 4. First, Fe(III) replaced Ca(II) from Ca(II)-NTMP to form Fe(III)-NTMP, where the free Ca(II) is released stoichiometrically. Afterward, the UV irradiation activates the photolysis of Fe(III)-phosphonate complexes based on LMCT process, resulting in the generation of Fe(II) and a phosphonyl radical (Kononova and Nesmeyanova, 2002). Then, PO(OH)O$ donates one electron, leading to the formation of the carbon-centered radical (PO(OH)2-(CH2)2-N-$CH2) and the release of phosphate (Barrett and McBride, 2005; Jaisi et al., 2016). The methylene radical is rapidly intercepted by O2 to form a peroxyl radical, which will be
Fig. 3. (a) EPR spectra of the Fe(III)/UV process under different irradiation duration (initial pH ¼ 3.0, [Ca-NTMP]0 ¼ 0.020 mM (1.81 mg TP/L), Fe(III) dosage ¼ 0.120 mM); (b) Effect of methanol on the formation of PO3 4 ([Ca-NTMP]0 ¼ 0.020 mM, initial pH ¼ 3.0, Fe(III) dosage ¼ 0.120 mM).
26
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Fig. 4. Proposed reaction pathway for transformation of Ca-NTMP for Fe(III)/UV unite of during the Fe(III)/UV/NaOH process.
insignificantly during the whole process, further validating the reliability of the above mechanism. A slight drop in the TP of the four P species during the initial reaction indicated the formation of other intermediate phosphonates, which were transformed to IDMP, AMPA and phosphate eventually.
Fig. 5. Profiles of the degradation products of Ca-NTMP during Fe(III)/UV process. [TP*] ¼ [PO3 4 ]þ[AMPA]þ[IDMP]þ[NTMP] (Fe(III) dosage ¼ 0.12 mM, initial pH ¼ 3.0, UV irradiation ¼ 1 h).
decomposed to superoxide and the iminium cation, and the latter one further reacts with H2O to form IDMP (Nowack and Stone, 2003). Simultaneously, another pathway, i.e., $OH-mediated degradation, occurred on the substrate. The initial cleavage of NTMP was identified at the C-N bond, leading to the immediate formation of IDMP and AMPA (evidence by HPLC analyses, as shown in Fig. 5). To further examine the reasonability of the proposed process, the mass balance of elemental P was conducted for photodegradation of Fe(III)-NTMP. As observed in Fig. 5, the TP of four P species, i.e., NTMP and IDMP, AMPA, and phosphate, varied
3.2.4. Co-precipitation for TP removal According to Fig. 2a, Fe(III)/UV/NaOH process could result in 90% TP removal at the Fe(III) dosage of 0.12 mM, while Fe(III)/NaOH could only result in limited TP removal (24%) from the Ca(II)-NTMP solution under similar conditions. Fe(III) precipitation and flocculation was effective for removal of phosphate, however, phosphonates tend to form complexes with typical Fe(III) and Al(III) flocculants, thereby compromising the final TP removal (Rott et al., 2017b). The above results further demonstrated that the UVirradiation unit was essential to enhance the final TP removal during precipitation. We also characterized the solid samples after co-precipitation in terms of SEM, XRD and elemental analysis (Fig. S3-S5). Apparently, the free metal cations, including Ca(II), Fe(II) and Fe(III), NTMP and its intermediate phosphonates, were co-precipitated with the formed phosphate. 3.3. Extension to other phosphonates To further examine the feasibility of the combined Fe(III)/UV/ NaOH process for other phosphonates, we employed Ca-PBTC, CaHEDP, Ca-EDTMP, Ca-DTPMP and Ca-HDTMP as the pollutants of concern. Their degradation properties by Fe(III) were examined in terms of phosphate formation and compared to the Fenton process (Fe(II)/H2O2). The results in Fig. 6 indicated that, for the Fenton process (pH ¼ 3), the concerned phosphonates could be converted
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27
classical Fenton process on the final TP removal during coprecipitation (pH ¼ 6), and direct Fe(III) coagulation was employed for reference. For all the tested six phosphonates (Fig. 7), pre-oxidation by both processes is essential to improve the TP removal of co-precipitation. Comparatively, the Fenton/NaOH process enhanced the final P removal for phosphonates (>50%), whereas the Fe(III)/UV/NaOH exhibited the most attractive performance, resulting in the final P removal higher than 90% with the residual total P below 0.5 mg/L (Level 1-A standard for Wastewater Treatment Plant in China, GB18918-2002) in all the test cases. 3.4. Validation of the combined process in treatment of real effluent
Fig. 6. The phosphate formation during the decomposition of phosphonates by Fe(III)/ UV and other processes. ([Ca]0 ¼ [phosphonate] ¼ 0.020 mM, initial pH ¼ 3.0, UV irradiation ¼ 1 h, H2O2 dosage ¼ 2 mM, other conditions are listed in Table S2).
Prior to the validation experiments, we examined the effect of natural organic matters (NOMs) and other metal ions (Cu2þ and Ni2þ) on the formation of phosphate and the final TP removal from the Ca(II)-NTMP solution. Results in Fig. S6 suggested that the presence of NOM and both metals at the test levels (5 mg TOC/L for NOM and 0.1 mM for metal ions) did not exert significant effect on the efficiency of the combined process. To validate the feasibility of the combined process in practical applicability, two authentic samples from a WWTP in Jiangyin City (China) was collected for test, the influent and the effluent. Direct Fe(III) coagulation and traditional Fenton followed by precipitation (namely Fenton/NaOH) were also employed for comparison. The initial TP of the test influent was 4.3 mg/L (1.82 mg/L in phosphate and 2.48 mg/L in DOP), while that for the effluent was 0.90 mg/L (0.24 mg/L in phosphate and 0.66 mg/L in DOP). Results in Table 1 indicated that Fe(III)/UV/NaOH exhibited more efficient phosphorous removal than the other two processes. Such satisfactory performance renders us to believe that the combined Fe(III)/UV/NaOH process is a promising option for water decontamination from phosphonates. 4. Conclusions
Fig. 7. Removal of total phosphorous from calcium-phosphonate solutions based on various processes ([Ca]0 ¼ [ligand] ¼ 0.020 mM, initial pH ¼ 3.0, UV irradiation ¼ 1 h, precipitation pH ¼ 6.0, H2O2 dosage ¼ 2 mM, other conditions were listed in Table S2).
to phosphate by the maximum of 20%, whereas for the Fe(III)/UV process, the same phosphonates exhibited much higher degradation efficiency (>60%), and the degradation of PBTC and HEDP even reached up to 90% and 80%, respectively. It seems that the elemental N present in phosphonate adversely affected the formation of phosphate, possibly because part of the hydroxyl radicals has to be consumed on the leakage of the C-N bond. Generally, Fe(III)/UV is more efficient than Fenton process to degrade phosphonates. Also, we further compared the effectiveness of Fe(III)/UV with
In the present study, the application of a proprietary combined process, i.e., Fe(III)/UV/NaOH, was extended to enhanced removal of phosphonate (in the form of Ca(II) complexes). The basic steps for the process included Fe(III) replacement with the complexed Ca(II) to form Fe(III)-phosphonate, UV-mediated degradation of phosphonate to cleave its C-N and C-P bonds and form phosphate and other breakdown products, as well as the final removal of TP through co-precipitation. By utilizing the process, the residual TP of the NTMP solution can be reduced from 1.81 mg/L to below <0.2 mg/L. Test on the treatment of real wastewater further validated the potential of the combined process in enhanced phosphorous removal from water. This study provides a highly efficient option for water decontamination from phosphonate, and further study is required on the pilot-scale operation of the combined process.
Table 1 Treatment of wastewater influent and secondary settling tank effluent by Fe(III)/UV/NaOHa and other processes. Sample
Fe(III)/NaOH
Influentb
c
Effluent
a
Initial TP (mg/L) Residual P (mg/L) TP removal (%) Initial TP (mg/L) Residual P (mg/L) TP removal (%)
Fenton/NaOH
4.30 (1.82 in phosphate and 2.48 in DOP) 2.31 0.93 46.3 78.4 0.90 (0.24 in phosphate and 0.66 in DOP) 0.63 0.38 30.0 57.7
Fe(III)/UV/NaOH 0.23 94.7 0.14 84.4
Initial pH ¼ 3.0, UV irradiation ¼ 1 h, precipitation pH ¼ 6.0. Sampled from a WWTP influent in Jiangyin City (China), the mixture of printing and dyeing wastewater (~40%) and domestic wastewater (~60%). (Fe(III) dosage ¼ 0.25 mM, Fe(II) dosage ¼ 0.25 mM, H2O2 dosage ¼ 2 mM). c Sampled from the secondary settling tank effluent of the identical WWTP in Jiangyin City (China). (Fe(III) dosage ¼ 0.05 mM, Fe(II) dosage ¼ 0.05 mM, H2O2 dosage ¼ 1 mM). b
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