Envlron. Sci. Technol. 1992, 26, 1375-1382
Gordon, G. E. Environ. Sei. Technol. 1988,22, 1132. Dzubay, T. G.; Stevens, R. K.; Gordon, G. E.; Olmez, I.; Sheffield, A. E.; Courtney, W. J. Environ. Sei. Technol. 1988, 22, 46. Olmez, I.; Gordon, G. E. Science 1985,229, 966. Kowalczyk, G. S.; Gordon, G. E.; Rheingrover, S. W. Environ. Sei. Technol. 1982, 16, 79. Kowalczyk, G. S. Ph.D. Dissertation, University of Maryland, College Park, MD, 1979. Tuncel, S. G.; Olmez, I.; Parrington, J. R.; Gordon, G. E.; Stevens, R. K. Environ. Sei. Technol. 1985, 19, 529. Mroz, E. J. Ph.D. Dissertation, University of Maryland, College Park, MD, 1976. Cahill, R. A. M.S. Dissertation, University of Maryland, College Park, MD, 1974. Mizohata, A. J. Aerosol Res., Jpn. 1986, 1, 274. Failey, M. P.; Anderson, D. L.; Zoller, W. H.; Gordon, G. E.: Lindstrom. R. M. Anal. Chem. 1979.51. 2209. Germani, M. S.;Gokmen, I.; Sigleo, A. C.; Kowalczyk, G. S.; Olmez, I.; Small, A. M.; Anderson, D. L.; Failey, M. P.; Gulovali, M. P.; Choquette, C. E.; Lepel, E. A.; Gordon, G.
E.; Zoller, W. H. Anal. Chem, 1980, 52, 240. (30) Olmez, I.; Sheffield, A. E.; Gordon, G. E.; Houck, J. E.; (31)
(32) (33) (34) (35)
Pritchett, L. C.; Cooper, J. A.; Dzubay, T. G.; Bennett, R. L. J. Air Pollut. Control Assoc. 1988, 38, 1392. Schacklette, H. T.; Hamilton, J. C.; Boerngen, J. G.; Bowles, J. M. Elemental Composition of Surficial Materials in the Conterminous United States. U.S. Geol. Surv. Prof. Pap. 1977, NO.574-0. Thompson, C. M. Ph.D. Dissertation, University of Maryland, College Park, 1985. Rahn, K. A. The Chemical Composition of the Atmospheric Aerosol;University of Rhode Island, Narragansett, RI, 1976; pp 151-177. Dodd, J. A.; Ondov, J. M.; Tuncel, G.; Dzubay, T. G.; Stevens, R. G. Enuiron. Sci. Technol. 1991, 25, 890. Yao, H. C.; Yao, Y. F. Y. J. Catal. 1984,86, 254.
Received for review November 5,1991. Accepted March 5,1992. This work was in part supported by the National Science Foundation under Grant ATM-82-19020.
Wet and Dry Deposition of Chlorinated Dioxins and Furans Carolyn J. Koester and Ronald A. Hltes*
School of Publlc and Environmental Affairsand Department of Chemistry, Indiana University, Bloomington, Indiana 47405 We investigated how wet and dry deposition contribute to the observed changes in polychlorinated dibenzo-pdioxins and dibenzofurans (PCDD/F) homologue profiles between sources and sinks. Sampling methods included wet-only samplers to collect rain and inverted frisbees and flat plates to collect dry deposition. Samples were analyzed by electron capture, negative ionization gas chromatographic mass spectrometry. Rain was collected in Indianapolis, IN. The average, total wet deposition flux of PCDD/F in Indianapolis (at 15 "C) was 220 ng/(m2 yr). Overall scavenging ratios for PCDD/F ranged from 15000 to 150000. The amount of PCDD/F which was particulate scavenged by rain increased as temperature decreased. Dry deposition fluxes were measured in Bloomington, IN, and in Indianapolis. The average dry deposition flux (at 15 "C) in Bloomington was 160 ng/(m2 yr) and in Indianapolis it was 320 ng/(m2yr). Dry deposition fluxes increased as temperature decreased. An average deposition velocity of 0.2 cm/s was calculated for PCDD/F homologues. Both wet and dry deposition, which are important mechanisms for the removal of atmospheric PCDD/F, contribute to the enhancement of 8D observed in the sediments. Introduction Polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/F) are toxic compounds that are introduced into the atmosphere from a variety of combustion sources; these include copper smelters (I), steel mills (I), automobiles (2-4), and municipal and industrial waste incinerators (3, 5-7). Once in the atmosphere, PCDD/F become dispersed throughout the environment and can be found in appreciable amounts at sites far from their sources (8, 9). Clearly, because of their ubiquity and toxicity, it is important to understand how PCDD/F move through the environment. A goal of our laboratory has been to understand the movement of PCDD/F from their sources, through the atmosphere, to their sinks (10). Previously, we have observed that the homologue profiles of PCDD/F sources and sinks are different (11). Source profiles, represented by municipal and industrial waste effluents, are characterized 0013-936X/92/0926-1375$03.00/0
by an almost uniform homologue profile of PCDD/F. In contrast, sink profiles, represented by sediments, are dominated by octachlorodibenzo-p-dioxin(8D). Both degradation and deposition processes could account for this change in homologue profile. In this paper, we consider both wet and dry deposition processes, which affect airborne PCDD/F. Wet deposition of PCDD/F has been studied by Eitzer and Hites (12). They concluded that both vapor-phase and particle-bound PCDD/F are removed from the atmosphere by rain. Rain scavenging was most efficient for particlebound PCDD/F and for PCDD/F with higher levels of chlorination. These factors contributed to the high abundance of 8D seen in the sediments. Other studies, in which only 8D could be routinely detected in rain, support this conclusion (13,14). Eitzer and Hites sampled rain in only one location: Bloomington, IN. We wished to verify their results by studying the rain scavenging of PCDD/F in another location. We chose to sample in a more urban location (Indianapolis, IN) because we expected to find higher PCDD/F concentrations. On the basis of air concentrations of PCDD/F, dry deposition should be a more important removal mechanism for atmospheric PCDD/F than wet deposition (15). We wished to test this hypothesis, and we report here the first direct measurements of PCDD/F dry deposition. Experimental Section Sampling Locations. The rain sampler was located on the roof of the School of Public and Environmental Affairs Building on the campus of Indiana UniversityPurdue University at Indianapolis, IN. Indianapolis has a population of 800 000 and is a more urban location than Bloomington, IN, where previous rain samples were collected. Bloomington, IN, which has a population of 50000, was the primary site of dry deposition collection. The samples were located on the deck of the School of Public and Environmental Affairs Building on the Bloomington campus of Indiana University. Some dry deposition samples were also collected in Indianapolis at the same site as the rain samples.
0 1992 American Chemical Society
Environ. Sci. Technol., Vol. 26, No. 7, 1992 1375
Rain Collection and Extraction. The rain collection and extraction procedures have been previously described (12). Briefly, rain was collected with a wet-only sampler (Aerochem Metrics, Bushnell, FL). The sampler's 1-m2 collection surface was made of stainless steel, and we collected volumes of 2-38 L depending on the intensity of the rain event. As the rain was collected, it was pumped through a series of three, precleaned, extra-thick, 47-mmdiameter glass fiber filters (Gelman Sciences, Ann Arbor, MI) which collected particles greater than l-pm diameter. The filtered rainwater was collected in a series of precleaned glass carboys. Three, unexposed glass fiber filters that were stored at the sampling site served as method blanks. A method blank was prepared and analyzed with every set of samples. The rainwater and filters were returned to the laboratory and spiked with [13C12] 1,2,3,7,8-pentachlorodibenzofuran (1,2,3,7,8-5F;Cambridge Isotope Labs, Woburn, MA) and with [13C12]8D(Cambridge Isotope Labs, Woburn, MA). The rainwater was extracted by stirring for 24 h with 1L of CH2C12(Omnisolv, EM Science, Gibbstown, NJ). The filters were Soxhlet extracted for 24 h with 1:l acetone: CH2Clz (both Omnisolv, EM Science, Gibbstown, NJ). Both extracts were passed through 60 g of preextracted, anhydrous Na2S0, (Fisher Scientific, St. Louis, MO) to remove excess water before sample cleanup. Dry Deposition Collection and Extraction. Dry deposition was collected using both inverted frisbee and flat plate samplers. The inverted frisbee samplers were based on the design of Hall and Upton (16). Because of their aerodynamic shape, they do not cause significant disturbances to wind flow, and their depth prevents particles from bouncing out of the samplers. The 40-cm-diameter frisbees (Discovering the World, La Mirada, CA) were covered with aluminum foil shaped to the contour of the inside surface and mounted parallel to the ground on 1.2-m-high poles. The aluminum foil was rinsed with CH2C12and coated with approximately 2 pL of mineral oil (Aldrich Chemical Co., Milwaukee, WI) per cm2 surface area (17). The mineral oil was dissolved in hexane at a concentration such that 10 mL of this solution was used to coat each frisbee sampler. As the hexane evaporated, a layer of mineral oil was deposited on the surface of the foil. The samplers' surfaces were covered with oil because wetted surfaces have better collection efficiencies (17,181. However, a possible disadvantage of the oil coating is that the oil could actively scavenge vapor-phase organic compounds, and thus, the deposition of the organic compounds would be overestimated (19). Because most atmospheric PCDD/F are bound to particles and because the method blanks (see below) showed no significant PCDD/F contamination, we assumed that scavenging of vapor-phase PCDD/F by the mineral oil was negligible. Flat plate samplers were also used. Two square samplers were constructed with 0.5-m2 glass collection surfaces mounted parallel to the ground on 1.2-m-high wooden frames. Before sample collection, the glass plates were removed and washed with Micro soap (International Product Corp., Trenton, NJ), rinsed with water, and rinsed with CH2C12. These samplers were coated with mineral oil at a concentration of 0.2 pL/cm2 surface area. Dry deposition was removed by wiping the inverted frisbee samplers twice with precleaned glass wool saturated with CH2C12,rinsing the frisbees twice with 20-mL aliquots of CH2C12,and finally wiping them twice again with CHzC12-saturatedglass wool. The glass wool and CHzC12 rinses were combined and placed in a single Soxhlet ex1376
Envlron. Sci. Technol., Vol. 26, No. 7, 1992
tractor, [13C1211,2,3,7,8-5Fand [13C12]8Dwere added as internal standards, and the samples were extracted for 24 h with 300 mL of CH2C12. Deposition collected in 3-8 frisbees, representing collection areas of approximately 0.5-1 m2, was combined into a single sample. This consolidation was necessary in order to measure detectable levels of PCDD/F. Four frisbees, prepared as above, covered and stored at the sample site served as method blanks. Method blanks were prepared and analyzed with every set of samples. Dry deposition was removed from the flat plate samplers by wiping the surface with glass wool saturated with CH2C12.The glass surface was divided into quarters, and each quarter was wiped 4 times with fresh, CH2C12-saturated glass wool. The glass wool from each sampler was placed in a Soxhlet extractor, the appropriate internal standards were added, and the sample was extracted with 300 mL of CH2C12. Measurements of dry deposition fluxes collected simultaneously with the inverted frisbees and with the flat plate samplers were approximately the same. A t-test comparing PCDD/F measurements made with two inverted frisbees and with two flat plate samplers indicated that, at the 95% confidence level, there was no difference between the average flux measured by each sampler type. We also examined the ratio of the fluxes collected by the frisbee samplers to the fluxes collected by the flat plate samplers for each PCDD/F homologue for four sampling periods during which both sampler types were used. If fluxes measured by these methods were the same, this ratio would equal 1. These ratios, for the 23 homologues measured, ranged from 0.22 to 1.91. The average ratio was 0.82, and the standard deviation was 0.4. The average ratio suggests that the flat plates are somewhat more efficient samplers of dry deposition than the inverted frisbees. Of the 23 ratios, 17 were less than 1. Using a x 2 test, we determined that, at the 95% confidence level (but not at the 99% level), it was not probable that 17 of the 23 ratios would be less than 1by random chance. This gives some indication, but not overwhelming evidence, that the flat plates are more efficient collectors than the inverted frisbees. For this study, given the weak statistical indications, we have assumed that the collection efficiencies of both the flat plate and inverted frisbee samplers are the same. Sample Analysis. All sample extracts were solvent exchanged to hexane and subjected to silica gel and alumina column chromatography as described elsewhere (8). Analyses were performed by gas chromatographic mass spectrometry (GC/MS) on a Hewlett-Packard 5985B GC/MS system equipped with an RTE-A data system. A 30-m X 0.25-mm, DB-5 capillary GC column (J&W Scientific, Rancho Cordova, CA), with helium (ultrapure grade, Air Products, Allentown, PA) as the carrier gas, was used to separate PCDD/F. The GC column temperature was programmed as follows: splitless injection at 60 "C, isothermal for 2 min, at 30 "C/min temperature ramp to 210 "C, followed by a 2 "C/min ramp to 280 "C. The injection port and the transfer line between the GC and the MS were held at 285 OC. The mass spectrometer was operated in the electron capture, negative ionization mode (ECNI) using methane (ultrapure grade, Air Products, Allentown, PA) at a pressure of 0.43 Torr (measured while the GC oven temperature was 60 "C) as the reagent gas. The ion source temperature was 150 "C. PCDD/F were detected by selected ion monitoring of two ions of the most intense chlorine isotope cluster of each PCDD/F homologue. A compound was classified as
+2
c5
100,
particle, re0.70
A
---c
80
v
v
50
6
1
........---.o. . . _ _ . _ _ . _ _ _ _ _ ~ _ ~ _ _ ~ .................
8
20
0
101
-
n
s
dissolved
100
0
y.
i-
e
8..
I
I
e
2o
e
k
0
10 20 30 Volume of Rain (Ll
n -
40
C
Results and Discussion Wet Deposition of PCDD/F by Rain. PCDD/F were measured in 15 rain samples, most representing a single rain event, collected in Indianapolis from December 1990 to June 1991. The volumes of rain collected and the temperatures during the rain events were recorded. Peakspecific PCDD/F concentrations for the dissolved and particle phases were measured and are reported elsewhere (20). Because the relative concentrations of PCDD/F isomers within a homologue are similar from sample to sample and because the relative ratios of the amounts of various PCDD/F congeners to the total amount of PCDD/F present in each sample are the same, we have chosen to quantitate PCDD/F as homologues. Total PCDD/F (the sum of all tetra- through octachlorinated congeners) in the rain's dissolved phase ranged from 8 to about 400 pg/L; total PCDD/F bound to particles range from 60 to 390 pg/L. These concentrations are comparable to those measured in Bloomington rain (12). Concentrations of particle-bound species scavenged by rain are dependent on rainfall amount (21, 22). Most particles containing pollutants are washed out of the atmosphere at the beginning of the rain event. Thus, smaller precipitation amounts are expected to have higher particle-bound concentrations of PCDD/F. We also observed
4F 5F 6F 7F 8F 4D 5D 6D 7D 8D
100 I
Flgure 1. Total PCDD/F concentration (pg/L) versus total volume (L) of rain for particulate-boundand dissolved-phase PCDD/F in 13 rain samples collected In Indianapolis, IN. Two dissolved data are missing due to a laboratory accident.
PCDD/F only if it fell within the correct retention time window and if the ratio of the responses of its quantitation ions to the responses of its confirmation ions were as predicted by isotopic abundances. Quantitation was based on comparison of the combined peak areas of the quantitation ions to the peak areas detected for a known amount of the internal standard, considering the daily calculated response factors. Response factors were determined by analyzing a PCDD/F standard which contained 23 congeners representing every level of chlorination (12). Average response factors for the tetradioxins/furans (4D/F), pentadioxins/furans (5D/F), and hexadioxins/furans (6D/F) were calculated relative to [l3CI2]1,2,3,7,8-5F;average response factors for heptadioxins/furans (7D/F) and octadioxins/furans (8D/F) were calculated relative to [l3CI2]8D.Ten response factors, one for each PCDD/F homologue, were determined. Typically, PCDD/F were quantified as homologues; all PCDD/F congeners of the same level of chlorination were assumed to have the same mass spectral response. This is not always true; for example, 2,3,7,8-4D has poor ECNI response. However, ECNI shows excellent sensitivity for most other PCDD/F, and we accepted this limitation.
t
'-
!?
i
8o 60
I
%dissolved particle
40 20 n "
4F 5F 6F 7F 8F 4D 5D 6D 7D 8D
Flgure 2. Average Indlanapolis (top) and Bloomington (bottom) raln profiles. PCDD/F are quantitated as homologues, and concentratlons are in picograms per liter of raln. Bloomingtonrain data are adapted from ref 23. Particulate-bound PCDD/F concentrations are represented by black bars; dissolved concentrations are represented by light bars. Numbers indicate level of chlorination; F indicates PCDF; D indicates PCDD.
this behavior for PCDD/F. Figure 1 shows a plot of total PCDD/F concentration versus volume of rain collected. The dissolved-phase and particle-bound PCDD/F concentrations have been plotted separately. The data have been fit to an exponential curve CPCDD/F = a ex~(bV) (1) where CPcDDF is the total PCDD/F concentration in rain, V is the voJ!ume of rainfall collected, and a and b are constants. Regression analysis of these data show that particle-bound PCDD/F concentrations are inversely correlated with rainfall amounts (r = 0.70, significant at the 95% confidence level; a = 250, b = -0.038) but that dissolved PCDD/F concentrations are not (r = 0.16). This is consistent with the expected differences between particulate and vapor scavenging processes. The overall average concentration (C,) of PCDD/F in rain was determined by
c, = cc,v,/cv,
(2)
where Ciand Vi are concentrations and volumes, respectively, of each rain event. This average concentration represents the ratio of the total mass of PCDD/F removed by rain to the total volume of rain collected during the course of our study. The average PCDD/F profile for Indianapolis rain, representing 15 rain events with an average temperature of 19 "C, is shown in Figure 2 (top). The total, dissolved-phase, and particle-bound PCDD/F average concentrations were 160, 30, and 130 pg/L, respectively. Figure 2 also compares the average PCDD/F profiles of Indianapolis and Bloomington rain. Bloomington data (see Figure 2, bottom), representing an average ambient temperature of 22 "C, were adapted from Eitzer (23) and were averaged by the same method as the IndiEnviron. Sci. Technol., Vol. 26, No. 7, 1992
1377
Table I. log 'cy," and log W,* for PCDD/F Homologues in Indianapolis (Indy) and Bloomington (Blm) RainC
PCDD/F
4F 5F 6F 7F
8F 5D 6D 7D 8D
w,
1% W" Indy
log W" Blm
1% Indy
log w p Blm
4.5 4.3 4.1 4.5 4.7 4.3 4.4 5.3 5.8
4.2 4.0 3.9 4.6 5.3 3.9 3.9 5.1 6.5
4.6 4.2 4.2 4.5 4.6 4.6 4.4 4.9 5.0
4.8 4.4 4.2 4.5 4.2 4.2 4.1 4.8 4.9
1378
Environ. Sci. Technoi., Vol. 26, No. 7, 1992
% on particles
0.2 0.4
33 000 18 000 15 000 32 000 41 000 30 000 26 000 91 000 150 000
24 35 74 79 87 67 69 78 60
0.8 0.9 0.5 0.7 0.9
0.9
5,000LL
We
0.7
a Vapor scavenging ratio. *Particle scavenging ratio. Calculated from data in Figure 2. scavenging ratio.
anapolis data. The two profiles are similar. The total, dissolved-phase, and particle-bound PCDD/F average concentrations were 150,40, and 110 pg/L, respectively. The scavenging ratio (W) is the ratio of the PCDD/F concentration in rain to the concentration in the atmosphere, and it measures the effectiveness of rain at removing PCDD/F (24). Ideally, air and rain concentrations should be measured simultaneously. However, because the sampling times required to measure PCDD/F in air are 2-3 days, this approach is not feasible. Thus, average PCDD/F air concentrations for Indianapolis (20) were used to estimate average scavenging ratios. The geometric average of the total (vapor plus particle-bound) PCDD/F concentration of each homologue in six air samples was determined (20). Vapor-phase and particle-bound PCDD/F concentrations were then calculated at the average rain temperature (19 "C) using equations describing the vapor-particle partitioning of PCDD/F (23, 25, 26). It was assumed that total, atmospheric PCDD/F concentrations did not change as a function of temperature. This conclusion is supported by Eitzer and Hites, who saw no seasonal variations in PCDD/F concentrations during a 3-year study of PCDD/F air concentrations in Bloomington (15). Using the measured, average Indianapolis rain concentrations and calculated average air concentrations at the average rain temperature, vapor scavenging ratios ( W,, which is the PCDD/F concentration in the dissolved phase of rain divided by the atmospheric vapor-phase concentration) and particulate scavenging ratios (Wprwhich is the particle-bound PCDD/F concentration in ram divided by the particle-bound PCDD/F concentration in air) were calculated. The values reported for scavenging ratios should be considered as estimates. W , and Wp may be highly variable between rain events (27) and are dependent on the atmospheric particle size distribution (which changes during a rainstorm), transient emissions from sources, temperature, cloud base height, and rainfall intensity (22). Our scavenging ratios are summarized by homologue in Table I, columns 2 and 4. Bloomington data (23) is also provided for comparison in columns 3 and 5. log W, for Indianapolis ranges from 4.1 to 5.8; log W ranges from 4.2 to 5.0. Overall, scavenging ratios measured in Indianapolis are comparable to those measured in Bloomington. When scavenging ratios in Indianapolis were regressed against those in Bloomington, the correlation coefficients ( r ) were 0.93 and 0.69 for W, and W,, respectively; both correlations were significant at the 95% confidence level. These scavenging ratios are also comparable to scavenging ratios measured for other semivolatile organic compounds (28,29). Because our scavenging ratios are lower than 106,below-cloud scavenging processes or scavenging caused by cold cloud processes such as ice
9d
Fraction bound to particles.
2m 3,000
-
Overall (total)
dissolved, r-0.42
A
+
4,000-
n
e
particle, r=0.7 1
'.__ '.__
...A....
'...
n 3 2,000 C
z
1,000 n
-5
10
15
20
25
30
Temperature (C) Figure 3. Total amount of PCDD/F removed in the dissolved phase and bound to particles in each rain event versus air temperature (in "C)during that event. Samples were collected in Indianapolis.
crystal formation and vapor accretion may be occurring (22)*
The relative contributions of vapor and particulate scavenging to overall rain scavenging can be determined by using the following relationship (12, 24) = wpO + W,(1 - 0 ) (3)
w
where W is the overall scavenging ratio and 0 is the fraction of PCDD/F bound to particles in the atmosphere. Columns 6 and 8 of Table I show 0 and the percentage of PCDD/F scavenged as particles ( WpO divided by W). On average, particulate scavenging removed 64% of the atmospheric PCDD/F in Indianapolis. For comparison, 68% of atmospheric PCDD/F were removed by particulate scavenging in Bloomington. Total rain scavenging efficiencies (see Table I, column 7) varied from 15000 to 150OOO; 7D and 8D were scavenged most efficiently. This is consistent with the sedimentary record where 8D is the predominant congener. Figure 3 shows a plot of the total mass of dissolved-phase and particle-bound PCDD/F removed by each individual Indianapolis rain event versus the ambient air temperature during the event. We observed that the total amount of PCDD/F removed by the dissolved phase in rain was not significantly correlated with temperature (r = 0.42). This lack of correlation does not imply that vapor-phase scavenging is not affected by temperature; instead, it may suggest that vapor scavenging has a more complex temperature dependence. Also, because vapor scavenging Is an equilibrium process, the total amount of vapor-phase PCDD/F removed by rain also depends on the rainfall amount. In contrast, the total amount of particle-bound PCDD/F removed by a rainfall event was inversely related to the
ambient air temperature (r = 0.71, significant at the 95% confidence level; slope = -130, intercept = 4200); see Figure 3. This behavior was expected because the amount of PCDD/F adsorbed to particles is determined by the ambient temperature (23),and if equal amounts of particles are removed by every rain event, the amount of PCDD/F removed by rain should vary with temperature. We assumed that each rain event removed an equal amount of atmospheric particles because we did not expect that the total suspended particulate concentration would change as a function of temperature (23). Also, the majority of atmospheric particles (about 80%) that are removed during a rain event are washed-out during the first 5-10 mm of rain (30,31). Because all of our samples (with one exception) fall within or above this range, we have assumed, to a first approximation, that each rain event removed an equal number of atmospheric particles. As expected, multiple regression analysis showed that both rainfall volume (V, in liters) and ambient air temperature (T, in "C) determined the particle-bound concentrations (Cr,p)in rainwater C,, = -8.6V - 10T 478 (4)
+
The correlation coefficient ( r ) was 0.82, which was significant at the 95% confidence level. It should be noted that we found no correlation between rainfall volume and ambient air temperature ( r = 0.42). Although we cannot model the temperature dependence of the particulate scavenging because it is a nonequilibrium process (W, is not constant and affected by rainfall amount), we expect that many semivolatile organic compounds in the atmosphere will be scavenged most efficiently in cold weather. This is because, at colder temperatures, there are greater percentages of PCDD/F bound to particles. Dry Deposition of PCDD/F. Total dry PCDD/F fluxes for 18 samples measured in Bloomington ranged from 3 to 760 ng/(m2 yr). The average PCDD/F dry deposition fluxes for Bloomington were determined by geometrically averaging the PCDD/F fluxes measured for each sampling date. The geometric average homologue profile, representing an average ambient temperature of 17 "C, for PCDD/F in dry deposition is shown in Figure 4,top. The average total PCDD/F flux is 100 ng/(m2 yr). This profile is similar to PCDD/F homologue profiles of Bloomington air samples 115, 23). As with the rain data, we observed that the dry deposition flux decreased as the temperature increased. This is because particle deposition dominates dry deposition flux (32-34) and more PCDD/F are adsorbed to particles at cooler temperatures (23). Thus, temperature influences the amount of PCDD/F that are bound to particles that are subsequently dry deposited. The relationship between the dry deposition flux and atmospheric temperature can be derived starting from the equation which describes vapor-particle partitioning in the atmosphere (25,26) log [Ca,v(TSP)/Ca,pl =
+ a,/T
(5)
where C,, and Ca,*are PCDD/F atmospheric concentrations in the vapor and particulate-bound phases, respectively; TSP is the total suspended particulate concentration in the atmosphere; T i s the temperature (K); and a. and a, are regression coefficients. Assuming vapor-phase dry deposition is negligible, the dry deposition flux is defined as flux = C,,V,
(6)
where Vd is deposition velocity, the speed at which par-
80
-s
I
50-
-
E 40P
.
$30~ c
-
I
g 20-
E
-
"
4F
5F
6F
7F
8F
4D 5D 6D 7D ED
Flgure 4. (Top) Average PCDDIF dry deposttion fluxes [in ng/(m2 yr)] In Bloomington, IN, measured at 17 OC. (Bottom) Average PCDD/F dry deposition fluxes for Indianapolis, IN, measured at 26 "C. Numbers indicate level of chlorination; F indicates PCDF; D indicates PCDD.
ticulate-bound PCDD/F are removed from the atmosphere. Assuming that PCDD/F concentrations are lognormally distributed, we have log C,,"
+ log C,,,
(7)
=b
where b is a constant. Solving eq 7 for log C,,v, substituting this expression into eq 5, and rearranging the result we obtain -2 log C,,p = a0 - b - log (TSP) + a i / T (8) Expressing eq 6 in logarithmic form: log flux = log C Q ,+~ log
vd
(9)
Solving eq 8 for log C,,, and substituting this into eq 9, we obtain log flux = ( b - 4 / 2 log (TSP)/2 + log Vd - ~ 1 / 2 T (10)
+
Combining the constants ( b - ao)/2, log (TSP)/2, and log v d , and calling the result k, we have log flux = k - a,/2T (11) Thus, the logarithm of the flux attributed to particle dry deposition should be proportional to inverse temperature. Figure 5 shows a plot of the logarithm of the total PCDD/F flux versus 1000/T (the average temperature during each sampling period) for nine dry deposition measurements in Bloomington. The correlation coefficient (r) is 0.82, which indicates a significant correlation at the 95% confidence level. The slope of the regression line ( 4 2 ) is 3400 and its intercept (k)is -9.6. This regression equation indicates that total PCDD/F dry fluxes will decrease by a factor of 17 as the temperature increases from 0 to 30 "C. The above dry deposition measurements were all made in Bloomington. Dry deposition was also sampled during July 1991 in Indianapolis. Figure 4, bottom, shows the geometric average of the four measured PCDD/F dry deposition fluxes. The Indianapolis dry deposition profile resembles an urban air sample (23) because it contains considerable quantities of homologues other than 8D. In Environ. Sci. Techno\., Vol. 26, No. 7, 1992
1379
5D 1.2"""""' 3.3 3.35
3.4
"
* " " "
3.45
3.5
" '
3.55
"
'
" '
3.6
1
3.65
1000/T (K)
Flgure 5. Logarithm of the total PCDD/F flux [in ng/(m2 yr)] versus 1OOO/ T (K) for the measurements of Bloomington dry deposition.
Table 11. Dry Particle Deposition Velocities ( Vd, in cm/s) for PCDD/F Homologues in Bloomington (Blm; Measured at 17 "C) and in Indianapolis (Indy; Measured at 26 "C)
PCDD/F
Blm
Indy
PCDD/F
Blm
Indy
4F
0.16 0.064 0.12 0.13 0.13
0.12 0.090 0.086 0.32 0.60
5D 6D 7D 8D
0.15 0.16 0.14 0.37
0.34 0.14 0.12 0.21
5F 6F
7F 8F
contrast, the dry deposition profile of Bloomington (see Figure 4, top) is dominated by 8D and resembles a more rural air sample (23). The average, total PCDD/F flux in Indianapolis is 130 ng/(m2 yr); this is two times greater than the Bloomington flux of 60 ng/(m2yr) calculated at the same temperature as the Indianapolis samples (26 "C) using eq 11. The flux in Indianapolis may be higher than in Bloomington because the total suspended particulate is greater in Indianapolis (80 pg/m3) than in Bloomington (50 pg/m3) (23) and because total PCDD/F air concentrations are higher in Indianapolis by a factor of 3 than in Bloomington (23). Deposition velocities can be estimated for each PCDD/F homologue from measured fluxes and atmospheric particle-bound concentrations; see eq 6. Deposition velocities calculated from the Bloomington and Indianapolis data are given in Table 11. Deposition velocities were calculated from the average fluxes measured in Bloomington at 17 OC (Figure 4, top) and the average Bloomington PCDD/F air concentrations reported by Eitzer (23). Deposition velocities were also calculated for the Indianapolis data based on the average flux at 26 OC (Figure 4, bottom) and the average Indianapolis air concentrations (20). Assuming vapor-phase dry deposition is negligible, the deposition velocity indicates how efficiently particle-bound PCDD/F are removed from the atmosphere. With two exceptions, the deposition velocities in Bloomington range from 0.12 to 0.16 cm/s. The average in Bloomington was 0.16 cm/s with a standard deviation of 0.08 cm/s. The deposition velocities in Indianapolis were more variable; the average was 0.23 cm/s with a standard deviation of 0.17 cm/s. The overall average deposition velocity was 0.19 cm/s. From these data we conclude that particle-bound PCDD/F are all removed with about the same efficiency. This also suggests that all atmospheric PCDD/F are bound to the same size particles in the atmosphere. Our calculated deposition velocities are in the range of those reported in the literature (33, 35). However, our values are lower than those calculated (-1 cm/s) using a mass balance approach for PAH by McVeety and Hites (36). There may be several reasons for this observed difference. First, water may be a more efficient dry deposition collector than the flat plate or frisbee samplers. 1380
Environ. Sci. Technol., Vol. 26, No. 7, 1992
Lake surfaces are not smooth because of wave motions; thus, impaction of particles on lake surfaces may occur to a greater extent than on surrogate surfaces. Also, lake surfaces are cooler than our sampler surfaces, which may cause some partitioning of PCDD/F from the particles into the vapor phase. It is possible that the particles to which PCDD/F are bound in the atmosphere have a different size distribution than those to which PAH are bound. Unfortunately, the size distribution of particles to which PCDD/F are bound is not known. These factors, in combination, may account for the observed differences in the deposition velocities.
Conclusions One of our goals was to understand the change in the PCDD/F homologue profile that occurs between sources, which have an almost uniform distribution of PCDD/F, and sinks, which have homologue profiles dominated by 8D. Both wet deposition and dry deposition of PCDDIF were measured, and the results can be incorporated into our understanding of the atmospheric fate of PCDD/F. Previously, we learned that wet deposition was most efficient for removing particle-bound PCDD/F and more highly chlorinated PCDD/F from the atmosphere (12). This will affect the homologue profile of PCDD/F seen in environmental sinks, such as sediments. Our observations of PCDD/F scavenging by rain in Indianapolis confirmed these conclusions. In addition, we observed that temperature affected rain scavenging of particle-bound PCDD/F. In colder weather, rain scavenging of particlebound PCDD/F was even more efficient because more atmospheric PCDD/F were bound to particles at lower temperatures. We also measured the dry deposition flux of PCDD/F. This was the first time that the dry deposition flux of PCDD/F had been directly measured. Temperature also influenced dry deposition. We also observed that the efficiency of removal by dry deposition for PCDD/F increased with increasing levels of chlorination. This also affects the homologue profile of PCDD/F found in environmental sinks. Having measured both wet and dry deposition, we calculated wet and dry deposition fluxes for Indianapolis and Bloomington to determine which process dominated PCDD/F deposition. We calculated total PCDD/F fluxes for 15 "C, the average global temperature. Wet deposition flux for Indianapolis was determined by calculating the particle-bound concentration, 190 pg/L (from eq 4, assuming an average rain volume of 16 L), adding the average dissolved-phase concentration of 34 pg/L, and assuming an average annual rainfall of 100 cm/yr (average Indianapolis rainfall for the past 30 years as reported by the National Weather Service). This gives a total wet flux of 220 ng/(m2 yr). The average Bloomington wet deposition flux was estimated to be 95% of the Indianapolis rain flux (see Figure 2) or a value of 210 ng/(m2 yr). The dry deposition flux in Bloomington was calculated (from eq ll) to be 160 ng/(m2 yr). The Indianapolis dry deposition flux was estimated to be twice that in Bloomington (see above) at 320 ng/(m2 yr). Total (wet and dry) deposition fluxes for Bloomington and Indianapolis were 370 and 540 ng/(m2 yr), respectively. These values are summarized in Table 111. PCDD/F fluxes to sediments determined by Czuczwa (8) are 230 ng/ (m2yr) to Siskiwit Lake, Isle Royale, MI, sediments and 1400 ng/(m2yr) to the average Great Lakes sediment. Total PCDD/F fluxes to both Indianapolis and Bloomington are lower (by factors of 3 and 4, respectively) than those to the average Great Lakes sediment. Our total
air
Table 111. Total PCDD/F Fluxes Measured from Wet and Dry Deposition in Indianapolis and in Bloomington (at 15 "C) and Measured to Aquatic Sediments" Indianapolis wet dry total Bloomington wet dry total Siskiwit Lake Great Lakes (av) Baltic Sea
x
0
source A
sediment
0
220 320 540 210 160 370 230 1400 1700
rain .r
d':;
Trout lake
(2)
"All units are ng/(m2 yr).
PCDD/F flux was also lower than the total PCDD/F flux, 1720 ng/(m2 yr), measured to sediment traps in the Baltic Sea (37). We speculate that these differences result from an underestimation of the dry deposition flux. As discussed above, water surfaces may be more efficient collectors of particle-bound PCDD/F than were our collectors. What do the results of this study contribute to our understanding of the changes that occur in the homologue profile as PCDD/F travel from sources to environmental sinks? To answer this question, we compared PCDD/F homologue profiles in different environmental compartments. We have used principal components analysis (Statistical Analysis System, Version 6.1, SAS Institute Inc., Cary, NC) to quantitatively compare homologue profiles of PCDD/F from sources (represented by fly ash samples) (a), air samples (23), wet deposition in Bloomington (23) and in Indianapolis, dry deposition in Bloomington and Indianapolis, and sediments (8). Total (vapor/dissolved-phase plus particle-bound PCDD/F) homologue concentrations have been used to obtain the PCDD/F profiles for air and wet deposition. Because the concentrations of all the samples differed, homologue concentrations were normalized to the total concentration of PCDD/F in each individual sample. This allowed us to focus on the differences in the PCDD/F profiles themselves without bias due to different absolute PCDD/F concentrations. Figure 6 shows a plot of principal component 2 versus principal component 1;these two principal components account for 54% of the variability in the data set. PCDD/F sources, represented by fly ash, lie to the right side of Figure 6, and sinks lie to the left. Urban air samples (Indianapolis) have profiles that more closely resemble sources than do rural air samples (Trout Lake). This indicates that as PCDD/F travel from their sources they are changed. In part, this change is a result of the partitioning of PCDD/F between the vapor and particulate phases by processes that are controlled by the compounds vapor pressure and the ambient temperature (15). We know that this change is not caused by photodegradation of PCDD/F bound to atmospheric particles (38);however, some of this change may be caused by photodegradation of vapor-phase PCDD/F (39) or by reactions of PCDD/F with atmospheric hydroxyl radicals (40). With four exceptions (three of which are Indianapolis dry deposition samples with principal component 2 greater than 3), PCDD/F homologue profiles for wet and dry deposition are similar to one another, which indicates that both wet and dry deposition contribute significantlyto the removal of atmospheric PCDD/F. The deposition profiles for PCDD/F are distinctly different from most air profiles. This is because scavenging by both wet and dry deposition efficiently removes particle-bound PCDD/F but not va-
0
2
4
Principal Component 1 Figure 8. Principal component 1 versus principal component 2 for PCDD/F homologue profiles of various envlronmental compartments. Refer to legend for identlflcation of sample type. Note that air samples for Indianapolis and Trout Lake are marked.
por-phase PCDD/F. For both wet and dry deposition, PCDD/F with higher levels of chlorination are scavenged most efficiently. The sediment samples lie close to the wet and dry deposition samples. Thus, both atmospheric degradation, which determines the PCDD/F profile in the atmosphere, and deposition, which controls the removal of PCDD/F from the atmosphere, contribute to the change in PCDD/F profile from source to sink. Acknowledgments We thank Louis Brzuzy for technical discussions and experimental assistance. Registry No. 4F, 30402-14-3; 5F,30402-15-4;6F, 55684-94-1; 7F, 38998-75-3; SF, 39001-02-0; 5D, 36088-22-9; 6D, 34465-46-8; 7D,37871-00-4; 8D, 3268-87-9.
Literature Cited (1) Kjeller, L.; Kulp, S.-E.; deWit, C.; Lex&, K.; Hasselsten, I.; Rappe, C.; Jonsson, P.; Jansson, B. Presented at the 11th International Symposium on Chlorinated Dioxins and Related Compounds, Research Triangle Park, NC, Sept 1991. (2) Ballschmiter, K.; Buchert, H.; Niemczyk, R.; Munder, A.; Swerev, M. Chemosphere 1986, 15, 901-915. (3) Markland, S.; Kjeller, L.-0.; Hansson, M.; Tysklind, M.; Rappe, C.; Ryan, C.; Collazo, H.; Dougherty, R. In Chlorinated Dioxins and Dibenzofurans in Perspective; Rappe, C., Choudhary, G., Keith, L., Eds.; Lewis Publishers, Inc.: Chelsea, MI, 1986. (4) Markland, S.; Anderson, R.; Tysklind, M.; Rappe, C.; Egeback, K.; Bjorkman, E.; Grigoriadis, V. Chemosphere 1990,20, 553-561. (5) Olie, K.; Vermeulen, P. L.; Hutzinger, 0. Chemosphere 1977, 6,455-459. (6) Lustenhouwer, J. W. A.; Olie, K.; Hutzinger, 0. Chemosphere 1980,9, 501-522. (7) Karasek, F. W.; Hutzinger, 0. Anal. Chem. 1986, 58, 633A-640A. (8) Czuczwa, J. M. Ph.D. Thesis, Indiana University, Bloomington, IN, 1984. (9) Oehme, M.; Fiirst, P.; m e r , C.; Meemken, H. A.; Groebel, W. ChemosDhere 1988.17. 1291-1300. (10) Hites, R. A: Acc. Cheh. Res. 1990, 23, 194-201. (11) Czuczwa, J. M.; Hites, R. A. Enuiron. Sci. Technol. 1986, 20, 195-200. (12) Eitzer, B. D.; Hites, R. A. Enuiron. Sci. Technol. 1989,23, 1396-1401. (13) Tashiro, C.; Clement, R. E.; Reid, N.; Orr, D.; Shackleton, M. Chemosphere 1989,19, 535-540. (14) Reid, N. W.; Orr, D. B.; Shackleton, M. N.; Lusis, M. S.; Tashiro, C.; Clement, R. E. Chemosphere 1990, 20, 1462-1467. Environ. Scl. Technol., Vol. 26, No. 7, 1992
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Eitzer, B. D.; Hites, R. A. Enuiron. Sci. Technol. 1989,23,
Ligocki, M. P.; Leuenberger, C.; Pankow, J. F. Atmos. Enuiron. 1985, 19, 1619-1626. Ambre, Y.; Nishikawa, M. Atmos. Enuiron. 1987, 21,
1389-1395.
Hall, D. J.; Upton, S. L. Atmos. Enuiron. 1988, 22, 1383-1394.
1469-1471.
Hessen, T. C.; Young, D. R.; McDermotbEhrlich, D. Atmos. Enuiron. 1979, 13, 1677-1680. Christensen, E. J.; Olney, C. E.; Bidleman, T. F. Bull. Enuiron. Contam. Toxicol. 1979, 23, 196-202. Murphy, T. J. Atmos. Enuiron. 1981, 15, 206-207. Koester, C. J. Ph.D. Thesis, Indiana University, Bloomington, IN, 1991. Gatz, D. F.; Dingle, A. N. Tellus 1971, 23, 14-26. Lindberg, S. E. Atmos. Enuiron. 1982, 16, 1701-1709. Eitzer, B. D. Ph.D. Thesis, Indiana University, Bloomington, IN, 1989. Ligocki, M. P. Ph.D. Thesis, Oregon Graduate Center, Beaverton, OR, 1986. Yamasaki, H.; Kuwata, K.; Miyamoto, H. Environ. Sci. Technol. 1982, 16, 189-194. Bidleman, T. F.; Foreman, W. T. In Sources and Fates of Aquatic Pollutants; Hites, R. A,, Eisenreich, S. J., Eds.; American Chemical Society: Washington, DC, 1987, pp
Lim, B.; Jickells, T. D.; Davies, T. D. Atmos. Enuiron. 1991, 25A, 745-762. Bidleman, T. F.; Christensen, E. J. J. Geophys. Res. 1979, 84, 7857-7862.
Eisenreich, S. J.; Looney, B. B.; Thornton, J. D. Enuiron. Sci. Technol. 1981, 15, 30-38. Farmer, C. T.; Wade, T. L. Water,Air, Soil Pollut. 1986, 29,439-452.
Sievering, H. Atmos. Enuiron. 1987, 21, 2179-2185. McVeety, B. D.; Hites, R. A. Atmos. Enuiron. 1988, 22, 511-536.
Broman, D.; N 8 , C.; Zebiihr, Y. Enuiron. Sci. Technol. 1991, 25, 1841-1850.
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Received for review December 17,1991. Accepted March 30,1992. This work was supported by U.S. Department of Energy Grant 87ER- 60530.
Evidence for Rapid, Nonbiological Degradation of Tributyltin Compounds in Autoclaved and Heat-Treated Fine-Grained Sediments Peter M. Stang,",t Richard F. Lee,$ and Peter F. Seiigmans Applied Technology Division, Computer Sciences Corporation, 4045 Hancock Street, San Diego, California 92 1IO, Skidaway Institute of Oceanography, P.O. Box 13687, Savannah, Georgia 31416, and Naval Ocean Systems Center, Code 52, Environmental Sciences Division, San Diego, California 92152-5000
Sterilized sediments, high in silt and clay content, from various sites in the United States rapidly degraded added 14C-labeledtributyltin (TBTX) or unlabeled TBTX to dibutyltin (DBTXJ, monobutyltin (MBTX,), and inorganic tin. This degradation was primarily abiotic, as documented by similar degradation rates between sterilized and nonsterilized sediment. Degradation occurred in two phases, with a rapid degradation phase (23-94%) after 2 days, followed by slower degradation rates of the remaining TBTX during the next 5-7 days. DBTXz is the primary degradation product found when TBTX is added to marine water, with very little production of MBTX,. For our sediment studies, as well as those of others, the primary degradation product when dissolved TBTX is added to fine-grained sediment was MBTX,. MBTX, is formed in the sediment and, because of ita hydrophilic nature, enters the water shortly after its formation. We suggest that production of MBTX, by bottom sediments is why the ratios of MBTX,/TBTX were significantly higher in bottom water than surface water in a San Diego Bay marina.
Introduction Tributyltin compounds have been and currently are in use as biocides in marine antifouling paints and have been found in marine waters and sediments in areas exposed *Author to whom correspondence should be addressed. Current address: 3631 Caminito Carmel Landing, San Diego. CA 92130. +Computer Sciences Corp. Skidaway Institute of Oceanography. 9 Naval Ocean Systems Center. 1382
Environ. Sci. Technol., Vol. 26, No. 7, 1992
to marine traffic (1-6). Several studies have reported degradation of tributyltin compounds in water (7-9) and in sediment and soil (1G12). Degradation by bacteria and algae was determined to be primarily responsible for the degradation of tributyltin compounds to less toxic products. Although tributyltin compounds in pure form may exist as the chloride, fluoride, hydroxide, oxide, or other species, we shall refer to tributyltin in the generic form. Therefore, we shall refer to tributyltin as TBTX, dibutyltin as DBTX2, and monobutyltin as MBTX,. In sediments, TBTX has often been associated with elevated concentrations of DBTX2 and MBTX, in the sediment. MBTX, and DBTXz were both found in conjunction with the presence of TBTX in harbor and marina sediments from Great Bay Estuary, New Hampshire (13). DBTX2 concentrations exceeded TBTX concentrations in the sediments of Poole Harbour, England, in five of seven survey periods (14). The authors suggested that microbial degradation of TBTX to DBTXz was taking place and presumed DBTXz degraded to MBTX, and Sn, although these compounds were not analyzed for. TBTX was detected in three of four sediment samples and DBTXz in all four samples collected from Lake Biwa, Japan (15). MBTX,, DBTX2, and TBTX were reported in the sediment of San Diego Bay, CA (IO),with TBTX being the dominant butyltin compound only in the vicinity of vessel repair/dry dock facilities, where the majority of TBTX is likely present in particulate, paint chip form. Areas where TBTX entered the sediment by sorption to particles from the dissolved state were dominated by increased levels of DBTX2and MBTX,. The half-life of TBTX in sediment has been reported as 5.5 months in marine sediments (10)
0013-936X/92/0926-1382$03.00/0
0 1992 American Chemical Society
