Adsorption of Dissolved Organics in Lake Water by Aluminum Oxide. Effect of Molecular Weight James A. Davis”?and Rolf Gloor Swiss Federal lnstitute for Water Resources and Water Pollution Control (EAWAG), CH-8600 Duebendorf, Switzerland
Dissolved organic compounds in a Swiss lake were fractionated into three molecular size classes by gel exclusion chromatography, and adsorption of each fraction on colloidal alumina was studied as a function of pH. Organic compounds with molecular weight (M,) greater than 1000 formed strong complexes with the alumina surface, but low molecular weight compounds were weakly adsorbed. Electrophoretic mobility measurements indicated that alumina particles suspended in the original lake water were highly negatively charged because of adsorbed organic matter. Most of the adsorbed organic compounds were in the M,range 1000 < M, < 3000. Adsorption of these compounds during the treatment of drinking water by alum coagulation may be responsible for the preferential removal of trihalomethane precursors. Adsorption may also influence the molecular-weight distribution of dissolved organic material in lakes. Introduction
Solid surfaces are the sites of important geochemical phenomena in natural aqueous systems. The formation, aggregation, and sedimentation of particulate matter and the simultaneous or subsequent adsorption of trace elements are controlled by processes which occur a t the solid/liquid interface. In sediments and interstitial waters, the rates of other surface reactions, e.g., nucleation and crystal growth, may be the rate-determining steps for early diagenetic processes. Since many of the solid phases in natural waters contain oxides, hydroxides, or aluminosilicate minerals, the study of surface chemical properties has focused mainly on reactions at the oxide/water interface (1-*3).Recent studies have noted the potential significance of adsorbed organic matter on the chemical and physical properties of colloidal particles suspended in natural waters (4-6). An adsorbed organic film can mask the chemical properties of the underlying solid and create a surface whose chemical behavior is dominated by functional groups of the adsorbed organic material ( 4 ) .Adsorption of trace elements by organically coated colloidal material may be substantially different from the adsorption behavior observed on “clean” hydrous oxide surface (6, 7). Similarly, the rates of other important surface-regulated processes, e.g., flocculation, sedimentation, and dissolution, may be influenced by adsorbed organic matter (8-12). Adsorption of natural organic compounds on oxide surfaces has also become a topic of interest to water-treatment engineers. The ubiquitous occurrence of chloroform and other orgahochlorine compounds in chlorinated water supplies is well documented (13-16) and has been attributed to the reactions of chlorine with dissolved natural organic matter (17-19). Pretreatment of natural waters with alum before chlorination has been suggested as an effective means of removing the precursors of organochlorine compounds (2023).
Despite the significance of the topic, adsorption of natural organic matter at the oxide/water interface is still poorly understood ( 5 , 2 4 ) .Although there may be several factors which regulate the binding of polar organic compounds a t the oxide t Present address: Water Resources Division, USGS-Mail Stop 65, Menlo Park, CA 94025.
0013-936X/81/0915-1223$01.25/0 @ 1981 American Chemical Society
surface, the present work will focus on the influence of molecular size and pH on the adsorption behavior of dissolved organic material of a Swiss lake. From a geochemical point of view, i t is important to know the molecular-weight distribution of adsorbed organic matter so that we may better assess its reactivity with trace elements. The study also serves as a first step in quantifying the role of adsorption in the geochemical cycle of organic carbon in lacustrine environments. For water-treatment practice, we need to determine whether molecular weight fractionation occurs during adsorption by aluminum oxide. Such a fractionation could be significant in the light of recent reports that chloroform and other organochlorine compounds are preferentially produced by particular molecular-weight fractions (25-27). Materials and M e t h o d s
Most experiments were conducted with a 65-L water sample collected at a depth of 2 m from Lake Greifensee (Switzerland) on January 8, 1980. The water was transported to the laboratory in large glass carboys, which had been previously washed with chromic acid, rinsed 3 times with double-distilled water, and finally rinsed with lake water. The lake water was filtered within 1h of collection with 0.45-ym membrane filters which had been prewashed with distilled water and 1L of lake water. Fifty liters were passed through a weak acidic cation-exchange column in the sodium form (Sephadex CM-50). The effluent from the cation-exchange column was concentrated 200-fold (to 250 mL) by vacuum rotary evaporation a t 45 “C. The dissolved organic carbon (DOC) concentrations of the filtered lake water and lake-water concentrate were 3.3 f 0.2 and 660 f 30 mg/L, respectively. Within the limits of experimental error, no organic carbon was removed by the cationexchange resin. Organic compounds in the concentrate were divided into three fractions by exclusion chromatography with Sephadex G-25 (fine, Pharmacia, Sweden) packed as a wet slurry in a 5-cm diameter column of 1-m length. The column was calibrated with glucose (completely permeated fraction) and dextran blue (completely excluded fraction). The eluent (0.01 M NaHC03 solution adjusted to pH 7.0) was delivered by an Altex 110 pump (Altex, Berkeley, CA) at a flow rate of 10 mL/min. Forty-milliliter aliquots of the concentrate were injected into the column and collected in three fractions corresponding to the following molecular-weight ranges: fraction I, greater than 4000; fraction 11,500-4000; and fraction 111,less than 500. The three size fractions were concentrated by rotary evaporation to -200 mL, acidified to pH 3 with HCl, and stored at 4 OC in the dark until used for adsorption experiments. In addition to the preparative column described above, a separate analytical column of Sephadex G-25 was used to measure the molecular-size distribution of the filtered lake water and other samples (28). The eluent (0.01 M sodium phsophate buffer solution, adjusted to pH 7.0) was delivered at a flow rate of 60 mL/h. The totally excluded and totally permeated peaks occurred a t 68 mL (V,) and 128 mL ( Vt). DOC concentration in the column effluent was monitored continuously by the on-line detector described by Gloor and Leidner (29). The eluent was kept under nitrogen to avoid Volume 15, Number 10, October 1981 1223
contamination by inorganic carbon. Sample preparation was identical with the methods discussed above for the preparative Sephadex column. Lake-water samples were passed through the cation-exchange column, concentrated 200-fold by vacuum rotary evaporation, acidified to pH 3, and stored at 4 "C in the dark. Before injection, samples were purged with nitrogen and adjusted to pH -7 with carbonate-free NaOH, and a 1-mL subsample was injected into the column with a loop injector. Other details of the preparative and analytical methods are described elsewhere (28). Adsorption experiments were conducted with a commercial product, Alon (Cabot Corp., Boston, MA), as adsorbent. According to X-ray diffraction analysis, Alon is predominantly y-Al203. The primary particles are nonporous and approximately spherical in shape, with an average diameter of 0.03 micron and a specific surface area of 120 m2/g (30). Before being used, the oxide was washed with dilute acid and base as described elsewhere (31,32). Adsorption experiments were performed in batch by the following methods: (1) Experiments with filtered lake water were begun within 2 h after collection of the sample. Aliquots (100 mL) of lake water were transferred to clean glass Erlenmeyer flasks, and the pH was adjusted to the desired value by small additions of HC1 or NaOH. y-Al2O3(100 mg) was added from a stock suspension of alumina (in distilled water), and then the flask was closed with a ground-glass stopper and shaken with mild agitation for 10 h at 20 OC. Preliminary experiments showed that 2-3 h was sufficient time to reach equilibrium. At the end of the equilibration period, the solid and liquid phases were separated by centrifugation at 10 000 rpm for 10 min, and equilbrium DOC concentrations and pH were measured. DOC was measured in triplicate with a TOC-Unor analyzer (Maihak Co., Hamburg, Germany). The minimum detectable concentration was 0.2 mg/L DOC. (2) Experiments with the lake water concentrate fractions were conducted by suspending 100 mg of alumina in 0.01 M NaCl solution, adjusting the pH to the desired experimental value, and allowing the suspension to equilibrate for 1h. Organic matter was then added from the stock solutions of fraction I, 11, or 111to yield an initial DOC concentration of 9.4 mg/L. Equilibration of the suspensions was carried out as above. All adsorption experiments were conducted in the pH range 4.5-9.5, where dissolution of alumina is less than M (33).Other details of the experimental methods are described elsewhere ( 5 ) . Electrophoretic mobility was measured at 25 "C in a cylindrical cell with the Mark I1 microelectrophoresis unit (Rank Brothers, London). A minimum of 10 particles were timed in each direction, and a mean velocity was computed from the measurements. The pH of each sample was measured immediately before being placed in the cell. Samples for the electrophoresis study were prepared by adding small volumes of stock y-Al2O3suspensions to filtered lake water samples (with pH adjusted) collected from Lake Greifensee on February 22, 1979. Adsorption of organic material from this sample on alumina was reported previously ( 5 ) . Samples were equilibrated with the lake water at 25 "C for 10 h before electrophoretic mobility was measured. A second set of mobility measurements was made with lake water exposed to ultraviolet (UV) light for 1.5 h, which reduced the DOC concentration from 3.3 to 0.5 mg/L. Pure oxygen gas was bubbled through the lake water during UV oxidation, but no other reagents or oxidants were added. The pH and bicarbonate concentration before UV oxidation were 8.0 and 224 mg/L as HC03-. After UV treatment, the values were 8.8 and 187 mg/L, due to COz stripping by the oxygen. T o restore the original inorganic composition, we slowly bubbled 5% C02 through the lake water until a pH of 8.0 was 1224
Environmental Science & Technology
attained. The UV-treated lake water was then divided into subsamples which were treated as described above before electrophoretic mobility measurements.
Results and Discussion Figure 1 illustrates the molecular-size distribution of organic compounds in the filtered lake water. The chromatogram shows the relative amount of DOC as a function of molecular weight, sugar and dextran compounds being used for calibration of the column (28). Since the detector responds directly to the carbon content of the effluent (as opposed to an indirect measurement, such as ultraviolet absorbance), the chromatogram respresents a quantitative distribution of organic carbon in the lake water. It is possible to estimate the amount of organic carbon in a particular size range; however, the estimate is limited somewhat by band broadening. For example, the organic-carbon peak of the totally excluded fraction in Figure 1 (nominally M , > 5000) is broadened to at least a value of 4000. Thus, in this sample, one can state that -18% of the DOC is present in compounds with a molecular weight greater than 4000. Several injections of the sample were made, and the chromatogram was very reproducible. Mass balances showed that all organic carbon injected to the Sephadex column was eluted. However, a slight tailing ( 4000 (fraction I) and those with M , < 500 (fraction 111).However, fraction I1 (nominally compounds with 500 < M , < 4000) still contained a significant amount of organic carbon which eluted a t volumes corresponding to compounds with M , less than 500. The chromatograms in Figure 2 do not correspond quantitatively to the original distribution shown in Figure 1,since each fraction was concentrated by a different factor after elution from the preparative column. Using the nominal
tI z 0
m LT
a 0 0
z a
IJ
I
I
I
'5000 3000 1000 600 4000, 500-4000, 4000) was decreased by 65% (see Figure 3),
8ol
90
70 n w
1-
1001
80 70
I I I I I I I ADSORPTION FROM FILTERED LAKE WATER Y-A120~ 10911, D O C 1 ~ ~ n ~ ~ 3 3 r n g("8180) ll
ADSORPTION CALCULATED FROM INDEPENDENT EXPERIMENTS WITH FRACTIONS I-E FRAYON
OF
EL
001
n
0
W
n
0
20
lo;O4
PH
Figure 3. percent DOC adsorbed as a function of pH by 1.0 g/L y-Alp03 for three size fractions isolated from Lake Greifensee water (1/8/80). Initial DOC concentration in each case was 9.4 mg/L in a 0.01 M NaCl solution.
PH
Figure 4. Percent DOC adsorbed as a function of pH for 1.0 g/L yA1203 suspended in Lake Greifensee water (circles) and a comparison with predicted values from independent experiments with the three size fractions.
Volume 15, Number 10, October 1981
1225
in fraction I1 (500 < M , < 4000) by 4076, and in fraction I11 ( M , < 500) by 20%. The actual changes in molecular-size distribution shown in Figure 5 are in reasonable agreement with these results. The results in both Figures 3 and 5 demonstrate an increase in affinity for the alumina surface as the molecular weight of the organic compounds increased. This trend was also observed within the distribution of organic compounds in fraction 11. Figure 6 shows the size distribution of compounds of fraction I1 before and after mixing with 1.0 g/L alumina a t pH 7.0. The upper chromatogram is identical with the one shown for fraction I1 in Figure 2. The lower chromatogram represents the size distribution pf compounds remaining in solution at the completion of the adsorption experiment. From Figure 3 we know that the aiea under the chromatogram should be reduced by -55% a t pH 7. Although DOC was decreased over the entire molecular-weight range, it i s clear that the percent reduction of DOC was much greater for the higher molecular weight compounds, e.g., -75% in the molecular-weight range 2000-3000.
AFTER ADSORPTION
z
0
m [L a
!.og/1
Y-AI203 DOCLAKE
u
3.3 m g l l
0 Z a
c3
LT 0
I
,5000 3000 1000 600 1000 in this sample had a very strong affinity for the alumina surface (Figure 7 ) .Ca. 85%of the DOC in compounds with M , > 5000 was removed, even a t pH 8.3. An explanation for these results will probably require a full understanding of the origin and nature of the high molecular weight fraction. In autumn the proportion of organic carbon in the completely excluded peak ( M , > 5000) increases relative to the low molecular weight compounds (R. Gloor and H. Leidner, unpublished results), as can be seen by a comparison of Figures 1and 7. The increase in high molecular weight organic carbon occurs during or after the usual bloom of plankton in the lake in early autumn. For a number of samples analyzed during 1979 and 1980, the high molecular weight organic carbon varied from ca. 10%to 30%percent of the t o t a l DOC, with the lowest values usually found in the winter. Veenstra and Schnoor (35) observed a similar pattern for seasonal variations of molecular-weight distribution in the Iowa River. The correlation of the increase in high molecular weight carbon with increase in algal productivity suggests that this material is of autochthonous origin. Most of the dissolved organic compounds released extracellularly during active growth of algae or upon autolysis a t death are decomposed
11)
,
Fl LT E R E T L A K E WAT
COMPOUNDS REMAINING IN SOLUTION AFTER ADSORPTION EXPERIMENT
I
1
>5000 3000 1000 600 '200 APPROXIMATE MOLECULAR WEIGHT (SEPHADEX G-25, dextran calibration )
Figure 6. DOC concentration as a function of molecular size in fraction II before and after mixing with a 1.0 g/L suspension of Y-Al203 in M NaHC03 solution. 1226
Environmental Science & Technology
,
>5000
I
3000
1000 600
4000) could be flocculated and removed from solution by acidification to pH 1,which suggested that this fraction may be composed of organic colloidal particles. No precipitate was visible in our high molecular weight fraction (MI > 4000), even when stored at pH 2 for several weeks. Some differences might be attributed to methods of sample preparation, since Veenstra and Schnoor (35) did not pass their river water through a cation-exchange column before concentration and acidification. Gloor et al. (28)have shown that inorganic precipitates form (with some associated organic matter) when Lake Greifensee water is concentrated ZOO-fold without removal of calcium ions. Results of the electrophoretic mobility experiment6 are shown in Figure 8. The average zeta potential of each batch suspension of particles is plotted vs the pH measured after equilibration. The upper curve shows that the alumina particles suspended in dilute NaCl solutions were positively charged at pH values less than 9, in agreement with published values of the isolectric point and surface charge of alumina in simple electrolyte solutions (32, 4 2 ) . However, when suspended in filtered lake water, the alumina particles were highly negatively charged over the entire p H range (4-8.5). Most of the change in mobility is due to adsorbed organics, because alumina particles suspended in the UV-treated lake water (0.5 mg/L DOC) exhibited significantly greater (more positive) mobilities. The mobilities of these particles were partly influenced by the remaining organic compounds and other anions capable of specific adsorption on alumina, e.g., sulfate ( 4 3 ) .The mobility results are in agreement with the results of other workers who have found negative mobilities for oxide particles suspended in filtered seawater ( 4 , 4 4 )and natural particulates in rivers and estuaries (45).The mobilities of various particles with greatly different surface and electrical properties converged to a narrow range of negative mobilities, indicating that the adsorbed organic matter dominates the subsequent surface chemistry ( 4 , 4 4 ) .Neihof and Loeb (44, 4 6 ) have presented some evidence that the organics adsorbed from seawater are of high molecular weight. Although the lake-water sample used for the mobility experiments (Figure 8) was collected on a different date (2/22/79), its adsorption behavior, reported previously ( 5 ) , was very similar to that shown in Figure 4.
50
3
LAKE WATER UNTREATED ( 3 3 m g / E D O C ) UV O X I D I Z E D ( O 5 m a / ~ w C )
.
0
4
5
6
7
8
9
40
41
PH
Figure 8. Electrophoretic mobility of y-AlpOa particles (50 mg/L) as a function of pH when suspended in 0.01 M NaCl solution (triangles), filtered Lake Greifensee water (2122179; circles), and UV-treated Lake Greifensee water (2/22/79; squares).
Geochemical Significance. In most lakes, the organiccarbon flux to the sediments is determined by the sedimentation of detrital organic material derived from autochthonous primary production by phytoplanktonic and littoral flora (411. Thus, the importance of organic adsorption is not as a contribution to the organic-carbon cycle, but rather as a process which modifies the surface properties of suspended particulate matter ( 5 ) .Adsorbed organic matter will have a significant effect on the adsorption of trace metals ( 4 , 7 , 4 7 ) and the coagulation behavior ( 4 5 ) of natural particulates. In our experiments with alumina suspended in Lake Greifensee water, we found that most of the adsorbed organic material is from fraction I1 (Figure 4).Moreover, the results of Figure 6 demonstrate that the bulk of this fraction removed was in the molecular-weight range 1000-3000. This is an important result, since it has been shown that this intermediate molecular weight fraction of natural organic matter forms the strongest complexes with transition-metal ions (48-51 ). Alumina and aluminosilicate surfaces in natural waters are probably covered with adsorbed organic matter, and this particulate material may be very effective in removing trace metals from solution by the formation of metal-organic complexes at the surface ( 7 ) .Thus, trace-metal adsorption models which consider natural organic ligands only as complexing agents in solution may be inadequate. In fact, the presence of the organic ligands at the surface may enhance the removal of trace metals from solution ( 4 , 6 ,7 ) . Although the role of adsorption in the total organic carbon cycle in lakes is probably small, adsorption could influence the molecular-weight distribution of the dissolved organic pool. A conspicuous feature of limnology is that in most lakes, including Lake Greifensee, the dissolved organic carbon concentration changes very little seasonally, even though the lake is experiencing dramatic metabolic changes throughout the year (41). Carbohydrates, proteinaceous compounds, fatty acids, and other low molecular weight compounds are labile and easily degraded by microorganisms. Thus, instantaneous measurements of the chemical mass of dissolved organic carbon, such as those in this study, are highly biased towards refractory compounds, even though the labile components may represent the major organic-carbon pathways and energy fluxes ( 4 1 ) .Since the DOC pool consists largely of refractory compounds and the concentration of DOC is relatively constant, the inputs of these compounds to the lake must be equal to their removal by water outflow, adsorption and sedimentation, and slow microbial degradation. Since we have not quantified the inputs and the outputs to the dissolved organic carbon pool in the lake, it is not possible to determine the actual importance of adsorption in influencing the molecular-weight distribution. Low molecular weight compounds (MI < 1000) are weakly adsorbed, so we can safely assume that adsorption plays a minor role in the cycling of this fraction. Although high molecular weight compounds were greatly removed by the addition of alumina, we cannot absolutely state that adsorption was the major removal mechanism, since heterocoagulation of organic colloidal material, e.g., broken cell walls, remains a possibility. For those compounds of intermediate size (1000 < M , < 3000),however, adsorption may be a significant removal mechanism which reduces the steady-state concentration of this fraction in the lake. Stabel and Steinberg ( 5 2 ) reached a similar conclusion after observing a dramatic reduction in the concentration of higher molecular weight compounds in a German lake following a large input of allochthonous inorganic particulate matter. Engineering Significance. Much of the attention on the chloriqation of drinking water has focused on the formation of chloroform and other trihalomethanes (THMs). Several authors have noted that this public and scientific concern Volume 15, Number 10, October 1981
1227
should be extended to include all organochlorine compounds formed, especially since the total organic chlorine produced is much greater than the chlorine present in THMs ( 1 4 , 1 5 , 19,53). Two studies have reported that high molecular weight organochlorine compounds are formed during the chlorination of surface waters (24,271, and the health risks of these compounds are unknown. Recent investigations have concluded that the yields of THMs are greater for reactions between chlorine and organic compounds of intermediate molecular weight (25, 3 5 ) . Our results show that this material (fraction 11) is removed to a greater extent than low molecular weight compounds. This is consistent with the observations that haloform potential is decreased by a greater percentage than DOC during alum coagulation (16,21).The explanation for this phenomena may be that the precursors of organochlorine compounds also contain the necessary functional groups for reaction with the alumina surface. Chlorine is presumed to react predominately with compounds which contain aromatic carboxylic groups with particular substitution patterns ( 17-19), and Kummert and Stumm ( 3 1 ) have shown that some of these compounds are likely to form strong complexes with the alumina surface. The pH dependence of adsorption in our experiments is similar to that found for organics removal by alum coagulation ( 2 0 , 2 1 , 2 3 ) .Our study focused on the maximum possible organics removal by adding excess surface area (120 m2/L) to the experimental systems. Only 50 mg/L of alumina (6 m2/L was added to the lake water for the electrophoretic mobility experiments (Figure 8),and the maximum DOC removed was -18% a t p H 5. In this case, the surface coverage of adsorbed organic matter is -12 mg of organic carbon/g of alumina. Maximum surface coverage for this alumina is presumed to be -30 mg of organic carbon/g of alumina, in the absence of competing anions (5). Thus, available surface area is probably a limiting factor for organics removal a t an alumina concentration of 50 mg/L. Another indication of the extent of surface coverage is the large negative mobilities of the alumina particles under these conditions (Figure 8). Such large negative mobilities for “clean” oxide particles would usually result in the formation of a stable colloidal suspension ( 1 ) .However, jar test experiments showed that the organically coated alumina particles coagulated rapidly a t all pH values. This may occur because the shear plane of the particle is extended farther from the oxide surface by the adsorbed organic film (W. Stumm, personal communication). Semmens and Field ( 2 0 ) noted that alum doses necessary to achieve optimum organics removal by coagulation are greater than is conventionally used for turbidity removal; doses of 100-200 mg/L alum may be needed. The higher alum doses are required because the predominant mechanism of organics removal is adsorption to an aluminuim hydroxide surface, and it is necessary to provide enough surface area for complete removal of the adsorbable compounds. (Note: Alum doses and y-A120aweight concentrations used in this study are not directly comparable because of differences in specific surface area.) Complete removal of the DOC is rarely achieved a t any alum dose; this study and others ( 5 , 2 1 , 2 3 ) have shown that a significant portion of the DOC (20-60%) cannot be removed from solution, even with very large doses of alumina or alum. This unreactive part of the DOC is composed of compounds which do not possess the functional groups necessary for the formation of strong complexes a t the alumina surface ( 5 , 4 0 ) .Other treatment alternatives will have to be considered to achieve greater removals of these compounds. Possibilities to consider are (1)ozonation before alum pretreatment, which may increase the reactivity of some compounds with the alumina surface, and (2) the use of granulated activated carbon (GAC) to achieve further organics removal 1228
Environmental Science & Technology
after alum pretreatment. Since the mechanism of adsorption on GAC involves the attraction of hydrophobic or less polar moieties to the carbon surface ( 5 4 , 5 5 ) ,alum coagulation and GAC adsorption have the potential to remove different fractions of the DOC pool. As has been noted by others ( 2 0 ) ,alum pretreatment could also reduce the cost of maintaining the GAC column operation. Future research in this area should focus on the various alternatives for organics removal prior to chlorination. The molecular weight and the functional groups present will largely determine which compounds are removed by alum coagulation and GAG adsorption. Commerical “humic substances’’ or organic compounds derived from alkaline soil or peat extracts may be poor models for the DOC of surface waters, since the molecular-weight distribution and functionality of these compounds may be very different ( 3 4 , 3 9 , 56, 57). Buffle et al. ( 3 4 ) showed that only distilled-water extracts of soils produced solutions where the molecularweight distribution of organic compounds was similar to that found in natural waters. Research should be applied directly to the DOC of real surface waters, rather than model “humic substances”, to be certain that the results are meaningful for water-treatment practice. Acknowledgment
We thank W. Stumm, K. Wuhrmann, H. Leidner, P. Baccini, P. Singer, C. O’Melia, P. Roberts, and E. M. Perdue for many helpful discussions during the course of this study and the preparation of the manuscript. The technical assistance of G. Sigg-Osman, H. Greutert, T. Fleischmann, and P. Egli is also gratefully acknowledged. ‘Literature Cited (1) Stumm, W.; Morgan, J. J. “Aquatic Chemistry”, 2nd ed.; Wiley-
Interscience; New York, 1981. (2) Schindler, P. W. In “Adsorption of Inorganic at the Solid-Liquid Interface”; Anderson, M. A., Rubin, A., Eds Ann Arbor Science Publishers: Ann Arbor, MI, in press. (3) James, R. 0.;Parks, G. A. Surf. Colloid Sei., in press. (4) Hunter, K. A. Limnol. Oceanogr. 1980,25,807. (5) Davis, J. A. In “Contaminants and Sediments”; Baker, R., Ed.; Ann Arbor Science Publishers: Ann Arbor, MI, 1980; Vol. 2, Chapter 15; pp 279-304. (6) Balistrieri, L.; Brewer, P. G.; Murray, J. W. Deep-sea Res., 1981, 28A, 101. (7) Davis, J. A.; Leckie, J. 0. Enuiron. Sci. Technol. 1978, 12, 1309. (8) Bovle. E. A.: Edmond. J. M.: Sholkovitz. E. R. Geochim. Cosmocf&. Acta’ 1977,41,1313. (9) Sholkovitz, E. R. Earth Planet. Sci. Lett. 1978. 41. 7 7 . (10) Chase, R.’R. P. Limnol. Oceanogr. 1979,24,417. ’ (11) Barcelona, M. J.; Atwood, D. K. Mar. Chem. 1978,6,99. (12) Suess, E. Geochim. Cosmochim. Acta 1973,37,2435. (13) Stevens, A. A.; Slocum, C. J.; Seeger, D. R.; Robeck, G. G. J . Am. Water Works Assoc. 1976,68,615. (14) Rook, J . J. J . Am. Water Works Assoc. 1976,68,168. (15) Sontheimer, H.; Heilker, E.; Jekel, M. R.; Nolte, H.; Vollmer, F. H. J . Am. Water Works Assoc. 1978,70,393. (16) Oliver, B. G.; Lawrence, J. J . Am. Water Works Assoc. 1979,71, 161. (17) Rook, J. J. Enuiron. Sci. Technol. 1977,11,478. (18) Norwood, D. L.; Johnson, J. D.; Christman, R. F.; Hass, J. R.; Bobenrieth, M. J. Enuiron. Sci. Technol. 1980,14, 187. (19) Larson, R. A.; Rockwell, A. L. Enuiron. Sci. Technol. 1979,13, 325. (20) Semmens, M. J.; Field, T. K. J . Am. Water Works Assoc. 1980, 72,476. (21) Babcock, D. B.; Singer, P. C. J . Am. Water Works Assoc. 1979, 71, 149. (22) Sontheimer, H. J . Am. Water Works Assoc. 1980, 72,386. (23) Kavanaugh, M. C. J . Am. Water Works Assoc. 1978,70,613. (24) Parfitt, R. L.; Fraser, A. R.; Farmer, V. C. J . Soil Sci. 1977,28, 289. (25) Schnoor, J. L.; Nitzschke, J. L.; Lucas, R. D.; Veenstra, J. N. Enuiron. Sci. Technol. 1979,13, 1134. (26) Oliver, B. G.; Visser, S. A. Water Res. 1980,14, 1137. ~
~~
(27) McCahill, M. P.; Conroy, L. E.; Maier, W. J. Enuiron. Sei. Technol. 1980, 24, 201. (28) Gloor, R.; Leidner, H.; Fleischmann, T.; Wuhrmann, K. Water Res., 1981,15,457. (29) Gloor, R.; Leidner, H. Anal. Chem. 1979,51,645. (30) Huang, C. P.; Stumm, W. Surf. Sei. 1972,32,287. (31) Kummert, R.; Stumm, W. J. Colloid Interface Sei. 1980, 75, 373. (32) Hohl, H.; Stumm, W. J . Colloid Interface Sei. 1976,55, 281. (33) May, H. M.; Helmke, P. A,; Jackson, M. L. Geochim. Cosmochim. Acta 1979,43,861. ’ (34) Buffle, J.; Deladoey, P.; Haerdi, W. Anal. Chim. Acta 1978,102, 339. (35) Veenstra, J. N.; Schnoor, J. L. J. Am. Water Works Assoc. 1980, 72, 583. (36) Reuter, J. H. “Abstract, 1977 Annual Meeting, Geological Society of America”, 1977; Vol. 9, No. 7, p 1140. (37) Parfitt, R. L.; Fraser, A. R.; Russell, J. D.; Farmer, V. C. J . Soil Sei. 1977,28,40. (38) Watson, J. R.; Posner, A. M.; Quirk, J. P. J . Soil Sei. 1973,24, 503. (39) Perdue, E. M. ACS Symp. Ser. 1979, No. 93, Chapter 5; pp 99-114. (40) Davis, J. A., submitted for publication in Geochim. Cosmochim. Acta. (41) Wetzel, R. Cr. “Limnology”; W. B. Saunders: Philadelphia, PA, 1975; pp 538-621.
(42) Wiese, G. R.; Healy, T. W. J. Colloid Interface Sei. 1975, 52, 427. (43) Davis, J. A.; Leckie, J. 0. J . Colloid Interface Sei. 1980, 74, 32. (44) Neihof, 12. A.; Loeb, G. I. Limnol. Oceanogr. 1972, 27, 7. (45) Hunter, K. A.; Liss, P. S. Nature (London) 1979,282,823. (46) Loeb, G. I.; Neihof, R. A. J . Mar. Res. 1977,35, 283. (47) Elliott, H. A.; Huang, C. P. Enuiron. Sei. Technol. 1980, 14, 87. (48) Baccini, P.; Suter, U. Schweiz. 2. Hydrol. 1979,41/42, 291. (49) Steinberg, C.; Stabel, H. Vom Wasser 1 9 7 8 , l l . (50) Sugai, S. F.; Healy, M. L. Mar. Chem. 1978,6, 291. (51) Baham, J.; Ball, N. B.; Sposito, G. J . Enuiron. Qual. 1978, 7,
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Receiued for review December 22,2980. Accepted June 15,1982
Thermal Decomposition Rate Constant of Peroxybenzoyl Nitrate in the Gas Phase Tomohiro Ohta” and lsao Mizoguchi Department of Environmental Sciences, Tokyo Metropolitan Research Laboratory of Public Health, Hyakunincho, Shinjukuku, Tokyo, 160, Japan
The temperature-dependent rate constant of peroxybenzoyl nitrate (PBzN) decay in 1atm of 0 2 Nz was determined by use of a new spectrometer, a second-derivative [JV-visible spectrometer with a 4-m optical path length. PBzN was prepared by the irradiation with black-blue fluorescence lamps (300 < h < 400 nm) of a mixture of Clz/NOz/C~H&HO/O~/N~. By adding excess nitric oxide to the reaction chamber after the photolysis, we found a first-order decay in PBzN during the dark period over the temperature range 290.3-304.7 K. The rate constant was determined to be 8.5 X 1014exp[(-25.2 f 3.0)/RT] s-l, which was fairly close to the value for peroxyacethyl nitrate (PAN).
+
Introduction It has been shown that peroxyacyl radicals add to NO2 to form peroxyacyl nitrates ( I ) . In recent years it has also been shown that peroxyalkyl radicals add to NO2 to form peroxyalkyl nitrates (2).Among the peroxy nitrates, HOzNOz ( 3 , 4 ) , CH3COO2N02 (5),ClCOOzNOz ( 6 ) , CH302NOz (7), CC1302N02 and CClzFOzNOz ( 8 , 9 ) , and t-BuOzNOa (10) were studied from kinetic and/or spectroscopic viewpoints. However, no rate constants of aromatic homologues have been studied. A kinetic study of peroxybenzoyl nitrate is necessary and interesting, not only because of its potential creation in the atmosphere and its possible role as an eye irritant ( I I ) , but also for understanding peroxy nitrate chemistry in general ( 1 2 ) . In the present paper, we report the measurement of the thermal decomposition rate constant of PBzN by use of a second-derivative UV-visible spectrometer (see, for example, ref (13). Experimental Section PBzN was prepared by the irradiation with a black-blue fluorescence lamp (300 < < 400 nm) of a mixture of Clz/ NOz/C6H&H0/02/Nz proposed first by Gay et al. (14). A 0013-936X/81/0915-1229$01,25/0
25.2-L Pyrex photochemical reaction chamber surrounded by four 40-W lamps was connected to a 7.57-L 1-m long stainless optical absorption cell of a second-derivative spectrometer (4-m optical path length, YANACO UO-2) by two pipes attached to both ends of the cell. Between the cell and the chamber, a Teflon diaphragm pump was placed to circulate the gases in the reaction chamber and the absorption cell at a flow rate of 10 L/min. The whole system led to a mercury-free greaseless gas-handling unit and a high-vacuum line. Benzaldehyde (2.5 pL) was injected with a syringe into a small glass tube attached to the evacuated chamber and the cell. Purified NO2 (Matheson) and Clz (Fujimoto Sanso) were measured in a calibrated tube and introduced into the cell by 50 torr of 0 2 and 600 torr of Nz.The initial concentrations for the photolysis were as follows: [CGH~CHO] = 18.4 ppm, [NOz] = 24.4 ppm, and [Clz] = 31.8 ppm. The mixture was prepared each time and irradiated for 11min to synthesize PBzN before each dark-reaction experiment. The initial concentrations . were determined after several trials to be sufficient to decompose most of the C6H&HO and make [NO10 = 0 and [NO210 = 5 ppm for the dark reaction. The major reaction scheme for the formation of PBzN is shown by reactions 1-4 (15).
The products of the photolysis in our system were condensed ahead of the circulation pump in a trap immersed in a dry ice-ethanol bath and were eluted by n-hexane. The liquidphase infrared spectra showed the formation of a compound
@ 1981 American Chemical Society
Volume 15, Number 10, October 1981
1229
