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table Table III-5.Summary of key tests of the fate, transport, and transfer models.  
expand table Table VX X-X

A summary of key discussions from the uncertainty evaluation is now presented. First is a summary of three exposure parameters common to all pathways:

1. Lifetime, Body Weights, and Exposure Durations:

Of these three parameters, the exposure duration is the most uncertain. The estimates of 9 and 20 years were made in this assessment for non-farming residents in rural settings, and farming residents in rural settings. These values were based on assumptions of time living at one residence.

A critical assumption of a constant soil concentration for contaminated soil sites should be carefully considered for site-specific assessments. Data on degradation indicates very slow rates of degradation, and only photolysis as a possible degradation mechanism, which would not impact residues below the surface.

A mass balance exercise on the demonstration of the off-site source category (where a 40,000 m2 area had soil concentrations averaging 1 ppb 2,3,7,8-TCDD) indicates that it would take 90 years to dissipate a reservoir of 2,3,7,8-TCDD extending 6 inches into the soil. An adult body weight of 70 kilograms and a lifetime of 70 years are standard assumptions for exposure and risk and, although variability is recognized for these parameters, these variations are not expected to add significant uncertainty in exposure estimates. The same is true for the 17 kg child body weight in the childhood exposure pattern of soil ingestion.

2. Soil Ingestion and Soil Dermal Contact:

Soil ingestion for older children and adults were not considered, which may have underestimated lifetime soil ingestion exposures. Pica soil ingestion patterns were not evaluated in this assessment. The ingestion rates (200 mg/day for central scenarios and 800 mg/day for high end scenarios, during ages 2-6) considering this appear reasonable. For the soil dermal contact pathway, key uncertain parameters include the soil adherence (0.2 mg/cm2-event for the central residential scenario and 1.0 mg/cm2-event for the high end farming scenario) and the absorption fraction (0.03 for dioxin-like compounds).

A major area of uncertainty for both pathways is the estimation of soil concentrations where the source of contamination is located distant from the site of exposure. For this assessment, this includes the off-site soil source category and the stack emission source category. Results from sensitivity analysis exercises for the erosion algorithm suggests that the 0.28 ppb soil concentration (within a 5-cm layer) used for soil ingestion and dermal contact, and which resulted from the 1 ppb nearby (150 m) soil contaminated site, may be high.

Specifically, when all parameters for the erosion algorithm remained constant except the dissipation half-life, which initially was 0.0693 yr-1 (half-life of 10 years) and then was reduced by a factor of ten to 0.00693 yr-1 (half-life of 100 years), the soil concentration 150 meters away at the site of exposure increased to slightly above 1.00 ppb. While dissipation of surface residues which have arrived at an exposure site from a distant source is an appropriate assumption, the outcome of a higher soil concentration 150 meters from a site of soil contamination when no dissipation is assumed (albiet assuming infinite time such that a steady state is reached) is questionable.

Key uncertain parameters identified include the dissipation rate (0.0693 yr-1), the mixing depth (5 cm), and the use of an enrichment ratio (equal to 3.0) which increases the concentration of dioxin-like compound on eroded soil relative to in-situ soil. This latter parameter was speculated to the one most likely to be inaccurate for evaluation of off-site soil impacts. Its assignment was not based on data specific to dioxin-like compounds, but rather to general literature data on enrichment ratios for soil nutrients and pesticides showing a range of between 1 and 5.

On the other hand, support for an enrichment ratio of 3.00 came in a data set including background soil and concurrent bottom sediment data in receiving water bodies in Connecticut (see Table III.5 for a summary of this data set). There, the ratio of sediment concentrations of 2,3,7,8-TCDD to soil concentrations was 2.8, suggesting that bottom sediments are enriched in comparison to surface soils. The model for bottom sediment impact from watershed soils includes the enrichment ratio, which was set at 3.00, and the demonstration scenarios did show a sediment:soil ratio of 2.8, like the observed data.

An uncertain outcome was also identified for the particle deposition algorithm used for the stack emission source category. An analysis suggests that the soil concentration in a 1-cm layer resulting from depositing particles may be underestimated by about an order of magnitude. The pertinent analysis for this observation came from the air-to-beef food chain model validation exercise conducted for dioxin-like compounds (further details of this exercise are found in Table III-5). There, a rural air profile of dioxin-like compounds were deposited onto soils, and the resulting concentrations of dioxin-like compounds were compared against observations from four United States reports on soil concentrations in rural areas.

Generally, the model underpredicted soil concentrations by about an order of magnitude. Suggested causes for this underprediction include: 1) the model does not consider vapor phase transfers to soils, 2) the model does not consider detritus contributions to soil, and 3) the half-life of 10 years may not be long enough for dioxin-like compounds.

In summary, principally identified uncertain parameters for the algorithms transporting eroding soil and depositing particles include: the mixing zone depth for untilled situation of 1 and 5 cm, the dissipation half-life of 10 years, the lack of consideration of vapor phase depositions and detritus additions to soils, and the use of an enrichment ratio for eroded soil of 3.0.

3. Ingestion of Water:

A comparison of alternate modeling approaches for estimating water concentrations showed similar results to the models adopted for this assessment. There also does not appear to be a wide range of possible values for water ingestion rate (1.4 L/day for central scenarios and 2.0 L/day for high end scenarios) and contact fraction (0.75 for central scenarios and 0.90 for high end scenarios), and these are not expected to introduce significant uncertainty into water ingestion exposure estimates.

4. Inhalation:

The inhalation rate assumed for both central and high end scenarios was 20 m3/day. The distinction in the scenarios was in the contact fractions: central scenarios assumed a contact fraction of 0.75 and high end scenarios had a 0.90 contact fraction. These fractions correspond to time at the home environment. These fractions and the inhalation rate are not expected to add significant uncertainty in inhalation exposure estimates.

Sensitivity analysis showed air concentrations resulting from soil emissions to be sensitive to Koc and H, and also to key source strength and delivery terms such as areas of contamination and wind speed. Assuming these non-chemical specific parameters can be known with reasonable certainty for site-specific applications, the most uncertainty lies with chemical specific data.

Alternate approaches for volatilization and air dispersion tested included the volatilization approach developed by Jury, et al. (1983) and the box model for dispersion calculations. The Jury model predicted about 1/3 as much volatilization flux (given the selection of parameters, made equal to or most analogous to the models of this assessment) as the Hwang, et al. (1986) model of this assessment. The box model predicted about 6 times higher air concentrations than the near-field dispersion approach of this assessment. This reasonable comparison lends some credibility to the models selected.

Approaches to estimate particulate phase concentrations are empirical and based on field data. They are based on highly erodible soils but are specific to inhalable size particles, those less than 10 m m. As such, they may overestimate inhalation exposures, but may underestimate the total reservoir of particulates, which becomes critical for the particle deposition to vegetation algorithms.

Another area of uncertainty is the assumption that volatilized contaminants do not become sorbed to airbone particles - this is also critical because vapor phase transfers dominate plant concentration estimation. A final key area of uncertainty is that transported contaminants from a contaminated to an exposure site via erosion are assumed not to volatilize or resuspend at the exposure site or from soils between the contaminated and the exposure site - air borne exposure site concentrations may be underestimated as a result.

5. Fruit and Vegetable Ingestion:

All ingestion parameters assumed are evaluated as reasonable for general exposure to broad categories of fruits and vegetables.However, great variability is expected if using these procedures on a specific site where home gardening practices can be more precisely ascertained.

Concepts of below and above ground vegetations were developed to accomodate soil to root algorithms and soil to air to vegetation algorithms. Protected vegetations - those with outer inedible protections such as citrus or corn - were assumed not to be impacted by dioxin-like compounds.

A key assumption in the vegetation algorithm, that dioxin-like compounds do not translocate from root to shoot, was verified by two experiments. Vapor-phase contributions to vegetation dominated the contaminated soil and stack emission source categories, with one exception. Particle depositions were more important for above ground fruit/vegetable concentrations for the stack emission source.

A critical empirical parameter was the above and below ground correction factors, VGag and VGbg, both set at 0.01 for fruits and vegetables. These factors were justified for dioxins based on the fact that the experiments for derivation of the below ground empirical transfer factor and the above ground empirical transfer factor were conducted with thin barley roots and azalea leafs, respectively.

Whole plant concentrations for these vegetations are likely to be much higher than whole plant concentrations of bulky fruits and vegetables; hence the introduction of the VG parameters. VG for grass was set at 1.00, which assumes that grass leaves and azalea leaves are analagous with regard to vegetative bulk. VG for cattle feed was set at 0.50, which assumes that some cattle feed is leafy (hay), while some is bulky (corn silage).

A different assumption for VG of fruits and vegetables, such as 0.10, would increase estimated concentrations and perhaps make plant:soil concentration ratios more in line with literature values (see Table III-5). Experimental evidence that a VGag for vapor transfers of dioxin-like compounds is justified came in a recent study by McCrady (1994). McCrady experimentally determined uptake rate constants, termed k1, for vapor phase 2,3,7,8-TCDD uptake into several vegetations including kale, grass, pepper, spruce needles, apple, tomato, and azalea leaves.

The uptake rate for an apple divided by the uptake rate for the grass leaf was 0.02 (where uptake rates were from air to whole vegetation on a dry weight basis). For the tomato and pepper, the same ratios were 0.03 and 0.08. The VGag was 0.01 for fruits and vegetables in this assessment. McCrady (1994) then went on to normalize his uptake rates on a surface area basis instead of a mass basis; i.e., air to vegetative surface area instead of air to vegetative mass.

Then, the uptake rates were substantially more similar, with the ratio of the apple uptake rate to the grass being 1.6 instead of 0.02; i.e., the apple uptake rate was 1.6 times higher than that of grass, instead of 1/50 as much when estimated on an air to dry weight mass basis. The ratios for tomato and pepper were 1.2 and 2.2, respectively. In his article, McCrady (1994) concludes, "The results of our experiments have demonstrated that the exposed surface area of plant tissue is an important consideration when estimating the uptake of 2,3,7,8-TCDD from airborne sources of vapor-phase 2,3,7,8-TCDD.

The surface area to volume ratio (or surface area to fresh weight ratio) of different plant species can be used to normalize uptake rate constants for different plant species." McCrady does caution, however, that uptake rates are only part of the bioconcentration factor estimation, and is unsure of the impact of surface area and volume differences on the elimination phase constant, k2 (personnal communication, J. McCrady, US EPA, ERL-Corvallis, Corvallis, OR 97333).

Still, his recent experiments do appear to justify the use of a VG parameter since the air-to-leaf transfer parameter was developed on an air-to-whole-plant-mass basis, and his results are consistent with the assignment of 0.01 for fruits and vegetables.

An uncertain experimentally derived empirical factor described the transfer of compounds from soil to below ground vegetables, the Root Concentration Factor, RCF. An analagous uncertain parameter describes the transfer of vapor-phase dioxin-like compounds from air to above ground vegetations, the air-to-leaf transfer factor, Bvpa.

Both of these parameters are estimated as functions of the contaminant properties; both used contaminant octanol water partition coefficient, Kow, and the Bvpa also used contaminant-specific Henry's Constant, H. The Bvpa was developed in a series of experiments by Bacci, et al. (1990, 1992) using 14 different organic contaminants and azalea leaves. Adjustments to the Bvpa as formulated by Bacci were suggested by the experiments on the transfer of 2,3,7,8-TCDD to grass leaves by McCrady and Maggard (1993).

The adjustments dealt with the impact of photodegradation, which was not considered in the experimental design of Bacci, and in the different plant species used by McCrady and Maggard. Those adjustments were made for the dioxin-like compounds in this assessment. The range of log Kow for 2,3,7,8-TCDD found in the literature was 6.15 to 8.5.

An alternate value of log Kow for 2,3,7,8-TCDD would more likely be higher than lower, given the selected value of 6.64. Increasing log Kow tends to decrease below ground vegetation, by as much as an order of magnitude, while increasing above ground vegetation by as much as an order of magnitude.

6. Ingestion of Fish:

The key exposure parameter for this pathway was the fish ingestion rate. The rates assumed in the demonstration scenarios were low in comparison to estimates given for subsistence fisherman or others who live near large water bodies where fish are commercially caught.

The justification for the lower ingestion rate for demonstration purposes was that the setting demonstrated was described as rural, containing farms and non-farm residences, where the emphasis is on agriculture. A relatively small watershed with a small impacted water body was assumed. Daily ingestion rates of 1.2 (central) and 4.1 (high end) g/day were assumed, based on an assumption of 3 fish meals per year (150 g/fish meal) obtained from the water body for the central scenario and 10 fish meals per year for the high end scenario.

Other fish ingestion rates that can be considered for exposure assessments include: 6.5 g/day characterized as a national average ingestion rate for freshwater and estuarine fish and shellfish (EPA, 1984), and 30 and 140 g/day, which are described as 50th and 90th percentile rates for recreational fisherman in areas where large water bodies are present (EPA, 1989).

Other models for estimating fish concentration based on water column concentrations, rather than suspended sediment concentrations, were described in EPA (1993) and demonstrated in this assessment. Results indicated that the water column approaches would predict similar whole fish concentrations compared with the sediment concentration approaches of this assessment.

However, the various models would respond differently to changes in model parameters. For example, a bioaccumulation parameter based on whole water concentration (total contaminant, the sum of sorbed and dissolved amounts, divided by water volume) will be mostly insensitive to changes in organic carbon content of sediments. In contrast, this is a critical parameter for bioaccumulation parameters which are based on sediment concentrations (as in this assessment) or dissolved-phase water column concentrations.

A key uncertain parameter for estimating fish tissue concentrations is the Biota Sediment Accumulation Factor, or BSAF, and the Biota Suspended Sediment Accumulation Factor, or BSSAF. A range of 0.03 to 0.30 for 2,3,7,8-TCDD is hypothesized for column feeding fish, while the Connecticut data (CDEP, 1992) and some other data on bottom feeding fish indicate higher BSAFs ranging up to 0.86 for 2,3,7,8-TCDD.

A value of 0.09 for 2,3,7,8-TCDD for BSAF and BSSAF is used in this assessments. Data is scarce for BSAF and BSSAF for other dioxin-like compounds, although available data does suggest that these parameter values decrease as the degree of chlorination increases. A key parameter is the fish lipid content, which can vary from below 0.05 to above 0.20. The model estimates a fish lipid concentration.

Multiplying fish lipid concentration by fish lipid content arrives at a whole fish concentration or an edible fish concentration, depending on the user's assignment and characterization of the fish lipid content variable. For this assignment, the fish lipid content was assigned a value of 0.07 for the demonstration scenarios, based on lipid content of fish in EPA's Lake Ontario study (EPA, 1990a).

7. Beef and Milk Ingestion:

The rates of beef and milk fat ingestion are 22 and 10.5 g/day, respectively. The median whole beef and whole milk ingestion rates are given as 100 and 300 g/day, respectively (EPA, 1989), and these were assumed for the demonstration scenarios. Beef fat and milk fat contents are assumed to be 22% and 3.5%, respectively. Only the high end demonstration scenarios included beef and milk ingestion pathways.

These scenarios were farm settings, and the assumption was that farming families would obtain a portion of their ingestion of these foods would come from home produced beef and milk.

The assumptions for contact fractions for beef and milk (fractions of their total consumption that comes from home supplies) was 0.44 and 0.40, respectively. These were average consumption fractions for farming families, whether or not the farm families home consumed, and were developed from a USDA (1966) survey of farming families.

Since exposure estimates from these pathways are linearly related to ingestion rate and contact fraction, these are critical exposure parameters for site specific applications. Comparison with earlier modeling approaches showed that the current approach to estimating beef and milk concentrations is the same as earlier approaches, although mathematically formulated differently.

Earlier approaches also estimated cattle dose of 2,3,7,8-TCDD from contaminated air (directly) and contaminated ground water - these earlier estimations showed these contributions to be minimal, and they were not considered in this assessment. Early efforts in the literature did not consider vapor transfers to vegetations; one later assessment did include vapor transfers, and a key result in that assessment, as well as this one, is that vapor transfers are critical for beef impacts.

Finally, earlier assessments considered the practice of fattening beef cattle prior to slaughter by feeding them residue-free grains. These efforts estimated over a 50% reduction in beef concentration due to residue degradation or elimination and/or dilution with increases in body fat. The demonstrations scenarios in this assessment did not consider this practice. However, this practice was considered in the air-to-beef food chain validation exercise. There, a 50% reduction in beef concentrations due to feedlot fattening was assumed.

Key uncertain and variable parameters for beef/milk concentrations include:

1) the assumptions concerning vapor/particle partitioning for the stack emission source category,
2) the air-to-leaf transfer parameter, Bvpa, for vapor phase contaminants,
3) beef cattle exposure assumptions,
4) the weathering factor for particles depositing on vegetations which cattle consume, and
5) uncertainties as discussed above for air to soil algorithms and soil to air algorithms.

REFERENCES FOR VOLUME III
  • Bacci, E.; Calamari, D.; Gaggi, C.; Vighi, M. (1990) Bioconcentration of Organic Chemical Vapors in Plant Leaves: Experimental Measurements and Correlation. Environ. Sci. Technol. 24: 885-889.
  • Bacci, E.; M.J. Cerejeira; C. Gaggi; G. Chemello; D. Calamari; M. Vighi (1992) Chlorinated Dioxins: Volatilization from Soils and Bioconcentration in Plant Leaves. Bull. of Env. Cont. and Tox. 48(3):401-408.
  • Bidleman, T.F. (1988) Atmospheric processes. Wet and dry depostion of organic compounds are controlled by their vapor-particle partitioning. Environ. Sci. Techol., 22:4, pp 361-367.
  • Briggs, G.A. (1975). Plume rise predictions. In: Lectures on air pollution and environmental impact analyses, American Meteorology Society.
  • Briggs, G.A. (1979). Plume rise. USAEC Critical Review Series. NTIS publication no. TID-25075.
  • CARB (1986) Subroutines for calculating dry depostion velocities using Sehmel's curves. Prepared by Bart Croes, California Air Resources Board.
  • CDEP (1992) Data on the Connecticut Department of Environmental Protection (CDEP) program to monitor soil, sediment, and fish in the vicinity of Resource Recovery Facilities. Data supplied by C. Fredette, CDEP, 165 Capitol Ave, Hartford, CT, 06106.
  • Dumbauld, R.K.; Rafferty, J.E.; Cramer, H.E. (1976) Dispersion deposition from aerial spray releases. Preprint volume for the Third Symposium on Atmospheric Diffusion and Air Quality. American Meteorological Society.
  • Fries, G.F. 1985. Bioavailability of soil-borne polybrominated biphenyls ingested by farm animals. Journal of Toxicology and Environmental Health 16: 565-579.
  • Fries, G.F.; Paustenbach, D.J. (1990) Evaluation of Potential Transmission of 2,3,7,8-Tetrachlorodibenzo-p-dioxin-Contaminated Incinerator Emissions to Humans Via Foods. J. Toxicol. Environ. Health 29: 1-43.
  • Hwang, S.T.; Falco, J.W.; Nauman, C.H. (1986) Development of Advisory Levels for Polychlorinated Biphenyls (PCBs) Cleanup. Exposure Assessment Group, Office of Research and Development, U.S. Environmental Protection Agency. EPA/600/6-86/002.
  • Jury, W.A.; Spencer, W.F.; Farmer, W.J. (1983) Behavior assessment model for trace organics in soil I. Model description. Journal of Environmental Quality 4:558-564.
  • McCrady, J.K.; Maggard, S.P. (1993) Uptake and photodegradation of 2,3,7,8-tetrachlorodibenzo-p-dioxin sorbed to grass foliage. Env. Sci. Technol. 27:343-350.
  • McCrady, J.K. (1994) Vapor-phase 2,3,7,8-TCDD sorption to plant foliage - a species comparison. Chemosphere 28(1):207-216.
  • McLachlan, M.S.; Thoma, H.; Reissinger, M.; Hutzinger, O. (1990) PCDD/F in an agricultural food chain. Part I: PCDD/F mass balance of a lactating cow. Chemosphere 20:1013-1020.
  • Paustenbach, D.; Wenning, R.; Lau, V.; Harrington, N.; Rennix, D.; Parsons, A. (1992)  Recent developments on hazards posed by 2378-TCDD in soil: Implications for setting risk-based cleanup goals at residential and industrial sites. J. Tox. Env. Health, 36:103-149.
  • Radke, L.F.; Hobbs, P.V.; Eltgroth, M.W. (1980) Scavenging of aerosol particles by precipitation. Journal of Applied Meteorology 19:715-722.
  • Rao, K.S. (1981) Analytical solutions of a gradient-trasnfer model for plume deposition and sedimentation. NOAA Technical memorandum. ERL ARL-109.
  • Rao, K.S.; Sutterfield, L. (1982) MPTER-DS. The MPTER model including deposition and sedimentation. U.S. EPA, Research Triangle Part, NC. EPA 600/8-82/024.
  • Reed, L.W.; Hunt, G.T.; Maisel, B.E.; Hoyt, M.; Keefe, D.; Hackney, P. (1990) Baseine assessment of PCDDs/PCDFs in the vicinity of the Elk River, Minnesota generating station. Chemosphere 21:159-171.
  • Sehmel, G.A. (1980) Particle and gas dry depostion: A review. Atmospheric Environ. 14, pp 983-1011.
  • Seinfeld, J.H. (1986) Atmospheric chemistry and physics of air pollution. New York, NY., John Wiley and Sons.
  • Turner, D.B. (1986) Fortran computer code/user's guide for COMPLEX I Version 86064: An air quality dispersion model in section 4. Additional models for regulatory use. Source file 31 contained in UNAMAP (VERSION 6). National Techical Information Service, Sprinfield, VA. NTIS PB86-222361/AS.
  • U.S. Department of Agriculture. (1966) Household food consumption survey 1965-1966. Report 12. Food Consumption of households in the U.S., Seasons and years 1965-1966. United States Department of Agriculture, Washington, D.C. U.S. Government Printing Office.
  • U.S. Environmental Protection Agency. (1984) Ambient water quality criteria document for 2,3,7,8-tetrachlorodibenzo-p-dioxin. Office of Water Regulations and Standards, Washington, D.C. EPA-440/5-84-007.
  • U.S. Environmental Protection Agency. (1986). Industrial source complex (ISC) dispersion model user's guide-second edition. Office of Air Quality Planning and Standards, Research Triangle Park, NC, EPA-450/4-86-005a.
  • U.S. Environmental Protection Agency. (1989) Exposure Factors Handbook. Office of Health and Environmental Assessment. EPA/600/8-89/043. July, 1989.
  • U.S. Environmental Protection Agency. (1990a) Lake Ontario TCDD Bioaccumulation Study Final Report. Cooperative study including US EPA, New York State Department of Environmental Conservation, New York State Department of Health, and Occidental Chemical Corporation. May, 1990.
  • U.S. Environmental Protection Agency. (1990b) Assessment of Risks from Exposure of Humans, Terrestrial and Avian Wildlife, and Aquatic Life to Dioxins and Furans from Disposal and Use of Sludge from Bleached Kraft and Sulfite Pulp and Paper Mills. Office of Toxic Substances and Office of Solid Waste, EPA 560/5-90-013. July, 1990.
  • U.S. Environmental Protection Agency. (1990c) USEPA/Paper Industry Cooperative Dioxin Study "The 104 Mill Study" Summary Report and USEPA/Paper Industry Cooperative Dioxin Study "The 104 Mill Study" Statistical Findings and Analyses Office of Water Regulations and Standards, July 13, 1990.
  • U.S. Environmental Protection Agency. (1990d) Methodology for Assessing Health Risks Associated with Indirect Exposure to Combustor Emissions. Interim Final Office of Health and Environmental Assessment. EPA/600/6-90/003. January, 1990.
  • U.S. Environmental Protection Agency. (1991) A Methodology for Estimating Population Exposures from the Consumption of Chemically Contaminated Fish. Prepared by Tetra Tech, Inc., Fairfax VA. for Office of Policy, Planning, and Evaluation and Office of Research and Development, US EPA. EPA/600/9-91/017.
  • U.S. Environmental Protection Agency. (1992a) Guidelines for exposure assessment. Office of Health and Environmental Assessment, Washington, DC. EPA/600-Z-92/001. published in Federal Register, May 29, 1992, p. 22888-22938.
  • U.S. Environmental Protection Agency. (1992b) National Study of Chemical Residues in Fish. Volumes I and II. Office of Science and Technology EPA 823-R-92-008a & 008b. September, 1992.
  • U.S. Environmental Protection Agency. (1992c) Dermal Exposure Assessment: Principals and Applications. Office of Health and Environmental Assessment. EPA/600/8-91/011B.
  • U.S. Environmental Protection Agency. (1993)