21,854 (1975). (4) Branson, D. R., Blau, G. E., Alexander, H. C., Neely, W. B., Trans. Am. Fish. SOC.,104,785 (1975). (5) Hamelink, J. L., Waybrant, R. C., Trans. Am. Fish. SOC.,105,124 (1976). -, (6) Clayton, J. R., Jr., Pavlou, S. P., Breitner, N. F., Enuiron. Sci. Technol., 11,676 (1977). (7) US. Geological Survey, “Water Resources Data for Indiana”, ReDort IN-74-1. 1974. (8) US. Geological Survey, “Water Resources Data for Indiana”, Report IN-75-1, 1975. (9) U S . Geological Survev. “Water Resources Data for Indiana”. Report IN-76-1, 1976. ” (10) Curry, K. D., Spacie, A., Indiana Acad. Sci., in press. (11) Parka, S. J., Worth, H. M., Proc. South. Weed Conf., 18, 469 (1965). (12) Probst, G. W., Tepe, J. B., “Degradation of Herbicides”, Kearney, P. C., Kaufman, D. D., Eds., Marcel Dekker, New York, 1969, pp 255-82. (13) Probst, G. W., Golab, T.,Herberg, R. J., Holzer, F. J., Parka, S. J., Van der Schans, C., Tepe, J. B., J Agric. Food Chem., 15,592 (1967). (14) Parr, J. F., Smith, S., Soil Sci., 115,55 (1973). (15) Golab, T., Amundson, M. E., 3rd International Congress of Pesticide Chemistry, Helsinki, Finland, July 3-9, 1974. (16) Sheets, T. J., Uradley, J. R., Jr., Jackson, M. D., Report No. 60, North Carolina State University, April, 1972. (17) Crosby, D. G., Leitis, E., Bull. Enuiron. Contam. Toxicol., 10, 237 (1973). (18) Willis, G. H., Wander, R. C., Southwick, L. M., J. Environ. Qual., 3,262 (1974). (19) Lawrence, L. M., Weed SOC.Am. Abstr., 89 (1966). \ - -
(20) Worth, H. M., 6th Inter-American Conference on Toxicology
and Occupational Health, Coral Gables, Fla., Aug 26-29,1968.
(21) Macek, K. J., Rodgers, C. R., Stalling,D. L., Korn, S., Trans. Am. Fish. SOC.,99,689 (1970). (22) Korn, S., Macedo, D., J . Fish. Res. Board Can., 30, 1880
(1973). (23) Data provided by T. Golab, Eli Lilly and Co., Greenfield Laboratories, Greenfield, Ind. (24) Data provided by G. Herr, Eli Lilly and Co., Tippecanoe Laboratories, Lafayette, Ind. (25) Roberts, J. R., deFrietas, A. S. W., Gidney, M. A. J., J. Fish. Res. Board Can., 34,89 (1977). (26) Spacie, A., Ph.D. Thesis, Purdue University, Lafayette, Ind., 1975. (27) Neely, W. B., Branson, D. R., Blau, G. E., Enuiron. Sci. Technol., 8, 1113 (1974). (28) Lu. P. Y.. Ph.D. Thesis. Universitv of Illinois. Urbana-Champaign; 1974.’ (29) Data provided by A. J. Leo, Pomona College, Claremont, Calif. (30) Macek, K. J., Lindberg, M. A., Sauter, S., Buxton, K. S., Costa, P. A,, “Toxicity of Four Pesticides to Water Fleas and Fathead Minnows”, U.S. Environmental Protection Agency, EPA-6001376-099, 1976. (31) Norstrom, R. J., McKinnon, A. E., deFreitas, A. S. W., J. Fish. Res. Board Can., 33,248 (1976). Received for review July 13, 1978. Accepted January 8, 1979. The study was supported by a grant from Eli Lilly and Co. This is Journal Paper No. 7222 from the Agricultural Experiment Station, Purdue University.
Smog Chamber Study of the Correlation of Hydroxyl Radical Rate Constants with Ozone Formation Arthur M. Winer”, Karen R. Darnall, Roger Atkinson, and James N. Pitts, Jr. Statewide Air Pollution Research Center, University of California, Riverside, Calif. 92521 H T h e continued need for more rational and practical as-
sessments of hydrocarbon reactivity for use in formulating control strategies has led to suggestions that OH radical reactivity be used as a criterion for predicting oxidant formation. In order for such a reactivity scale t o have maximum utility, correlations between the rates of reaction of hydroxyl radicals with organics and rates and levels of ozone formation must be established. T o this end, 9-h irradiations of NO,-hydrocarbon-air mixtures consisting of a n alkane, an alkene, and a n aromatic were carried out in a 6400-L all-glass smog chamber. T h e “standard” mixture consisted of n-pentane, rn-xylene, and trans-2-butene, since the rate constants were known both for their reactions, and that of at least one isomer, with the OH radical. In one case the respective isomers were substituted sequentially at the same initial concentration and the effect of differences in O H rate constants on 0 3 production for constant carbon number was examined. In a second set of runs, the initial concentrations of organics were changed so as t o maintain constant the O H reactivity (Le., [HC] X koH). T o a good approximation the rate of formation and yields of both 0 3 and PAN in the earlier stages (56-9 h) of irradiation depended primarily on the OH radical reactivity and to a much lesser extent on the amount of carbon present.
With mobile source emission control programs how well established in the U S . ( I ) , attention is again focusing heavily on reducing stationary source emissions of organics (2) as a means of approaching or achieving the federal air quality 822
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standard for oxidant (ozone). For more than a decade the concept that organics may be differentiated on the basis of their potential for producing photochemical oxidant has received general acceptance as a cost-effective basis for developing oxidant control strategies. In the past, hydrocarbon reactivity classifications have been based on smog chamber measurements of secondary smog manifestations (e.g., rates of hydrocarbon consumption, NO to NO2 conversion, ozone formation, etc.) during irradiations of hydrocarbon-oxides of nitrogen mixtures. Not surprisingly, significant differences have occurred among studies of this type conducted in various laboratories. T h e available data from such studies have been summarized and discussed in detail previously (3-5). Recent efforts a t the state (6) and federal ( 4 , 5 ) levels in the area of hydrocarbon reactivity have been t o develop revised reactivity scales which group organics by classes reflecting their relative reactivity. Such scales then provide a basis for the substitution of less reactive organics than those presently employed in industrial processes which lead t o emissions into the atmosphere. Other workers (7) have attempted to evaluate whether reductions in hydrocarbon reactivity could be as effective in limiting future ozone maxima as proportional across-the-board reductions in NMHC concentrations. In view of the difficulties encountered in formulating hydrocarbon reactivity scales based on this type of smog chamber data, we have proposed a supplementary approach which does not suffer from such difficulties (8-10). Specifically, we formulated a reactivity classification for organics based on the rates of their reaction with the hydroxyl radical (OH). The 0013-936X/79/0913-0822$01 .OO/O@ 1979 American Chemical Society
basic premise for such a reactivity scale is t h a t the dominant loss process for alkanes and aromatics (and to a lesser extent for alkenes) in the polluted troposphere is via reaction with the OH radical. However, as we have pointed out (8-11 ) the applicability of such an OH radical reactivity scale depends upon quantitatively linking OH reaction rates for organics to their oxidant-forming potential. We have carried out an investigation concerning the relationship between rates of reaction with OH and rates and amounts of ozone formed using a model system involving combinations of nine organic compounds of atmospheric relevance. Thus we report here results from a matrix of 9-h, NOx-hydrocarbon-air smog chamber irradiations employing this model system, a NO,-n-pentane-m-xylene-trans-2butene-air mixture being chosen as the “standard” system. In this system organics were chosen such that rate constants for both their reactions, and that of at least one isomer, with the hydroxyl radical are known. Two types of experiments were conducted: one in which the respective isomers were substituted sequentially at the same initial concentration, and the effects of differences in OH rate on ozone production for constant carbon number were examined, and the second in which the concentrations were changed so as t o keep constant the reactivity (i.e., the product of concentration times the OH rate constant). Experimental Experiments were carried out in a 6400-L all-glass (Pyrex) chamber, with irradiation provided by two externally mounted, diametrically opposed banks of 40 Sylvania 40-W BL lamps, backed by arrays of Alzak-coated reflectors. Prior to each experiment the chamber was flushed with dry purified air ( 1 2 ) for -2 h a t a flow of -12 cubic feet per minute (cfm). T h e chamber was then flushed with humidified pure air for -1 h to achieve the desired initial relative humidity and the chamber temperature was adjusted to the desired operating temperature by means of infrared lamps. During all flushing procedures, two sonic pumps were in operation to promote release of materials from the chamber walls. The matrix air used during the flushing procedure and for the final fill generally contained @xylene: eth B, ethylbenzene. Based primarily on the data for 3 and 6 h of irradiation.
though it might be expected, a priori, that the amount of PAN formed would depend primarily on the structure of the hydrocarbons, and hence that a relationship should not necessarily exist between OH radical reactivity and PAN formation, the present data show that such a relationship does in fact exist, a t least to a first approximation. Other interesting observations from these data (Table 111) are the formation of peroxypropionyl nitrate (PPN) in the irradiations containing 824
Environmental Science & Technology
1-butene [as observed previously ( 2 5 ) ]and the formation of biacetyl from irradiations containing o-xylene (the mechanistic implications of this latter observation are discussed elsewhere (26)).In the following sections these data are discussed with respect to the two types of experiments carried out.
Experiments at Constant Concentration with Differing OH Radical Rate Constants. Most of the major effects
on the formation of O:, and PAN are observed in the constant concentration (type C) runs, where the isomers were substituted sequentially a t the same initial concentration and the effect of differences in the OH radical rate constants for a constant carbon number were examined. I t can be seen from Tables I and IV that the reactivity trends in these type C runs are generally consistent with the OH radical rate constants. Of particular interest is the observation that substitution of o-xylene, p-xylene, or ethylbenzene for m-xylene gives rise to a marked decrease in reactivity-more so than substitution of 1-butene for trans-2-butene (for which there is approximately the same relative decrease in khqiHdrocarbon).T h e reason for this effect is probably due to the fact t h a t the highly reactive alkenes such as trans- 2-butene have almost totally reacted within the first hour or two of irradiation, while the aromatics have not been depleted after 9 h of irradiation. Thus, the effect of substituting 1-butene for trans-2-butene will be most evident in the very early stages of the irradiation. Because of the lower reactivity of the alkanes, changing the alkane OH radical rate constants has less effect on the overall reactivity of the mixture. Thus, the aromatic hydrocarbons are anticipated to be the major contributor to the observed oxidant formation over the total period of the irradiations. It is also noteworthy that the o-xylene rim is more reactive than the p-xylene run even though these two xylenes have very similar OH radical rate constants (Table I). Clearly, although to a good approximation OH radical reactivity is a dominant factor in determining oxidant production, there are other factors involved. These are almost certainly related to the reaction mechanisms subsequent to the initial OH radical attack. Of interest with respect to OH radical rate constant values were results obtained from the substitution, a t a similar concentration, of isopentane for n-pentane (Table IV). From the literature data available a t the time the present study was initiated (18, 19) it appeared that n-pentane should react with t,he OH radical a factor of two faster than isopentane. However, it can be seen from Table IV that substitution of isopentane for n-pentane resulted in an almost identical time for NO:! maximum and an almost identical 0 3 time-concentration profile. This is totally consistent with our subsequent observation (23) that n-pentane and isopentane have essentially identical rate constants for reaction with the OH radical at 300 f 1 K (Table I). Since n-pentane and isopentane were thus shown to have the same reactivity toward the OH radical, neopentane was substituted for isopentane, since the OH radical rate constants for these two isomers differ by a factor of -4 (Table I). The only type C run carried out for neopentane was, in fact, a double substitution, Le., consisting of a neopentane-l-butene-m-xylene mixture. By reference to Tables I11 and IV it can be seen that this mixture was somewhat less reactive than the n-pentane-1-butene-m-xylene mixture which may be regarded as the “standard” for this particular run. Although not conclusive, this result is again consistent with ozone production being predominantly determined by OH radical reactivity. Experiments at Constant Reactivity. For these type R runs the hydrocarbon concentrations were altered so as to keep the reactivity constant for each class of compound (Le., the product of the concentration and OH radical rate constant). In these constant reactivity experiments comparable variability was observed in the O3 and PAN time-concentration profiles, and, as in the type C runs, substitution of p-xylene and ethylbenzene for m-xylene leads to lower O3 yields. This again demonstrates that factors other than the OH radical rate constant play a role in determining rates and levels of 0:j production. In the case of the two neopentane substitution runs [carried out with neopentane concentrations
applicable to the two somewhat differing literature values for , much the neopentane OH radical rate constant ( 2 0 , 2 3 ) ]the higher carbon content of the mixture may be the reason for the higher reactivity observed relative to the standard hydrocarbon system. Implications and Comparison with the Literature. The experiments carried out a t constant concentration (type C runs) show conclusively that the rates of formation of both 0 3 and PAN in the earlier stages (56-9 h) of NO,-hydrocarbon-air photooxidations correlate well with the reactivity of the hydrocarbon mixture toward the OH radical. However, it is also apparent that the maximum O3 yield is essentially independent of isomeric substitution. While O3 maxima were not attained for the runs containing p-xylene and ethylbenzene, it is likely, from inspection of the data in Tables I11 and IV, that the O3 maxima for these compounds would be very similar to those for the other type C runs if the irradiations for the p-xylene and ethylbenzene systems continued to the O3 maximum. The present experimental observation of a constant maximum ozone yield, but with a variation in the rate of 0 3 production (consistent with the OH radical reactivity), is in complete accord with the modeling study carried out by Bufalini and co-workers (27). In the constant reactivity (type R) runs, the experimental observation of a reasonably constant maximum 0 3 yield probably arises from the essentially complete consumption of the initial NO,, rather than being due to a lack of dependence on the hydrocarbon concentration (expressed in terms of carbon number). Thus, this observation is also consistent with the data of Bufalini and co-workers (27). With respect to applications to the ambient atmosphere, the reduction of 0 3 (and PAN) production rates by lowering the OH radical reactivity of the mixture will allow more dispersion and dilution of the reactant hydrocarbon-NO, mixture, which may then be expected to lead to lower observed maximum O3 yields. This differs from the constant maximum 0 3 yields observed in the present low dilution experimental system. Hence, it would appear t h a t lowering the reactivity Of the reactant mixture toward the hydroxyl radical (especially, based on results from the present study, for hydrocarbons with OH radical rate constants in the range -5 x 10-12’ cm3 molecule-1 s-1) will cm3 molecule-1 s-l to -3 x lower both the maximum 0 3 yields and the production rates of 0 3 and PAN, thus improving the air quality in the region -0-6 h downwind from the major emission sources. Finally, as we have pointed out previously ( I I ) , the ultimate usefulness of the rates of OH radical reactions with organics as a reactivity index for formulating emission control strategies for stationary sources will depend upon obtaining a greater understanding of the reaction pathways leading to ozone formation after the initial OH radical attack. Acknowledgments The authors thank G. C. Vogelaar and F. R. Burleson for carrying out the gas chromatographic analyses and W. D. Long for valuable assistance in conducting the chamber experiments. Literature Cited (1) Clean Air Act Amendments of 1977, Public Law 95-95, Aug 7, 1977. ( 2 ) California Air Resources Board Bulletin, 1978. ( 3 ) Altshuller, A. P., Hufalini, J. J., Enuiron. Sei. Techno/., 5 , 39 (1971). (4) U.S. Environmental Protection Agency, “Proceedings of the Solvent Reactivity Conference”, Research Triangle Park, N.C., EPA-65013-74-010,Nov 1974. (5) “Air Quality Criteria for Ozone and Other Photochemical Oxidants’’, Vol. I, U.S. Environmental Protection Agency, Office of Research and Development, Washington, D.C. 20460. (6) “Adoption of a System for t h e Classification of Organic ComVolume 13, Number 7, July 1979
825
pounds According to Photochemical Reactivity”, California Air Resources Board Staff, Report No. 76-3-4, Feb 19,1976. (7) Miller, D. F., Sverdrup, G. M., Jones, K., “Factors Affecting the Design of Oxidant Control Strategies”,Battelle Final Report to US. Council on Environmental Quality, March 28,1977. (8) Darnall, K. R., Lloyd, A. C., Winer, A. M., Pitts, J. N., Jr., Enuiron. Sci. Technol., 10,692 (1976). (9) Pitts, J. N., Jr., Lloyd, A. C., Winer, A. M., Darnall, K. R., Doyle, G. J., presented at the 69th Annual Meeting of the Air Pollution Control Association,Portland, Oreg., June 27-July 1,1976, Paper NO.76-31.1. (10) Pitts, J. N., Jr., Winer, A. M., Darnall, K. R., Lloyd, A. C., Doyle, G. J., “International Conference on Photochemical Oxidant Pollution and Its Control, Proceedings”, Vol. 11, B. Dimitriades, Ed., EPA-600/3-77-001b,Paper 14-2,Jan 1977, pp 687-704. (11) Pitts, J. N., Jr., Winer, A. M., Doyle, G. J., Darnall, K. R., Enuiron. Sci Technol., 12, 100 (1978). (12) Doyle, G. J., Bekowies, P. J., Winer, A. M., Pitts, J. N., Jr., Enuiron. Sei. Technol.. 11.45 (1977). (13) Holmes, J. R., O’Brien, R. J., Crabtree, J. H., Hecht, T. A., Seinfeld, J. H., Enuiron. Sei. Technol., 7, 519 (1973). (14) Winer, A. M., Peters, J. W., Smith, J. P., Pitts, J. N., Jr., Enuiron. Sci. Technol., 8, 1118 (1974). (15) Smith. R. G.. Brvan. R. J.. Feldstein. M.. Levadie. B.. Miller. F. A,, Stephens, E: R.,“h7hite,N.G., Health Lab. Sci., 7(l)’Suppl.,’87 (1970). (16) Stephens, E. R., Price, M. A,, J . Chem. Educ., 50,351 (1973). (17) Stephens, E. R., “Hydrocarbons in Polluted Air”, Summary Report, Coordinating Research Council, Project CAPA-5-68,June 1973;NTIS No. PB230 993/A5. (18) Wu, C. H., Japar, S.M., Niki, H., J . Enuiron. Sci. Health, A l l , 191 (1976).
(19) Lloyd, A. C., Darnall, K. R., Winer, A. M., Pitts, J. N., Jr., J . Phys. Chem., 80,789 (1976). (20) Greiner, N. R., J . Chem. Phys., 53,1070 (1970). (21) Atkinson, R., Pitts, J. N., Jr., J. Chem. Phys., 63, 3591 (1975). (22) Perry, R. A,, Atkinson, R., Pitts, J. N., Jr., J . Chem. Phys., 64, 5314 (1976). (23) Darnall, K. R., Atkinson, R., Pitts, J. N., Jr., J. Phys. Chem., 82, 1581 (1978). (24) Atkinson, R., Darnall, K. R., Pitts, J. N., Jr.,J . Phys. Chem., 82, 2759 (1978). (25) Heuss, J. M., Glasson, W. A., Environ. Sci. Technol., 2, 1109 ( 1968). (26) Darnall, K. R., Atkinson, R., Pitts, J. N., Jr., J . Phys. Chern., submitted for publication. (27) Bufalini, J. J., Walter, T. A., Bufalini, M. M., Enuiron. Sci. Technol., 10,908 (1976). Receiued for review August 28, 1978. Accepted January 22,1979. The authors gratefully acknowledge the financial support of California Air Resources Board Contract No. A6-172-30.
Supplementary Material Available: Table I I I containing concentration-time profiles for reactants and products i n the runs carried out in this study will appear follouing these pages in the microfilm edition of this volume of the journal (38 pages). Photocopies of the supplementary material from this paper only or microfiche (105 X 148 mm,24X reduction,.negatives) containing all of the supplementary material for the papers in this issue may be obtained from the Business Operations Office,Books and Journals Division, American Chemical Society, 1155 16th St., “V. W., Washington, D.C. 20036. Remit check or money order for $6.50 for photocopy or $3.00 for microfiche, referring to code number ES&T-79822.
Naturally Occurring Organic Phosphorus Compounds in Aquatic Plants Walter C. Weimer’ Battelle-Northwest Laboratories, 329 Building/300 Area, Richland, Wash. 99352
David E. Armstrong Water Chemistry Program, University
of
Wisconsin, Madison, Wis. 53076
T h e inositol polyphosphate esters may represent the largest distinct class of acid-resistant organic phosphorus compounds in aquatic plants. They comprise a major fraction of the residual organic phosphorus in the extracts of three macrophyte species, an aquatic angiosperm, and an alga. The inositol di- through tetraphosphate esters are present in greater concentrations than are the inositol penta- plus hexaphosphate esters. This enrichment in the lower phosphate esters in the plant extracts is quite similar t o the lower ester enrichment reported in lake sediments. T h e annual growth of algae and macrophytes in many eutrophic lakes may produce a very significant quantity of phosphorus-containing biomass that ultimately becomes a portion of the lake sediments, since these plants or some of their degradation products eventually settle t o the lake bottom. In view of the fact that up to 70% of the phosphorus in some lake sediments may be in the form of organic phosphorus (I),the organic phosphorus components of aquatic plants may have comprised a considerable fraction of the total phosphorus content of the living plants. There has, however, been limited research characterizing the chemical nature of the organic phosphorus constituents of aquatic plants. Inositol hexaphosphate (the hexaphosphate ester of hexahydroxycyclohexane), a compound of considerable significance in terrestrial systems since it is the principal phosphorus storage form in higher plants, has been isolated 826
Environmental Science & Technology
from Lernna gibba (duckweed) (2,3). Roberts and Loewus (3) have also observed inositol phosphate production by another aquatic angiosperm, Wolffiellafloridana (water meal). While the presence of free inositols in algae has been reported ( 4 , 5 ) , no free phosphate esters of inositol have been detected as algal constituents. Phosphatidylinositols (generally the monoesters) have, however, been reported to comprise 15-20% of the total algal phospholipid content (6, 7). There has been some additional research toward a characterization of dissolved organic phosphorus components of lake waters. Minear (8) has characterized a fraction of the dissolved organic phosphorus compounds in both lake water samples and laboratory algal cultures. Small quantities of soluble deoxyribonucleic acids (DNA) and DNA fragments were identified in both the laboratory cultures and the natural water samples. Herbes et al. (9) have reported that since u p to 50% of the dissolved organic phosphorus in samples from two Michigan lakes was hydrolyzable by the enzyme phytase, the presence of inositol polyphosphates in the pool of dissolved organic phosphorus was suggested. During our current research concerning the chemical characterization of a fraction of the organic phosphorus components in lake sediments ( I O ) , several species of aquatic plants were examined as potential sources of organic phosphorus constituents to these lake sediments. Some of the chemical forms of phosphorus found in these plants were compared with the forms found in the sediments of the lakes from which the plants were collected.
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