Environ. Sci. Technol. 2006, 40, 3057-3063
Phenols and Amine Induced HO• Generation During the Initial Phase of Natural Water Ozonation MARC-OLIVIER BUFFLE AND URS VON GUNTEN* Swiss Federal Institute of Aquatic Science & Technology (EAWAG), Uberlandstrasse 133, Postfach 611, CH-8600 Dubendorf, Switzerland
The initial phase of ozone decomposition in natural water (t < 20 s) is poorly understood. It has recently been shown to result in very high transient HO• concentrations and, thereby, plays an essential role during processes such as bromate formation or contaminants oxidation. Phenols and amines are ubiquitous moieties of natural organic matter. Naturally occurring concentrations of primary, secondary, and tertiary amines, amino acids, and phenol were added to surface water, and ozone decomposition as well as HO• generation were measured starting 350 milliseconds after ozone addition. Six seconds into the process, 5 µM of dimethylamine and phenol had generated ∫HO•dt ) 1 × 10-10 M‚s and 1.8 × 10-10 M‚s, respectively. With 10 µM dimethylamine and 1.5 mgO3/L, Rct, (∫HO•dt/ ∫O3dt) reached 10-6, which is larger than in advanced oxidation processes (AOP) such as O3/H2O2. Experiments in the presence of HO•-scavengers indicated that a significant fraction of phenol-induced ozone decomposition and HO• generation results from a direct electron transfer to ozone. For dimethylamine, the main mechanism of HO• generation is direct formation of O2•- which reacts selectively with O3 to form O3•-. Pretreatment of phenol-containing water with HOCl or HOBr did not decrease HO• generation, while the same treatment of dimethylamine-containing water considerably reduced HO• generation.
Introduction Ozone is a strong oxidant used in water treatment for taste, odor and color removal, oxidation, and disinfection. Current research also demonstrates that emerging contaminants such as cyanotoxins, hormones, and antibiotics are efficiently oxidized and biochemically inactivated by ozone (1-5). Biphasic Kinetics of Ozone Decomposition. Ozone decomposition in natural water can be kinetically and mechanistically divided into an initial and a second phase (6-8) (Figure 1a, adapted from ref 7). During the initial phase (t < ∼20 s), rapid direct reactions of ozone with specific NOM moieties and some inorganic compounds consume a large fraction of the added ozone (often called instantaneous ozone demand). During this phase, ozone does not follow an apparent first-order rate law as during the second phase (Figure 1a). In fact, k′O3 [s-1] increases following a power function with t f 0 (6). Very high yields of HO• are generated and Rct ()∫HO•dt/∫O3dt) * Corresponding author phone: ++41 44 8235270; fax: ++41 44 8235210; e-mail:
[email protected]. 10.1021/es052020c CCC: $33.50 Published on Web 03/31/2006
2006 American Chemical Society
also increases following a power function with t f 0 (7). Interestingly, although HO• transient concentrations are very high during the initial phase, ozone decomposition seems not to be controlled by the radical chain reaction as in the second phase (6). During the second phase, ozone decomposition follows an apparent first-order rate law, i.e., k′O3 [s-1] remains constant and is 10-100 times smaller than during the initial phase. Rct is also 10-100 times smaller and constant (7). The most reactive moieties of NOM have reacted with ozone during the initial phase so that ozone decomposition in the second phase is mostly controlled by a radical chain reaction and not by its direct reaction with those moieties (Figure 1b). The radical chain reaction is described exhaustively elsewhere (9, 10). In short, the reaction is partially initiated by O3 reaction with HO- to form HO2-. O3 reacts again with HO2- and generates O2•-, which reacts very selectively with another O3 to form O3•-, readily generating HO•. HO• reaction with certain NOM moieties (promoters) leads to carbon centered radicals which upon O2 addition form more O2•-, and so on. This part of the reaction sequence is called propagation. The chain reaction is terminated upon the reaction of HO• with compounds (inhibitors, e.g., carbonate, tert-butyl alcohol) that do not lead to O2•- formation. Importance of the Initial Phase. Given the characteristics of the initial phase-rapid ozone decomposition and very high relative concentration of HO•-, its investigation is essential to better comprehend a number of key mechanisms involved in water ozonation. For example, the formation of potentially carcinogenic bromate during ozonation of bromide-containing water is in part due to HO• generated during the initial phase (11). By reducing HO• exposure during the initial phase through pre-chlorination and ammonia addition, bromate formation could be significantly decreased (11). Another example is that of certain compounds considered refractory to ozone oxidation (e.g., iopromide) but nevertheless showing significant degradation during wastewater ozonation (3). This effect can be well explained by very high HO• exposures measured during ozone decomposition in wastewater, which is mechanistically similar to the initial phase in natural water (6, 7). A further example is ozone’s specificity which, though advantageous when selectively oxidizing biochemically active moieties of pharmaceutical compounds, might be compromised during the initial phase if large fractions of compounds are oxidized by unselective HO•, thus inducing primary metabolites that might not be biochemically inactive (for an extensive discussion see ref 5). In this paper we investigate mechanisms responsible for the initial phase and propose moieties of the NOM that might simultaneously generate high ozone decomposition rates and high HO• yields. Presence of Ozone-Reactive Moieties in NOM. Natural organic matter (NOM) consists of an infinite variety of organic molecules, ranging from low ( 20 s (solid squares) in Lake Zurich water at pH 8 and 2.4 mgO3/L (50 µM). (b) During the initial phase some NOM moieties react directly with ozone to form O2•- or O3•-, while during the second phase some NOM moieties promote the radical chain induced ozone decomposition by reacting with HO• and subsequently O2 to release O2•- (9). typically not known, with the exception of hexosamines which have been measured in the range of 0.18-0.85 µM in low productivity lakes (12). Hexosamines, however, are typically acetylated in the natural environment, which essentially inactivates them with regard to ozone oxidation; ozone reacts very slowly with amides and saccharides in general. Dissolved organic nitrogen (DON) might give a rough estimate of the maximum possible amine concentration; it has been measured in the range of 0.07-35 µM for low productivity lakes (12). Unfortunately, DON is typically obtained by subtraction of nitrate from total dissolved nitrogen (TDN) resulting in rather poor estimates given the innocuous presence of comparatively high concentrations of nitrate in surface water. Moieties of importance to ozone, such as olefins, are not reported. In summary, it can be assumed that natural waters contain concentrations in the µM range of phenolic and amine moieties in the NOM. However, it is clearly not possible at this point to provide an exact description of the concentration distribution of ozone-reactive moieties in natural water matrixes. Ozone-Reactive Functional Groups. Ozone is a specific oxidant, which reacts readily with a limited number of functional groups such as olefins, amines, and activated aromatic systems (10). As discussed above, some of these moieties can be found in µM concentrations in lake NOM. Ozone reacts readily with olefins through cycloaddition (e.g., for nonsubstituted olefins: k′′ ∼105 M-1s-1). These reaction rates are nearly independent of pH and neither O2•- nor O3•is generated in the process (14). The apparent reaction rates of phenols and amines with ozone display very strong pH dependencies. Protonated species react many orders of magnitude slower than the deprotonated (phenolate) or neutral species (amine). Recent investigations have shown that pure solutions of amines and phenols generate HO• through formation of O3•- which, at neutral pH, instantaneously decays into HO• and O2. An indirect pathway leads to the generation of O2•- which reacts quickly and selectively with O3 to form O3•- and finally HO• (15-17). Tertiary amines generate ∼10% O3•- upon an electron transfer to ozone, following eq 1. However, the main mechanism leads to the formation of oxylamine and oxygen, 3058
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following eq 2 (15).
R3N + O3 f R3N•+ + O3•-
(1)
R3N + O3 f R3NO + 1O2
(2)
In contrast, the main mechanism of the reaction between ozone and secondary amines, seems to induce O2•- (eq 3), while only 20% leads to hydroxylamine and singlet oxygen (eq 4) (18).
R2NH + O3 f R2NO• + H+ + O2•-
(3)
R2NH + O3 f R2NOH + 1O2
(4)
In the case of primary amines and amino acid, the mechanisms are less well understood, but are probably akin to the secondary amine mechanism. Based on the above mechanisms, one can expect the reaction of ozone with tertiary amines to generate HO• more rapidly but in smaller yields than in the reaction with secondary amines. Oxidation of secondary amines should also consume significantly more ozone than that of tertiary amines due to the O2•- intermediate which reacts with an additional ozone molecule prior to generating HO•. A fraction of the reaction of phenol (i.e., phenolate at neutral pH) with ozone also generates HO• following an electron transfer forming O3•- with a yield of 22%, as indicated by eq 5 (17). Subsequent oxidation of its product with the generated HO• can lead to formation of O2•-, accelerating ozone decomposition.
PhO- + O3 f PhO• + O3•-
(5)
In this investigation, naturally occurring concentrations of primary, secondary, and tertiary amines, amino acids, and phenol were added to surface water and ozone decomposition, as well as HO• generation were measured starting 350 milliseconds after ozone addition.
FIGURE 2. (a) Ozone decomposition and (b) HO• exposures during ozonation of Lake Zurich water at pH 8 and 1.5 mgO3/L (31 µM). With addition of 5 µM glycine, 2.5 µM phenol, 10 µM trimethylamine, 10 µM glucosamine, and 10 µM dimethylamine. Percentage values associated with solid gray symbols represent the calculated degree of oxidation of compounds.
Materials and Methods
Results and Discussion
Reagents. All reagents were of analytical grade. All solutions were made with milliQ water with a resistivity > 18 mΩ cm. Ozone stock solution was made by sparging an ozone/oxygen gas mixture through ice-cooled water. The solution was diluted to reach O3 ) 500 µM, acidified with 1 mM H2SO4, and kept at 1 °C in an ice bath for the duration of the experiments. Indigo reagent contained 312 µM indigo and 10 mL/L H3PO4 (85%) for all experiments. Fulvic and humic acid isolates were obtained from the International Humic Substance Society (IHSS, Nordic aquatic FA and HA).
Effect of Phenolic and Amino Compounds on HO• Generation and Ozone Decomposition. In Figure 2a, ozone decomposition in Lake Zurich water at pH 8 (solid triangles) shows that 10% of ozone is consumed prior to 350 milliseconds and 40% prior to 20 s. This is in agreement with earlier research (7). In a standard ozone batch experiment with Lake Zurich water at pH 8, the so-called “instantaneous ozone demand” (IOD) would, therefore, be calculated as 40% of the ozone dose. The curved line displayed by the data shows that unlike what is typically observed during the second phase, ozone kinetics in the initial phase is not of apparent first order. Figure 2b shows the increase of HO• exposure as a function of O3 exposure (∫HO•dt versus ∫O3dt); the slopes of the curves represent Rct.
Natural Waters. Lake Zurich water (pH 7.8, alkalinity 2.4 mM, DOC 1.4 mg/L) was collected from the raw water intake of Zurich drinking water treatment plant, 30 m below the lake’s surface. The water was filtered at 0.45 µm and kept at 4 °C. Waters were buffered with 0.5 mM borate, which increased the pH to 8.6. After mixing with ozone solution (1:10), the final pH was 7.95 ( 0.05. Para-chlorobenzoic acid (pCBA, 1 µM) was added to all waters to serve as a HO• probe. Methods. HO• exposure was back-calculated based on the extent of oxidation of pCBA, analyzed with HPLC (8). Ozone was measured online with a Varian Cary 100, either directly at 258 nm ( ) 3000 M-1cm-1) or with indigo at 600 nm ( ) 20 000 M-1cm-1) (19). When HO• induced reactions needed to be excluded, 50 mM tert-butyl alcohol (tBA) was added to the solution to serve as an HO• scavenger (>99% HO•-scavenging as tested with pCBA). The rapid measurement of ozone decomposition was performed with a continuous quench flow system (CQFS). The system which has been described and characterized previously (6, 7) rapidly mixes a stream of aqueous ozone with one of natural water in a contact loop and quenches residual ozone with the indigo reagent. CQFS allows a first measurement 115 milliseconds after ozone addition. For this project, the system was slightly modified with the indigo double-syringe pump being replaced by a large volume (0.266 L) single-syringe pump (ISCO260D) which reduced noise induced by syringe’s switchover. Measurements with CQFS display the following statistics: 90% confidence intervals of ozone concentration measurements in ozone decomposition kinetic experiments are, on average, 11% off the mean values, and pCBA displays 90% confidence intervals on average 5% off the mean values (7).
Phenol addition to Lake Zurich water (2.5 µM, crosses in Figure 2) increases the ozone decomposition rate prior to the first measurement at 350 milliseconds. Accordingly, given kPh-O3 ) 1.8 × 107 M-1s-1 at pH 8 (20), phenol must be completely oxidized well before the first measurement at 350 milliseconds (i.e., 99% at 8 milliseconds). Following this rapid reaction, however, the rate of ozone decomposition is similar to that in Lake Zurich water. This indicates that 2 orders of magnitude of time after phenol oxidation is completed, the radical chain reaction is not substantially accelerated. The generation of HO• in Figure 2b (crosses) concurs with this description; a very high generation of HO• occurs prior to 350 milliseconds followed by a rather sluggish one, similar to that of Lake Zurich water. Trimethylamine (open triangles)sthe most reactive amine investigated here (kR3N-O3 ) 5 × 104 M-1s-1 at pH 8 (20))shas reacted substantially prior to 350 milliseconds (48%) and almost completely at 3 s (95%). This can be observed in the ozone decomposition profile (Figure 2a) for which the rate decreases considerably around 3 s and becomes very similar to that of Lake Zurich water afterward. Similarly to phenols, this indicates that trimethylamine does not increase the radical chain induced ozone decomposition once its own primary oxidation is completed. Moreover, Figure 2b indicates that trimethylamine increases the HO•-exposure during its reaction with ozone but not substantially once the oxidation is completed (curve becomes parallel to Lake Zurich water). VOL. 40, NO. 9, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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Glycine (solid diamonds) and glucosamine (open circles) are considerably slower in their reaction with ozone. Only 70% of glycine (kgly-O3 ) 1600 M-1s-1 at pH 8; ref 20) is oxidized prior to the last measurement point at 20 s. The rate constant for glucosamine and ozone is not known, but as a primary amine, it is likely to be on the order of ∼103 M-1s-1 (e.g., kbutylNH2-O3 ) 340 M-1s-1). Ozone decomposition profiles in the glycine and glucosamine experiments reflect these lower reactivities. Nevertheless, the extent of ozone decomposition after 20 s of reaction with glucosamine suggests that O2•- is generated during the reaction. The induced HO•exposure also increases steadily compared to Lake Zurich water during the time window of their oxidations by ozone. Dimethylamine (solid square) (kR2NH-O3 ) 2×104 M-1s-1 at pH 8; ref 20) is slightly slower to react with ozone than trimethylamine and is, therefore, oxidized principally during the time window investigated here (25% has reacted prior to 350 milliseconds and 99% at 20 s). Dimethylamine induces a very strong increase in HO•-exposure. In comparison to Lake Zurich water (26 × 10-11 vs 5 × 10-11 Ms) the enhancement is so large that the oxidation by HO• of an ozone-refractive compound such as atrazine (katrazine-O3 ) 6 M-1s-1, katrazine-HO• ) 3 × 109 M-1s-1 ref 10) would increase from 15 to 55% in the first 20 s if 10 µM (0.24 mgC/L) dimethylamine were added to Lake Zurich water. For the same experiment, Rct () ∫HO•dt/∫O3dt) was constant at ∼10-6 during the first 20 s, which is large even for standard O3/ H2O2 AOP, and 2 orders of magnitude larger than for natural water ozonation (7). The larger HO• exposures obtained during ozonation of dimethylamine concur with earlier mechanistic investigations in synthetic water indicating an 80% O2•- yield (18). Ozone decomposition is initially more rapid with trimethylamine but, finally, more extensive with dimethylamine. This difference is due to the additional reaction of ozone with O2•- generated upon ozonation of dimethylamine but not generated upon ozonation of trimethylamine. Sorbic acid (10µM, data not shown) (ksorbic-O3 ) 9.6×105 M-1s-1 at pH 8; ref 4) was also added to Lake Zurich water to model a compound that should not impact HO• generation. Sorbic acid should react with ozone mainly through a cycloaddition to its double bonds (no O3•- or O2•- generated) (14). As predicted, data showed similar HO• generation after addition of 10 µM sorbic acid to Lake Zurich water as in unmodified Lake Zurich water (data not shown). Spiked concentrations of some compounds were varied to confirm by trends the observations made above. Figure 3 shows the effect on HO• exposure of 5 and 10 µM addition of trimethylamine, glycine, and dimethylamine to Lake Zurich water. There is a clear correlation between the compounds concentrations and HO• exposures (at same ozone exposures ∼1.2 × 10-4 Ms). Dimethylamine displays the highest HO• exposure of all amines tested. However, the above values cannot be used to directly compare the efficiency of one molecule to generate HO• versus another, because for the same ozone exposure, compounds will have reacted to various degrees depending on the magnitude of their rate constants. Spiked concentrations of phenol are also directly related to HO• exposures generated. For the investigated ozone exposure, the addition of 5 µM phenol increases HO• exposure 385% when compared to unmodified Lake Zurich water. Effect of Humic and Fulvic Acid on Ozone Decomposition and HO• Generation. As clearly demonstrated by the above data, certain compounds and hence specific moieties in the NOM not only display high reactivity with ozone but also generate high yield of HO• upon their oxidation. It is of interest to compare the profile of HO• generated by these compounds to those generated by fractions of NOM. 3060
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FIGURE 3. HO• exposures at ∫O3dt ) ∼1.2 × 10-4 Ms following ozonation after standard addition of various compounds to Lake Zurich water at pH 8 and 1.5 mgO3/L (31 µM). Addition of 0 µM, 5 µM, and 10 µM trimethylamine, glycine, dimethylamine, and 0 µM, 2.5 µM, 5 µM phenol. In Figure 4, the addition of 5 µM (60 µgC/L) fulvic acid (solid circles) and humic acid (open circles) to Lake Zurich water shows the importance of those NOM fractions on initial HO• generation. Although their ozone decomposition profile cannot be differentiated and are only marginally faster than ozone decomposition in unmodified Lake Zurich water, the initial increase in HO• exposures compared to Lake Zurich water is substantial. When compared to the simple model compounds discussed above (represented by trend lines in Figure 4b), it seems that the main effect of humic and fulvic acid on HO• generation takes place prior to 350 milliseconds. This demonstrates that the critical moieties in humic and fulvic acids inducing HO• upon ozonation have rate constants > ∼ 50′000 M-1s-1. Given the known high degree of aromaticity of fulvic and humic acid and the rapid drop in absorption at 285 nm, during the first 350 milliseconds following ozone addition (6), those moieties can be hypothesized to be phenolic or/and other activated aromatics systems. Effect of Generated HO• on Ozone Decomposition. During an earlier investigation of ozone decomposition in wastewater, the addition of an HO• scavenger did not modify ozone decomposition kinetics substantially, suggesting that the radical chain induced ozone decomposition may not have a strong influence on the initial phase (6). In Figure 5, an HO• scavenger (50 mM tert-butyl alcohol, tBA) was added to unmodified Lake Zurich water. The HO•-probe, pCBA, did not decrease during ozonation which confirmed that all HO• reacted with tBA (data not shown). Similarly to what has been observed earlier (6), addition of a scavenger did not stabilize ozone during the first 20 s (compare open vs solid diamonds). In water spiked with 2.5 and 5 µM phenol, however, an important reduction of the initial ozone decomposition (
