Environ. Sci. Technol. 1990, 24, 1027-1032
Biogeochemistry of Butyltins in an Enclosed Marine Ecosystem Davld Adelman,? Kenneth R. Hinga,’ and Michael E. 0. Pllson
Graduate School of Oceanography, University of Rhode Island, Narragansett, Rhode Island 02882 ~~
Tributyltin (TBT) and its degradation products were studied by introducing radiolabeled tributyltin into a 13-m3 marine enclosure (a MERL mesocosm) with near-natural water column and benthos. TBT and its degradation products were monitored for 278 days. TBT concentrations in the water column (initially 590 f 20 ng L-l) decreased at a rate of 0.20 day-l for 15 days and then slowed to 0.10 day-’. Most TBT was lost from the water column through biodegradation, which occurred at a rate of 0.08 day-’. Two-thirds of the degradation proceeded through debutylation to dibutyltin (DBT), which in turn degraded to monobutyltin (MBT) at 0.04 day-l. One-third of the TBT was degraded directly to MBT. There was no evidence for degradation of MBT in the water. Another portion of the TBT removed from the water column was transported to sediments. TBT in the sediments did not appear to measurably degrade. A portion of the TBT was apparently transported rapidly to the air-water interface and then was lost from the tank.
-
Introduction Concern over the use of tributyltin (TBT) in antifouling paints results primarily from its extreme toxicity to marine organisms, especially mollusks. In order to mitigate problems evident in a variety of marine populations, several countries have banned the use of TBT-containing paints on boats less than 25 m in length. TBT use on larger vessels is still permitted. The risk to the environment from TBT use is a function of both the toxicity of the chemical and its persistence and behavior in the environment. When a chemical is subjected to rapid degradation, calculation of its risk to the environment is especially sensitive to factors other than toxicity. Previous work indicates that TBT is fairly rapidly degraded in marine waters (at least at warm temperatures) and is subject to scavenging to benthic sediments and to enrichment at the air-water interface (1,2). But, there are still relatively few studies that contribute to a picture of TBT persistence and behavior in marine environments (I,.% The objective of the research reported here was to determine the behavior of TBT and its degradation products under conditions representative of a temperate, shallow, coastal marine ecosystem. Radiolabeled TBT was added to a marine enclosure mesocosm at the Marine Ecosystems Research Laboratory (MERL), University of Rhode Island, and the distribution of label followed for 278 days. When operated without deliberate manipulations, the biological communities and chemical processes in the MERL mesocosms are very similar to those observed in adjacent lower Narragansett Bay, even when operated in “batch mode” for up to 8 months (3-6). A chemical introduced into a MERL mesocosm should behave as it would in a natural environment with similar characteristics. The MERL mesocosms have been usefully applied to the study of the behavior of a variety of chemicals (7-13). In this case the mesocosm is used to see if indit Present address: Environmental Sciences Services, Providence,
RI 0013-936X19010924-1027802.5010
vidual processes, as measured in the laboratory, are operative in such a relatively complete ecosystem. Experimental Section The 1.8 m diameter X 5.5 m high mesocosm contained sediment and water transferred from adjacent Narragansett Bay. A 30 cm deep layer of sediment, complete with its active biological community, was collected intact as a series of box cores and transferred to the mesocosm tank. Water from the bay was added to the mesocosm with a nondisruptive bellows pump. Nutrients and plankton in the mesocosm were representative of typical conditions in the adjacent bay in early July. At 2030 h, 13 July 1987,480pCi of Tri-n-[1-l4C]butyltin chloride dissolved in 10 mL of MeOH was added to a MERL mesocosm. The TBT was made available through the courtesy of D. Rudnick and the US. EPA Gulf Breeze Laboratory. The manufacturer (Amersham Corp.) gave the specific activity of the compound to be 21 mCi mmol-’. The TBT as received had -7% contaminants and was cleaned by preparatory thin-layer chromatography (TLC). The spike resulted in an initial tributyltin chloride concentration of 590 f 20 ng L-l, as determined by liquid scintillation counting (LSC) of three replicate samples taken 1.5 h after spiking the mesocosm. This concentration is about 0.01 of the saturation value for TBT in seawater for the temperature at the spike (14). Samples taken for analysis were as follows: whole water, solvent-extractable and nonextractable fractions of whole water, dissolved inorganic carbon (C[CO,]), suspended particles, &water interface, and sediment. Water column samples were taken during mixing periods when the water column is homogeneous (the mesocosm was mixed 2 h out of every 6). In all samples, 14Cconcentrations were determined by LSC. All liquid scintillation counts were corrected for background and counting efficiency for each type of sample and were counted for sufficient time to reduce the uncertainty of activity in the sample to less than 1%. A brief description of analytical procedures is provided here. Further details may be found elsewhere (14, 15).
Total activity in the water (before extraction) was measured by combining 10 mL of sample with 10 mL of scintillation fluid in a vial and counting. The concentration of total 14C02(C[’4C02]),generated from the degradation of labeled butyl groups, was measured by acidifying 500 mL of whole water samples, stripping with Nz gas (which was passed through an activated carbon filter to remove volatilized butyltins), and trapping with phenethylamine (PEA). Activity in the PEA was then counted. Processing of water for the quantification of activity associated with individual butyltin species consisted of the following: extraction of an acidified water sample twice with hexane, concentration by vacuum evaporation, separation by thin-layer chromatography (TLC), removal of appropriate sections of TLC media, and activity determination by LSC. Total activity remaining in the acidified water after hexane extraction (which also removed [CO,] from samples) was measured by combining 10 mL of extracted water with 10 mL of scintillation fluid and counting.
0 1990 American Chemical Society
Environ. Sci. Technol., Vol. 24, No. 7, 1990
1027
Preparatory work for this experiment showed that the three butyltins were extracted from the water with different efficiencies and were subjected to different losses during sample concentration (14, 15). The hexane extraction efficiencies were 95 f 3% for TBT, 55 f 15% for DBT, and 11 f 3% for MBT. Losses of TBT and DBT during vacuum evaporation were 1 2 f 6% and 67 f 3%, respectively. Losses of MBT were great enough that MBT could not be adequately evaluated on the TLC plate after evaporative concentration. The activities found as tri-, di-, and monobutyltin were calculated by a simple system of simultaneous equations and the constants given above. This procedure is based on the assumption that the only components remaining in the sample after the extraction process are TBT, DBT, and MBT. The concentrations of the three butyltins may be calculated as f = (TBTTLc/0.88)/(TBTTLc/0.88 + DBTTL~/O.~~)
Month
. S
J - A
D
0 . N
J
F
M
- o l0
200
100
0
Day P
1.0
P
9
8 -
0
0.8
s
+
'5 0.6
Total Extracted Nonextracted
.-c
0.4
O
0.2
E
0.0
Li
-
100
0
200 Day
[TBT] = (Aex - 0.11[MBT])f/0.95 [DBT] = [(A,, - O.ll[MBT])(l - f)]/0.55 [MBT] = [R,, - (0.05[TBT] + 0.45[DBT])]/0.89
In these equations the constants represent the extraction efficiencies discussed above and f is the fraction of activity in the organic phase attributed to TBT after correction for losses during vacuum evaporation; TBTTLC is the activity in the TBT region of the TLC plate; DBTTLCis the activity in the DBT region of the TLC plate; [TBT] is the activity of TBT in the original sample; [DBT] is the activity of DBT in the original sample; [MBT] is the activity of MBT in the original sample; A, is the activity extracted from the sample; and Ra, is the activity remaining in the aqueous phase after extraction. Molar concentrations of TBT, DBT, and MBT were then calculated from the activity of each compound. Given the symmetry of the TBT molecule, all butyl groups are equivalent. As each butyl group is lost, the specific activity of the remaining compound is reduced. DBT has 2/3 and MBT has the specific activity of the parent TBT. For TBT, the manufacturer-provided specific activity gave a conversion factor of 1.0 dpm g-' = 0.024 nmol L-l. Appropriate DBT and MBT factors are then 1.0 dpm g-' = 0.036 nmol L-l and 1.0 dpm g-l = 0.072 nmol L-', respectively. Both the raw activity of the fractions and the calculated molar concentrations are useful parameters and will be discussed below. The butyltins on suspended particles were analyzed by filtering samples through two stacked glass fiber filters (Whatman GF/C). The measured activity absorbed by the second filter paper was used to correct for butyltins adsorbed directly from the water sample onto the first filter. (The fraction of TBT adsorbed from the water was small, so both filters were exposed to essentially the same dissolved concentration of TBT.) The blank correction was necessary as 50 f 20% of the activity found on the first filter was absorbed directly from the water. The top filter and blank were extracted separately with hexane for 16-20 h, and the extract was separated by TLC, each fraction counted, and a correction made for loss during solvent evaporation. Concentrations of TBT and DBT on particles were calculated from the total activity measured on particles and from TBT to DBT ratios extracted from particles. Total weights of suspended particles were determined by filtration through a 0.4-pm Nucleopore filter. Distribution coefficients [Kdsin units of (pg/kg)/(pg/L)] were calculated for TBT and DBT from the particulate extract data, total wet concentrtions, and measured sus1028 Envlron. Sci. Technol., Vol. 24, No. 7. 1990
"
0
"
I
'
"
'
, .
I
100
,
200
Day
Flgure 1. (A) Water temperature in the mesocosm and month. (B) Total activity in the mesocosm and extractable and nonextractable fractions. (C) Activity of radiolabeled x [ C O , ] In the water column.
pended particle concentrations. The Kd for MBT was determined from total activity in the water and on particles after concentrations of TBT and DBT had fallen below 10% of total activity (days 40-52). During the first 30 days of the experiment, the surface microlayer was sampled, both between and during mixing cycles, by use of a stainless steel Garrett screen, and the total activity determined. The screen used is reported to sample the upper 440 f 20 pm of water (11). Sediment cores were obtained with a corer that does not lose the flocculant layer at the surface of the sediment (16) and were immediately frozen. Total activity was determined in the upper 5 cm of sediment. In other experiments with radiolabeled substances in the MERL mesocosms, little activity had penetrated to this depth (7). Cores were first extracted with one of a variety of extraction procedures [aspart of an evaluation of techniques, (15)]. The extracted sediment was then combusted and the evolved C02 trapped in phenethylamine (PEA) to determine activity remaining in the sediment. The extracted activity and the 14C02activity were summed to comprise total activity in the cores. Results
The experiment started when the temperature was near the summer maximum and progressed through the winter and into the spring (Figure 1A). The total activity in the water column was initially 78 f 2 cpm g-' of water. Both total and solvent-extractable activities decreased immediately after the spike, with the extractable fraction decreasing more rapidly (Figure 1B). This was accompanied by a rapid increase in nonextractable activity (which does not include 14C02). After day 50, activity in all three fractions remained relatively constant. The concentration of 14C02increased through day 50 and then slowly decreased (Figure 1C). The decrease in
Table I. Distribution Coefficient (Kd)of ButyltinsO
,
0
-20%/d
10
20
I ..
# " '
30
40
Day Rate
U
ref
comment
10-3~~ 0.6-8.2 6.2-55 0.2-1.6 2.1-26 0.7-2.0 1.8-29 2.8-4.5
TBT TBT TBT DBT DBT MBT MBT
3.4-9.3 60 f 30 0.6 30 f 20 2.6 2.9 f 0.5
TBT TBT DBT DBT MBT MBT
DBT
Sediment dependent on sed. type in situ dependent on location in situ dependent on location in situ dependent on location Suspended Particles dependent on time of year
20 21 22 21 22 21 22 22 this study 22 this study 22 this study
aValues in h a l k d / ( u d L ) .
significant amounts of DBT and MBT were found. The spike was added during a 2-h mixing cycle. The mixers homogenize the mesocosm water column in 15 min, so TBT should have been uniformly mixed into the mesocosm after 2 h. Total activity on particles after 2 h was 3.42 dpm g-' of water, which represented 4.6% of activity in the water. The concentration of particles in the water column was 680 Fg L-l. The partition coefficient for TBT with this approach was 7.1 X lo4, which agrees with the partition coefficient determined from the exact procedure. In a similar fashion, a partition coefficient for MBT was calculated from data when MBT accounted for over 90% of the butyltins in the water (days 40-52). By use of total activity on particles and total activity in the water (four determinations), the Kd for MBT was estimated to be (2.9 f 0.5) X lo3. The distribution coefficients found here for TBT and DBT are somewhat higher than previously reported values, while the & for MBT is similar (Table I). On the first day of the experiment, activity in samples of the surface microlayer (air-water interface) was 220 f 40 dpm g-l. The concentration was higher than that in the underlying water column by a factor of 3.0 f 0.1. Over the first 30 days the surface layer was sampled eight times. The enrichment factor decreased steadily with time, reaching a value of 0.75 f 0.01 by day 29. Activity in the sediment, after the first 2 weeks, accounted for 30% of activity added to the mesocosm (Figure 3A). There was considerable variability between replicate cores, which may obscure some trends, but the activity in the sediment did not appear to change significantly after day 14 (43 cores were taken at 11sampling dates). Activity in the sediment, combined with activity retained in the water column, indicated that approximately 70% of the activity remained in the mesocosm.
-
-4 0
0.0 0
10
20
30
40
Day
Figure 2. ButyRin concentrations with time. (A) Filled circles are measured concentrations of TBT in the water column. Lines were fit to the data and show the first-order removal rates. (6) Filled circles are the measured concentrations of DBT in the water column. Lines are models for DBT decay rates of 4, 6, and 8% per day assuming a 8 % day-' degradation rate of TBT. (C) Filled circles are the measured concentrations of MBT in the water column. The lines are modeled MBT concentrations by assuming 33% of the TBT decays directly to MBT and DBT degradation rates of 4, 6, and 8% per day.
l4COZafter day 50 reflected loss to the atmosphere, which was then greater than generation of 14C02from degradation of labeled butyl groups. The loss rate of 14C02from MERL mesocosms to the atmosphere has been found to be as high as 0.01 day-' in the summer and considerably less in winter. The decrease in the mesocosm concentraday-'. tion for days 50-250 was 7 X Figure 2 shows the concentrations of TBT, DBT, and MBT over the first 42 days. After that time, activity in the extract was too low to allow a determination of the TBT and DBT activities. The sum of water column activities of TBT, DBT, MBT, and COz accounted for 90 f 10% of the total measured activity over this time. Calculated TBT concentrations decreased rapidly after the spike (Figure 2A). For the first 15 days the total decrease in TBT concentration appeared to be a first-order process with a rate constant of 0.20 day-'. The rate then slowed to 0.10 day-' for days 15-42. The concentration of DBT increased from an initial mole fraction of 0.05 to a maximum of 0.24 on day 10 (Figure 2B) and decreased thereafter. The concentration of MBT increased throughout the first 50 days and then stabilized (Figure 2 0 . The distribution of TBT and DBT activity between the dissolved state and that associated with suspended particles was calculated by using data from days 2-19 (13 values). The distribution coefficients (K@)for TBT and DBT were (6 f 3) X lo4 and (2 f 2) X lo4, respectively, and had no apparent trend with time. An independent calculation of the TBT Kd was made by examining total activity on particles immediately after the spike and before
Discussion Of the activity added to the mesocosm, 60% had left the water column after 50 days. Activity leaving the water column could have gone to any of three places: the sediment, the atmosphere, or the mesocosm walls. After the first 2 weeks, the radiolabel in the sediment accounted for -30% of the activity added to the mesocosm. Similarly, after the first 2 weeks, the radiolabel in the water appeared not to change except for the loss of radiolabeled COz to the atmosphere. There did not appear to be any remobilization of activity in sediments back to the water column. The sediment extractions and quantification of fractions did not remove enough of the radiolabel (typically Environ. Sci. Technol., Vol. 24, No. 7, 1990
1029
; 1.0 U
2 *
.5
Total
0.5 0.6
Water
L
0.4 0.2
ri I ; 0.0 50
0 0.6
0.4
100 Day Activity
1
150
in aediment
1
1
0
10
20
0
10
20
30
40
50
30
40
50
Day
Figure 3. Radiolabeled butyitin activity. (A) Total activity in the sediments, total activity in the water column, and the sum of activity in water and sediment, which is the total actitivy in the mesocosm. (9) Filled circles are measured activity in the sediment. Lines are predicted sediment concentrations for TBT transport rates to the sediment of 4 and 5 % per day. Error bars represent 1 standard deviation. (C)Open circles are the total meSOcosm activity. Lines are expected activities for various TBT loss rates to the atmosphere.
only 50%) to rigorously interpret the concentrations of individual compounds (15). However, nearly all the radiolabel extracted was TBT and there did not appear to be any change in the ratio of TBT to the total sediment activity with time. The simplest interpretation is that once TBT was incorporated into the sediments, its degradation was greatly slowed or stopped. This may be in part due to the low temperatures for much of the later part of the 280-day experiment. When the amount of degradation that would have been observable given the procedures used was estimated, the degradation of TBT in the sediment was less than -1 X lo-" day-l. The other 30% of the activity leaving the water column was also lost quickly. The total activity in the water column decreased steadily and rapidly for the first few days of the experiment (Figure 1B). At the end of 7 days, when the first sediment samples were taken, the s u m of sediment and water column activity was only 70 f 20% of the total found in the water column immediately after the spike (Figure 3A,C). Hence, a 30% loss of activity from the mesocosm appeared to occur during the first week of the experiment. It was possible that some of the lost activity was adsorbed to the mesocosm walls. However, there was no indication that any activity was desorbed from the walls during the experiment. At the end of the experiment, the mesocosm was drained and portions of the walls scrubbed with either detergents or methanol. A rough pad was used so some wall material was removed in the washes. The washes were sampled for total activity. Only 0.9% of the activity added to the mesocosm was accounted for in 1030
Environ. Sci. Technoi., Vol. 24, No. 7, 1990
material washed from the walls. The mesocosm walls did not appear to represent a significant long-term sink for butyltins. Each of the three potential reservoirs in the mesocosm was sampled by a technique that quantitatively collected and counted the total 14Cin the sample. Since all reservoirs were sampled, we must conclude that the activity that could not be found in sediments or water was lost to the atmosphere. The rate of this loss was 2-3 times faster than has been found for volatilization of gaseous alkanes in the MERL mesocosms (IO). As volatilization of TBT would be expected to be considerably slower than nonionic gases, this high rate is not consistent with models that describe volatilization. These models require the chemical to diffuse through a nonmixed, diffusive, layer at the air-water interface. A different process may have been operating here. For the first few days of the experiment, the surface samples, taken both during and between mixing cycles, were clearly enriched in total activity relative to the bulk water column. Assuming all the activity was TBT on day 1, the concentration in the surface microlayer of the mesocosm was 1.7 f 0.3 pg of TBT L-l. This concentration is similar to maximum reported environmental surface layer concentrations of 1.2 (17)and 2.7 pg of TBT L-I (18). Such enrichments were probably the result of hydrophobic chemicals accumulating at the air-water interface, as has been reported for many compounds. The surface enrichment in this experiment was no longer found after TBT concentrations in the water decreased to low levels. It appears that there was some mechanism that scavenged TBT (but not enough DBT to be important or any MBT) to the surface layer. What is of note here was the rapidity with which the TBT was brought to the surface layer. After rapid transfer to the surface microlayer, the material must have volatilized, which resulted in a significant loss of labeled material from the system. Whether TBT volatilized directly, or was first photodegraded, is a question that cannot be answered with our data. It does not appear necessary to invoke photodegradation prior to volatilization. We have noted in procedural checks in the laboratory that spots of radiolabeled TBT placed on a variety of surfaces completely volatilized overnight when there was good air circulation. TBT photodegradation rates in water are reported to be slow (1,2);however, these may not apply to TBT at the air-water interface. Losses from the water column through surface microlayer enrichment and volatilization may be an important process in the field. This mechanism was also suggested to help explain petroleum hydrocarbon behavior in these mesocosms (11). Since the processes that control the extent and rate of surface enrichment are not known for either the field or mesocosm situations, we do not know whether they may be enhanced in the mesocosm environment. Further work will be required to establish the importance of the process in the field. The total decrease of TBT concentration in the water column was consistent with first-order kinetics. Initially the rate constant was 0.20 day-' (half-life, 3.5 day). A rate change to 0.10 day-l (half-life, 6.9 day) occurred around day 15. The change in removal rate roughly correspond to a change in chlorophyll concentration over the same period, from 15 f 6 to 8 f 4 pg L-l. A similar correlation of degradation rate to chlorophyll concentrations has previously been reported (19). The gross removal rate of TBT from the water column was the result of biological degradation, scavenging to sediment, and the presumed loss to the atmosphere. If we
Table 11. The Half-Life of Butyltins in Freshwater and Seawater comment
ref
4 months 2 months 16 days 9 days 8 days 7 days 6-19 days 6 days 6-7 days 4-13 days
Degradation of TBT sterile freshwater/sediment mix marine sediment aerobic freshwater/sediment mix seawater at 5 O F freshwater aerobic microbes seawater/sediment at 20 "C freshwater anaerobic microbes estuarine microbes estuarine waters freshwater at 5 O F estuarine water estuarine water
24 25 26 27 26 this study 26 28 29 27 28 19
6-12 days
Total Removal Rate seawater/sediment at 20 O F
this study
60 days 12-18 davs
Degradation of DBT seawater at 5 O F seawater/sediment at 20 O F
27 this studv
half-life >11 months 5.5 months
assume each of these processes followed first-order kinetics, the experimental results can be modeled to provide rate constants for each process. Transport rates to the sediment were assumed to occur in the ratio of the distribution coefficients of TBT and DBT (Table I). By use of this assumption, and the assumption that there were no net losses from the sediment, a transport rate for TBT of 0.045 day-' predicts the total activity found in sediments (Figure 3B). This scavenging rate was consistent with previous experiments in MERL with other compounds of similar Kds (7, 8). The scavenging rate required a replacement rate for suspended particulate matter of 1.0 day-'. Which is similar to replacement rates previously found both in the MERL tanks and in Narragansett Bay in summer (23). The volatile loss was modeled by assuming loss of TBT only. A rate constant of -0.075 day-' fits the data for total activity in the mesocosm (Figure 3C). The biological degradation rate for TBT may be calculated as the difference between gross removal rate and the two transport rates calculated above. The calculated difference was 0.08 day-' (half-life, 9 days). This value is similar to degradation rates reported by others for marine waters (Table 11). The rate of disappearance of DBT after day 10 equaled the rate of increase of MBT. Attempts to fit degradation rates to the data were not possible with the assumption that TBT degraded by successive debutylations through DBT to MBT. In the first 10 days, no degradation rate of DBT could describe both the DBT and MBT data. It was necessary to assume that a fraction of the TBT degraded directly to MBT, as has been found in experiments with several forms of microbes (30). By assuming that one-third of the TBT degraded directly to MBT, the data could be fit with simple first-order equations. A degradation rate of 0.05 f 0.01 day-' for DBT then predicted the overall trends of measured DBT and MBT concentrations (Figure 2B,C). The calculated DBT and MBT concentrations in the first 6 days were poorly described by modeling the degradation rates. This may have resulted from interference in the analysis by a degradation product other than COz that was not measured by our procedures. Evidence for such a compound, or compounds, was found by considering the 14C02concentrations. The concentration of DBT reached a maximum of 10 dpm g-' on day 10. The concentration of 14C02resulting
-
from complete remineralization of the butyl group released during DBT production should have been 5 dpm g-l. The concentration of MBT on day 10 was 7.5 dpm g-', indicating that 15.0 dpm g-' C02 should have been produced from the butyl groups released during MBT production. The measured concentration of C02 was -7 dpm g-' or one-third of the total expected value. This imbalance did not persist for more than a few days. The results could be explained by assuming that butene was formed, as has been suggested previously (19). Butene itself decays at rates of 0.027-0.043 day-' in these systems in summer (10) and could have temporarily held some of the labeled carbon. Conclusions
TBT may be removed from marine waters by three process: biological degradation, scavenging to sediments, and probably by transport to the air-water interface followed by volatilization or photodegradation and volatilization. Under the summer conditions at the start of this experiment, these removal processes were all important. Predictions that rely only on biological degradation for removal of TBT are likely to overestimate TBT concentrations. Butyltins in the sediment did not appear to appreciably degrade. Cold temperatures during much of the experiment may have been partly responsible for this observation. The concentrations of DBT and MBT could only be described by assuming that a significant portion of TBT degrades directly to MBT. It should not be assumed that the loss of TBT will be equal to the appearance of DBT. MBT did not degrade once formed, nor did it appear to be subject to other removal processes. Acknowledgments
We thank J. Quinn for his helpful suggestions during the course of the experiment. We also thank the staff at MERL for their help and support. Registry No. DBT, 14488-53-0;MBT, 78763-54-9;tributyltin chloride, 1461-22-9. L i t e r a t u r e Cited (1) Clarke, E. A.; Sterritt, R. M.; Lester, J. N. Environ. Sci. Technol. 1988, 22, 600. (2) Maguire, R. J. Appl. Organomet. Chem. 1987, 1, 475. (3) Pilson, M. E. Q. J . Mar. Res. 1985, 43, 849. (4) Hunt, C. D.; Smith, D. L. In Marine Mesocosms: Biological and Chemical Research in Experimental Ecosystems; Grice, G. D., Reeves, M. R., Eds.; Springer-Verlag: New York, 1982; p 111. (5) Santschi, P. H. In Marine Mesocosms: Biological and Chemical Research in Experimental Ecosystems; Grice, G. D., Reeve, M. R., E&.; Springer-Verlag: New York, 1982; p 63. (6) Donaghay, P. L. In Concepts in Marine Pollution Measurements; White, H. H., Ed.; Maryland Sea Grant College: College Park, MD, 1984; p 589. (7) Hinga, K. R.; Pilson, M. E. Q. Enuiron. Sci. Technol. 1987, 21, 648. (8) Santschi, P. H.; Amdurer, M.; Adler, D.; OHara, P. 0.;Li, Y.-H.; Doering, P. J . Mar. Res. 1987, 45, 1007. (9) Wade, T. L.; Quinn, J. G. Mar. Environ. Res. 1980,3, 15. (10) Bopp, R. F.; Santschi, P. H.; Li. Y.-H.; Deck, B. L. Org. Geochem. 1981, 3, 9. (11) Gearing, P. J.; Gearing, J. N. Mar. Environ. Res. 1982,6, 133. (12) Lee,R. F.; Hinga, K. R.; Almquist, G. In Marine Mesocosms: Biological and Chemical Research in Experimental Ecosystems; Grice, G. D.; Reeve, M. R., Eds.; Springer-Verlag: New York, 1982; p 123. Environ. Sci. Technol., Vol. 24, No. 7, 1990
1091
Environ. Sci. Technol. 1990, 24, 1032-1030
(13) Wakeham, S. G.; Canuel, E. A.; Doering, P. H. Environ. Sci. Technol. 1986,20, 574. (14) Adelman, D. A. M.S. Thesis, University of Rhode Island, 1988. (15) Adelman, D. A.;Hinga, K.R.; Pilson, M. E. Q. Environ. Monit. Assess., in press. (16) Frithsen,J. B.; Rudnick, D. T.; Elmgren, R. Hydrobiologia 1983, 99, 75.
(17) Hall, L. W., Jr.; Lenkevich, M. J.; Hall, W. S.; Pinkney,A. E.; Bushong, S. J. Mar. Pollut. Bull. 1987, 18, 78. (18) Cleary, J. J.; Stebbing, A. R. D. In Oceans '87; IEEE and Marine Technology Society: Washington, DC, 1987;p 1405. (19) Lee, R. F.; Valkirs, A. 0.;Seligman, P. F. In Oceans '87; IEEE and Marine Technology Society: Washington, DC, 1987; p 1411. (20) Unger, M. A,; MacIntyre,W. G.; Huggett, R. J. In Oceans '87; IEEE and Marine Technology Society: Washington DC, 1987; p 1381. (21) Stang, P. M.; Seligman, P. F. In Oceans '86; IEEE and Marine Technology Society Washington, DC, 1986, p 1256. (22) Valkirs, A. 0.; Seligman, P. F.; Lee, R. F. In Oceans '86; IEEE and Marine Technology Society: Washington, DC, 1986; p 1165.
(23) Hinga, K. R. Mar. Environ. Res. 1988,26, 97. (24) Slesinger, A. E.; Dressler, I. In The Organotin Workshop; Good, M., Ed.; University of New Orleans: New Orleans, LA, 1987; p 115. (25) Stang, P. M.; Seligman, P. F. In Oceans '86; IEEE and Marine Technology Society Washington, DC, 1986;p 1256. (26) Maguire, R. J.; Tkacz, F. J. J.Agric. Food Chem. 1985,33, 947. (27) Thain, J.; Waldock, M. J.; Waite, M. E. In Oceans '87;IEEE
and Marine Technology Society: Washington, DC, 1987;
p 1398. (28) Olson, G. J.; Brinckman, F. E. In Oceans '86; IEEE and Marine Technology Society Washington, DC, 1986;p 1196. (29) Seligman, P. F.; Valkirs, A. 0.;Lee, R. F. Enuiron. Sci, Technol. 1986,20, 1229. (30) Barug, D. Chemosphere 1981, 10.
Received for review May 9,1989. Revised manuscript received December 11, 1989. Accepted March 7, 1990. This work was funded primarily through a cooperative agreement between the Centers Program of the U.S. Environmental Protection Agency and the Marine Ecosystems Research Laboratory of The Univeristy of Rhode Island.
Nonlinear Pesticide Dissipation in Soil: A New Model Based on Spatial Variability David I . Gustafson' and Larry R. Holden
Monsanto Agricultural Company,+700 Chesterfield Parkway, St. Louis, Missouri 63 198 In both laboratory and field studies, the dissipation of pesticides in soil often fails to follow simple first-order reaction kinetics. Rather than being linear when plotted as In C versus time, the dissipation data are curved, typically concave upward. This nonlinear behavior has now been described successfully through the use of a new first-order, nonlinear kinetic model. The nonlinear model is based on the assumption of a spatially variable firstorder rate constant, and it reduces to the linear case when the rate constant is spatially uniform. Excellent fits to both laboratory and field data are obtained for all pesticides modeled. Interestingly, the relative variability found for the rate constant is similar for laboratory and field studies, suggesting that the length scale of the spatial variability is very small, possibly on the order of pore-size dimensions.
Introduction Soils are heterogeneous (1-6). In spite of this, most theoretical descriptions of pesticide dissipation kinetics in soil have relied upon models that assume some degree of homogeneity (7-16). This fundamental discrepancy may be responsible for the general failure of these models to adequately characterize observed dissipation curves (17-19).We take a new approach in this paper and show, by explicitly acknowledging the heterogeneity or spatial variability of soil, that one can derive a simple, nonlinear, first-order kinetic model of pesticide dissipation giving excellent fits to observed data. As used in this paper, the term dissipation refers to the group of processes that reduces the concentration (e.g., percent remaining or pounds/acre) of a pesticide following its application to a laboratory aliquot of soil or an experiment field plot. As such, it includes a host of biological, A unit of Monsanto Co. 1032
Environ. Scl. Technol., Vol. 24, No. 7, 1990
chemical, and physical phenomena that are in turn dependent on a number of soil, environmental, and cultural factors. Processes most often associated with dissipation are microbial degradation, chemical hydrolysis, volatilFactors ization, runoff, wind erosion and photolysis (7,12). affecting the rates of these processes include microbial population density, temperature, amount and intensity of precipitation, solar intensity, soil properties (moisture, organic matter, texture, sorptive capacity), tillage, and cropping practices (7,12).Several of these processes and factors are either unimportant, controllable, or ignored in laboratory studies, but all play a role in field studies to some degree. All of the factors affecting dissipation rates are spatially variable to some degree (1-6). For instance, the spatial variations in temperature across a newly planted cornfield are generally moderate, but the soil organic matter content can vary by at least an order of magnitude (1,20).Certainly at the microscopic level (where most dissipative processes occur) there are large variations in water content and soil sorptive capacity as different size soil pores are encountered ( I ) . If the applied pesticide were able to diffuse readily throughout the entire treated field or laboratory beaker, such spatial variability would not be as critical because the system would behave more like a single, well-mixed reactor. However, diffusion rates in soil are notoriously slow for all but the most volatile of pesticides (21),and the applied material is thereby confined to rather small regions of soil, which may be very different from others only a short distance away. Given the apparent complexity of dissipation, it is reasonable to question whether any single model could adequately characterize its rate in a quantitative manner. In fact, in a recent U.S.Environmental Protection Agency (EPA) document (17)it is stated that "it is doubtful that any single rate equation will ever be found which is applicable to all or most pesticides in soil". This caveat does
0013-936X/90/0924-1032$02.50/0
0 1990 American Chemical Society
