An adjustment is made to the sediment delivery ratio, SD₁, considering the size discrepancies between the contaminated site and the exposure site. For example, if the contaminated site is larger than the exposure site, then the amount of eroded soil delivered 150 meters downgradient would not all mix with soil at the exposure site. On the other hand, if the contaminated site were smaller than the exposure site, than the full amount of eroded soil delivered 150 meters downgradient would be contained within the exposure site. A simple correction factor, equaling the ratio of a side length of the exposure site (assumed square-shaped) and a side length of the contaminated site size (also assumed square shaped), is used to adjust the sediment delivery ratio:

where:

SD_1a = adjusted sediment delivery ratio corresponding to SD₁, unitless

SD₁ = sediment delivery ratio reducing the amount eroding from the contaminated site to be delivered to the exposure site, unitless

CF = A_ES^0.5/A_LS^0.5 if A_ES < A_LS

= 1 if A_ES > A_LS

A_ES = area of exposure site, m²

A_LS = area of contaminated site, m²

Similar considerations are pertinent to the land area between the contaminated and exposure site. Remember that the algorithm assumed that some "clean" (D₂) and some "contaminated" soil (D₁) erodes onto the exposure site, and that a similar amount of soil entering the exposure site (R, which equals D₁ + D₂) leaves the exposure site so as to maintain a mass balance. The amount of clean soil eroding from upgradient sources mixing with exposure site soil can be larger than the amount of contaminated soil if the exposure site is larger than the contaminated site. If the exposure site is smaller than the contaminated, and similar to the solution for SD_1a above, then only the small corridor defined by the size of the exposure site contributes clean soil. Either way (i.e., the exposure site is larger or smaller than the contaminated site), the size of the land area contributing clean soil is defined by the size of the exposure site. A_BLE can be estimated as the product of the distance between the exposure and contaminated site, and the side length of the exposure site:

where:

A_BLE = land area between contaminated and exposure site, m²

DL = distance from landfill to exposure site, m

SL = side length of exposure site, m

= (A_ES)^0.5

A_ES = area of exposure site, m².

Estimating the dispersion and resulting exposure site concentrations of air-borne contaminants, originating at the site of contamination in a vapor or a particle phase, requires a different solution for the off-site as compared to the on-site situation. A simplified solution, given as a virtual point source model, can be found in Turner (1970). This model approximates the dispersion that occurs from an area source by using an imaginary point source. This point is located upwind of the actual source at a distance calculated to create a lateral dispersion at the site equal to its width:

where:

C_air = concentration of contaminant in air, µg/m³

FLUX = average contaminant flux rate, g/cm²-s

A_s = area of contaminated site, m²

FREQ = frequency wind blows from source to receptor, unitless

VD = virtual distance, source center to receptor, m

S_z = vertical dispersion coefficient, m

U_m = average wind speed, m/s

10¹⁰ = converts g/cm² to µg/m².

The term, FREQ, has been added to this equation to appropriately account for changing wind directions, and hence, obtain a more accurate annual average air concentration. The vertical dispersion, S_z, is estimated as an empirical function of the distance from the source center to receptor:

where:

S_z = vertical dispersion coefficient, m

X = actual distance from source center to receptor, m.

The virtual distance, VD, is an empirical function of the width of the contaminated area and the actual distance from source center to receptor:

where:

VD = virtual distance, source center to receptor, m

a = width of contaminated area perpendicular to wind direction - defined previously as side length for assumed square-shaped contaminated area, m

X = actual distance from source center to receptor, m.

Prior guidance on windspeed (Section 4.3.2) indicated windspeeds ranged from 2.8 to 6.3 m/sec, and suggested a mid-range of 4.0 m/sec in the absence of better information. Where the wind blows from all directions equally, then it will blow from one compass sector about 15% of the time. On these bases, a windspeed of 4.0 m/sec and a FREQ of 0.15 were used in the example scenarios in Chapter 5. In most places, however, wind direction is much less variable, and the appropriate value is best determined with site specific information.

This section provides background information on specific sites of soil contamination which have been studied for the presence and impact of dioxin-like compounds. These include landfills used for disposal of ash from municipal waste combustion facilities, the disposal of sludge from pulp and paper mills, and sites of soil contamination typified by the sites monitored in the National Dioxin Study (in many cases, Superfund sites or sites that were in some stage of being considered for inclusion on the NPL list at the time of the study; EPA, 1987). Discussion of these particular sites does not imply that they represent the bulk of such sites nationally, or that they are discussed here based on any critical environmental or exposure rational. They are discussed because they present unique issues for emissions and fate and transport of dioxin-like compounds from sites of soil contamination, and because they have been studied. Issues discussed below are pertinent for other types of off-site soil contamination sites.

Particular issues regarding landfills receiving ash include: the impact of soil cover on releases, the ash concentrations, the size of such landfills, the quantity of ash generated by incinerators, and the fugitive emissions that result from ash management. Key sources providing information for this section include a methodology document describing approaches to estimating environmental releases and exposures to ash (EPA, 1991), and a contractor report applying these types of methodologies using site-specific data from several ash landfills (MRI, 1990). Each of the identified topics will be discussed in turn.

• Landfill Cover: Whether or not ash is covered once it is disposed of at the landfill is critical in determining releases and subsequent exposures. Currently, practices at operating landfills vary from no coverage after disposal on active portions of the landfill to daily coverage of disposed ash. MRI (1990) visited six facilities disposing ash, including ash monofills and municipal solid waste landfills. In one of the facilities, an ash monofill located at the site of the combustor, the disposal area encompassed 15 acres and did not use daily cover until final elevation was reached. At that time, a clean cover of 2 feet of soil would be applied. At a second facility, located at the site of the combustor but landfilling municipal solid waste as well as ash, ash was used for different purposes, including a subbase roadbed material, as soil substitute for earth work, and as a daily cover for MSW receipts. An assumption of bare surfaces (i.e., no vegetation) during the period of landfilling activity, with concentrations of dioxin-like compounds equal to concentrations in the ash would appear to be appropriate assumptions for practices at these two landfills.

Where daily cover is employed, however, appropriate assumptions are not straightforward. Of the remaining four sites studied by MRI, two employed daily covers ("clean cover material" in one case and a "HDPE liner" in the other, sic), and daily coverage practices were not discussed for two sites. Approaches described for airborne emissions and erosion losses would have to be modified when daily cover is applied. First, losses of contaminants via overland soil or wind erosion could not be expected to occur when cover (soil cover or otherwise) is in place, although the active part of the landfill would be subject to erosion during an operating day. Even in that case, however, site-specific practices might include little or no ash disposal during periods of soil-erosion-producing storms. Depending on site-specific practices, one might estimate annual erosion losses using methodologies described in this assessment, and then empirically reduce erosions losses based on these practices and scientific judgement.

Air emissions from active portions of the landfill, as in wind erosion and volatilization, also are obviously impacted by cover practices. These emissions would occur during the actual disposal. Wind erosion and volatilization fluxes could be estimated as given in earlier sections, and then reduced by two-thirds, which might correspond to an assumption of disposal during 1/3 of a day or a year, etc. When covered by soil or a synthetic cover, wind erosion losses would not occur. However, buried residues may diffuse through layers of clean soil and be released via volatilization.

Estimates of volatilization release via diffusion through clean cover have been made. A rigorous approach for such estimates is detailed in Hwang, et al. (1986). Use of this approach requires a computer to iteratively solve a partial differential equation, expressed in terms of a Fourier series. It can be shown, with these equations, that the vapor emission rate through such a cover will not reach steady state for hundreds of years. Hwang's approach was applied to an earlier assessment for 2,3,7,8-TCDD (EPA, 1988b). Calculations were performed for 2,3,7,8-TCDD contamination with a thickness of contamination of 8 ft, and clean caps ranging from 10 to 25 cm. The results of this exercise suggest that the average emission rate of a 70-year period are 1/4 to 1/5 of what they would be without the cap. Based on this exercise, a simple assumption might be made that a clean cap will reduce the average emission rate calculated without a clean cap by 80%. However, these results are not consistent with those described in Jury, et al. (1990). The analytical solution developed by Jury was demonstrated on 35 organic compounds. One exercise conducted by Jury was to estimate the cover thickness required to restrict volatilization to less than 0.7% of the mass incorporated in soil. For 2,3,7,8-TCDD, the thickness was estimated at 0.7 cm for a sandy soil and 0.2 cm for a clayey soil. This appears to contradict the work of Hwang since it shows an essentially insignificant loss for a cap much less thick than the 10-25 cm cap in the exercises using Hwang's approach. However, Jury's approach allows for assumptions on degradation of the buried compound. For that exercise, Jury assumed that the half-life for 2,3,7,8-TCDD was 1 year. This is a very rapid degradation rate, given information that the dioxin-like compounds resist degradation, particularly when not exposed to sunlight. On the other hand, the Hwang model assumes no degradation loss, and as such, the generalization from his exercise might be an overestimate. Hwang's exercise might also have overestimated since it assumed a rather thick 8-ft layer of subsoil contamination. From these arguments, it would appear that neither exercise appropriately evaluated the difference in volatilization in a no cover versus a cover situation.

The above discussions concerned flux calculations when cover practices are used. One set of adjustments discussed reduced a total potential flux of volatilized or wind eroded losses based on a portion of the time that the ash would be uncovered. A second discussion indicated that some loss via volatilization might be modeled with a clean cap. In any case, it is clear that cover practices will reduce losses. Cover practices must be considered when evaluating the exposure to ash disposed of in landfills.

• Ash Concentrations: A key consideration, of course, in modeling transport of dioxin-like compounds from an ash landfill is the concentration on the ash. Ash concentrations of dioxin-like compounds have been found to vary widely, from non-detect (generally less than 0.1 ppb) to the hundreds and thousands of part per billion. Table 4-5 appears in EPA (1991) and summarizes concentrations of dioxin-like compounds and PCBs found in fly, bottom, and combined ash. These data are a summary of 19 references, ranging in publication date from 1974 to 1990. It should be noted that, except for 2,3,7,8-TCDD and 2,3,7,8-TCDF, results listed are for congener groupings defined by degree chlorination.

• Size of Landfill and Amount of Ash Landfilled: The size of the landfill and the amount of ash applied daily or over time are both required for estimating exposures nearby. These can both be obtained from site-specific observations. Amounts of daily disposed ash are required to estimate fugitive particulate emissions, as will be discussed shortly. Amounts of daily or ultimate disposal are also tied to landfill size, or the portion of a landfill that is active on a daily basis. One common practice is to fill cells of a landfill one at a time, and once filled, to cover with a 2-ft (or so) layer of clean soil. The appropriate size in this case is the average size of a landfill cell. If daily coverage is applied, than the size for modeling purposes corresponds to the area over which daily coverage occurs. This can also vary depending on the depth of disposal during a day. A six-inch daily coverage, for example, would take twice as much space as a 1-ft depth of daily disposal. If the intent of a day's disposal is to cover over the entire area of an active cell, then depth of coverage need not be considered in determining landfill size.

Determination of landfill size (or the size of the active portion of the landfill) may be required in the absence of site-specific information, such as in the planning stages for a new incinerator. This is where details on landfill management need to be determined. One important detail, as already noted, is the amount of ash generated for daily disposal. Cook (1991) assumes that bottom and fly ash combined comprise about 11% of total receipts on a volume basis. However, a relationship between ash generated and solid waste

Table 4-5. Ranges of concentrations of PCDDs, PCDFs, and PCBs in municipal waste combustor ash (results in ng/g or ppb).

 

 

 

Constituent Fly Ash (ref) Combined Ash Bottom Ash

 

 

MCDD 2.0 ND NR

DCDD 0.4-200 ND-120 NR

T₃CDD 1.1-82 ND-33 NR

T₄CDD ND-250 0.14-14 <0.04-410

PCDD ND-722 0.07-50 ND-800

H₆CDD ND-5,565 0.07-78 ND-1,000

H₇CDD ND-3,030 0.07-120 ND-290

OCDD ND-3,152 0.07-89 ND-55

2,3,7,8-TCDD ND-330 0.02-0.78 <0.04-6.7

Total PCDD 5-10,883 6.2-350 ND-2,800

MCDF 41 1.1 NR

DCDF ND-90 ND-42 NR

T₃CDF 0.7-550 ND-14 NR

T₄CDF ND-410 2.3-9 10.1-350

PCDF ND-1800 1.6-37 0.07-430

H₆CDF Tr-2,353 1.2-35 ND-920

H₇CDF Tr-887 0.62-36 ND-210

OCDF ND-398 0.18-8.4 ND-11

2,3,7,8-TCDF 0.05-5.4 0.41-12 ND-13

Total PCDF 3.73-2,396 6.14-153.9 ND-1,600

Mono CB 0.29-9.5 ND ND-1.3

Di CB 0.13-9.9 0.126-1.35 ND-5.5

Tri CB ND-110 0.35-14.3 ND-80

Tetra CB 0.5-140 16.5 ND-47

Penta CB 0.87-225 ND ND-48

Hexa CB 0.45-65 ND-39 NR

Hepta CB ND-0.1 ND NR

Octa CB ND-1.2 ND NR

Nona CB ND ND NR

Deca CB ND ND NR

Total PCB ND-360 ND-32.15 ND-180

 

 

ND = not detected at the detection limit Tr: 0.01<Tr<0.1 ng/g NR = Not reported

Source: EPA (1991).

received by an incinerator on a mass basis is more useful for estimating daily disposal amounts. In a recent EPA (1990f) report on ash characterization, ash mass was estimated as an average of 29.5% of municipal solid waste received in five facilities studies, with a narrow range of 25-35%. This mass was estimated on a wet weight basis. Ash is wetted when exiting the incinerator, and water comprises 20-30% of the total weight at that point. If the ash is immediately trucked for landfill disposal, its total weight includes the weight of this quench water. Often ash is stored at the incinerator site in piles prior to disposal, that storage ranging from hours to days. In this circumstance, much of the quench water would have drained off or evaporated, and then the total weight hauled would be closer to a dry weight estimate. In summary, the amount of ash generated to be disposed of a daily basis can be estimated as: the daily receipt of municipal solid waste (tons) * a wet weight ash fraction (0.25-0.35) * a wet to dry weight conversion if appropriate (wet weight * 0.80, e.g.).

• Fugitive Particulate Emissions: Fugitive emissions can occur from the time ash exits the incinerator for temporary storage at the facility site (or immediate loading onto trucks for disposal) until ultimate disposal. Approaches to estimate fugitive releases from incinerator ash management are described in EPA (1991), and will be summarized here.

As noted, ash can be wet when exiting the quench tank. If stored at the facility site prior to disposal in a landfill, leaching from piles can occur. Because dioxin-like compounds are strongly hydrophobic, however, the impact of leaching is unlikely to occur much beyond the soil beneath and near the storage piles. If loaded onto trucks when very wet, leaking onto roadways may also occur. If these storage piles are left uncovered, they would of course be subject to erosion losses, which might move residues further from the piles than just leaching of water from the piles.

Of more concern than water-borne losses due to ash management are fugitive emissions of dry ash. Wind erosion, which can occur from open storage piles or uncovered portions of the landfill, is a fugitive emission that has been discussed for soil contamination. Specific practices in the management of ash can also result in fugitive emissions. Such practices include: 1) loading onto and dumping out of trucks, 2) truck transport from the incinerator facility to the landfill site, 3) truck or other traffic over paved or unpaved roadways at the incinerator site, at the landfill site, or other roadways containing contaminated dust, and 4) spreading and compacting of ash at the landfill site. A set of empirical emission factor equations for estimating fugitive particulate emissions, called "AP-42" equations, have been developed by EPA's Office of Air Quality Planning and Standards (EPA, 1985a; EPA, 1988a). Specifics on applying these equations for ash management are described in EPA (1991). An example of their application using site-specific information for ash management is detailed in MRI (1991). An abbreviated listing of emission factor equations that have been used in these two publications are:

• Vehicular traffic over unpaved roadways. Dust on the surfaces of roads, both unpaved and paved, can become suspended due to vehicular traffic. When these roadways are near ash storage piles or within the landfill, that dust can become contaminated. The emission factor equation for emissions from unpaved roadways is:

where:

E_up = emission flux for unpaved surfaces, kg/VKt (VKt equals vehicle kilometer traveled)

k_unp = particle size multiplier specific to the unpaved road emission flux equation, unitless

s = silt content of unpaved roadway, %

Vs = vehicle speed, km/hr

W = vehicle weight, kg

nw = number of wheels per vehicle, unitless

P = number of days with at least 0.254 mm (0.01 inch) precipitation per year, unitless.

• Emissions off trucks in transit. Although no emission factor equations have specifically been developed for trucks while in transit from the incinerator facility to the landfill, such emissions can occur if the ash is dry, and partially or completely uncovered. The following equation for estimating emissions from open storage piles has been suggested for use in estimating fugitive emissions from trucks in transit (EPA, 1991; the emission factor equation from EPA, 1985a). Note that use of this equation will require specific management assumptions in order to estimate the number of uncovered hectares per day: the number of trucks in use per day, the surface area of trucks, the percent of uncovered area if a tarpaulin is used, the moisture content of ash, and so on.

where:

E_t = particulates emitted from trucks in transit, kg/day/hectare

s = silt content material of ash, %

P = number of days with >0.25 mm precipitation per year

f = percentage of time that the unobstructed wind speed exceeds 5.4 m/s.

• Loading and unloading. The unloading operations at the disposal site may result in the release of fugitive dust. The following emission factor equation provides emission factors for kilograms of particulate emitted per megagram (metric ton, or 1000 kg) of soil loaded and unloaded:

where:

E_lu = emission factor for loading and unloading, kg fugitive dust/MT ash

k_unl = particle size multiplier, dimensionless

U_m = wind speed, m/s

M = material moisture content, %.

• Spreading and compacting of ash at the landfill. An emission factor specifically for ash spreading and compacting has not been developed. However, emission factor equations for similar applications have been applied for estimating fugitive emissions due to spreading and compacting. MRI (1990) used an AP-42 emission factor developed for dozer moving of overburden in western surface coal mines. Kellermeyer and Ziemer (1989) assumed that the spreading and compaction of ash was analogous to vehicular transport on unpaved surfaces, and used the emission factor for that process. A third possible assumption is that the processes of spreading and compacting are analogous to agricultural tillage. That emission factor equation for agricultural tillage is:

where:

E_at = emission factor for agricultural tillage, kg/ha

k_at = particle size multiplier, dimensionless

s = silt content, %.

When applying such equations, there are further key issues to consider. These include:

• Concentrations on fugitive ash emissions: When such an emission occurs from ash surfaces, such as from storage piles, off trucks in transit, in spreading and compacting, and so on, than there is a good argument to assume that such concentrations on such emissions are "enriched" in comparison to an ash average. The argument here is similar to the argument for enrichment assumed for eroded soils: processes resulting in fugitive air emissions favor lighter particles with more surface area and hence more sites for binding. No data could be found to assign a value to an ash enrichment ratio. MRI (1990) did, however, take data on municipal waste combustor facility roadway dust, and based on that data and statistical evaluations, speculated that fly ash constituted the principal source of lead and cadmium found on paved surfaces. Since fly ash is finer than bottom or combined ash, one hypothesis for this finding is that fugitive emissions from ash management at the combustor site transported these finer particles to roadway surfaces. This is not to imply, however, that concentrations in dust suspended from roadways due to traffic should be higher in concentration than concentrations in ash - this enrichment concept only applies to ash surfaces themselves. Rather, the concentration on roadway suspended dust should be lower than on the ash. This is because contaminated dust on roadways mixes with clean dust from other sources. As noted, MRI (1990) did take roadway dust samples, and their data appears to place such a dilution factor (concentration on roadway dust divided by concentration on ash) in the range of 0.1 to 0.3. Specifically, they took particulate samples from landfill haul routes while at the same time taking samples of incinerator ash being delivered for disposal the same day. Each paired sample (roadway particulate and ash), were measured for four metals: As, Cd, Cr, and Pb. Several paired samples were taken on both paved and unpaved haul routes. Ratios were then generated for roadway particulate metal concentrations over ash metal concentrations. Results were: As - paved and unpaved ratios were similar and consistently near 0.1 (roadside particulate concentrations of As were 10% of ash concentrations of As), Cd - paved and unpaved ratios were similar and ranged between 0.0 and 0.4, Cr - paved ratios ranged from 0.3 to 0.6, while unpaved had a wide range of 0.3 to 2.0, Pb - paved and unpaved ratios were similar between 0.0 and 0.2. For analogous situations - daily deliveries of contaminated ash - one might assume a dilution factor in the 0.1-0.2 range.

• Selection of values for emission factor equations: As noted, all these equations are empirical equations. They were developed from data on sites where such emissions occur, such as strip mining sites. EPA (1988a) describes the range of conditions over which such equations were developed. What is meant by "conditions" are such factors as the range of vehicle weights in the data set, the range in number of wheels on such vehicles, and so on. Application of these equations for situations not included within these ranges should be done cautiously. Very critical also is the selection of the particle size multiplier variable, k. These values range from about 0.10 to no higher than 1.0. Lower k values are used to estimate emissions of the smallest sized particles; generally particles less than 5 m m in diameter. Higher k values are used to estimate emissions of all sized particles less than a higher diameter, usually either 15 or 30 m m. If these equations are used to only estimate particulate inhalation exposures, than the k value corresponding to 10 m m sized particles, or inhalable sized particles, should be used. When used to estimate total emissions, than the highest k value listed should be used. Such estimations are appropriate when also evaluating impacts to off-site soils or vegetation.

• Controls for fugitive emissions: All these equations were developed when no fugitive emission controls were in place. Common controls for roadway dust suppression include wetting or use of a chemical dust suppressant. Ash transported in trucks is commonly wetted and/or a tarpaulin is used to control emissions off trucks. There is no guidance or data on the effectiveness of such controls, but they must be considered. In demonstrating these procedures, EPA (1991) assumed that controls on emissions resulted in 90% reductions in potential emissions. If a control is known to be in place and used on a regular basis, than this percent reduction is probably a reasonable starting assumption.

This discussion focuses on an assessment on the land application of sludge from bleached kraft and sulfite pulp and paper mills (EPA, 1990e). Focusing on this source of sludge does not imply that pulp and paper mills produce more sludge than other industries, or that sludge from pulp and paper mills contains more dioxin-like compounds than other sludges. However, it is known that dioxin-like compounds are found in pulp and paper mill sludges. Also, because of the 104-mill study in 1988, much information is available on the content and disposal of this sludge (further information on the 104-mill study can be found in EPA (1990c) and EPA (1990d)). Some of the issues briefly discussed below for pulp and paper mill sludges would also pertain to sludges containing dioxin-like compounds from other sources.

EPA (1990e) described frequency distributions of concentrations of 2,3,7,8-TCDD and 2,3,7,8-TCDF for 79 mills reporting this information and also broke out the data based on disposal option. Although EPA (1990e) used the disposal option breakout of concentrations in their assessment of the impacts of the various options, it is not felt that the disposal option of choice is based on concentration. Over all options, the median (50% percentile as given in EPA (1990e)) and maximum 2,3,7,8-TCDD concentrations found in sludges were 51 and 3800 ng/kg (ppt), respectively. The median and maximum 2,3,7,8-TCDF concentrations found were 158 and 17100 ppt.

Fate and transport for contaminants is sludge is dependent on disposal means. Of the approximate 2.5 million metric tons of pulp and paper mill sludge generated annually (as estimated in the 1988 104-mill study), five principal options for disposal were noted: landfilling (44% of all sludge disposed), surface impoundments (24%), land application (12%), incineration (12%), and distribution and marketing (8%). Impacts by incineration were not discussed in EPA (1990e) and are not discussed in this section. Key issues pertaining to each disposal issue are now discussed.

• Landfilling: The issue of coverage as discussed above for ash landfills is relevant for any landfill. However, fugitive particulate emissions during sludge handling and disposal is not an issue as it was for disposal of ash from incinerators due to the differences in moisture content. Sludge is much higher in moisture at the time it is disposed of in comparison to ash - with moisture contents as high as 90%.

• Surface Impoundments: It was assumed in EPA (1990e) that sludge disposed of in surface impoundments have a higher moisture content as compared to sludge disposed of in landfills. Surface impoundments were located at the mill site, explaining the assumption for a higher moisture content. A surface impoundment in the EPA (1990e) assessment was defined as a facility in which the sludges are stored or disposed on land without a cover layer of soil. For this type of management, soil cover would not be an issue. Concentrations would be those measured in the sludge. Also, vegetative cover would be expected to be minimal, which would influence parameters associated with soil erosion.

• Land Application: Twelve percent of all sludge produced annually was land applied. Four of the 104 mills applied the sludge to forest land, two mills land applied the sludge to agricultural land, and two mills used the sludge for abandoned mine reclamation. The high organic matter content (EPA (1990e) assumed a 25% organic carbon fraction in sludge) and high fraction of clay-sized particles make sludge an attractive soil amender. Sludge is either applied to the land surface with or without incorporation. When not incorporated, sludge can be assumed to replace surface soils and concentrations would be those in the sludge. When incorporated, soil concentrations can be estimated simply as (in mg/kg): (mass of contaminant added, mg)/(mass of sludge added, kg + mass of soil in mixing zone, kg). One key issue when incorporated is the number of years of such treatments. Most of the land application uses of paper and pulp mill sludges reported in EPA (1990e) made applications in only one year. As easily seen in the above suggested equation, higher concentrations result with more years of incorporation. The other key issue with incorporation, of course, is the depth of incorporation. For agricultural applications, the depth of incorporation assumed in EPA (1990e) was 15 cm, similar to the 20 cm incorporation assumed for home vegetable gardening in this assessment. For silvicultural uses, the assumption in EPA (1990e) was 2.5 cm, which corresponds to some but minimal mixing. For abandoned mine reclamation, the assumption was 0 cm incorporation. Routes of exposure might also vary from focuses in this document depending on land application choice. When applied to agricultural land, impacts to food crops would demand particular attention (the procedures in this assessment were demonstrated with home grown vegetables, although of course impacts to food crops are critical when agricultural field soils are impacted by dioxin-like compounds). When applied to forest land, ecological impacts might warrant particular attention, as was discussed and demonstrated in EPA (1990e). A final issue to consider when land applying sludge to land is a rate of dissipation/degradation of dioxin-like compounds. Landfills and surface impoundments have ongoing surface applications and over time, the total depth of applications in the range of meters, so an assumption of a constant source strength over a period of exposure, as was assumed in this assessment for soil contamination sources, is reasonable. However, if only a few centimeters of surface soil are impacted, which might be the case for single applications to land and/or surface applications with no incorporation, an assumption of dissipation may be warranted. EPA (1990e) assumed no degradation of 2,3,7,8-TCDD in their assessment of impacts from land applications.

• Distribution and Marketing Uses: The volume of sludge distributed and marketed was approximately 8% of the total amount of sludge generated for the 104-mill study. For this use, sludge was composted and then sold as a soil amendment in residential, agricultural, and commercial settings. More attention to the dermal contact pathway appears appropriate for this usage. Site-specific factors, and the values for these factors used in EPA (1990e), include: 1) depth of incorporation - 0, 15 and 25 cm in assumptions characterized as high, best, and low estimates, 2) garden size - 0.016 and 0.022 hectares characterized as low/best estimate and high, and referencing a national gardening survey, 3) rate of application - between 5 and 20 dry metric tons per hectare references a USDA publication on use of sewage sludge compost for soil improvement and plant growth, and 4) years of using such compost - 20 without specific reference. The years of application is needed for estimating soil concentrations during and after the period of exposure, using a simple ratio as discussed above in land application.

The National Dioxin Study (EPA, 1987a) focused on sites of known or suspected contamination of soil by 2,3,7,8-TCDD. There were 7 "Tiers" of investigation, with roughly decreasing expectations of finding 2,3,7,8-TCDD. Tiers 1 and 2 included 2,4,5-TCP production and associated disposal sites (Tier 1) and sites where 2,4,5-TCP was used as a precursor in the manufacture of pesticidal products and associated disposal sites (Tier 2). These tiers had the highest expectation for finding 2,3,7,8-TCDD. There were originally thought to be 450 sites that would fall in Tiers 1 and 2, but after investigation, only 100 sites were included for study. Some were downgraded into Tier 3. Of the 100 sites studied, 20 were on or were proposed for inclusion in the Superfund National Priorities List. Tiers 3 and 5 were associated with 2,4,5-TCP formulation (Tier 3) and use (Tier 5). Tier 6 were organic chemical or pesticide manufacturing facilities were 2,3,7,8-TCDD was suspected of being present. Tier 4 included combustion sources and are not discussed further in this section. Tier 7, basically an examination of background areas, are also not discussed here.

Issues that are identified as important in fate and transport modeling for this subcategory of off-site sources include concentrations, the possibility of ground water contamination, and site-specific characterization. These are discussed in turn.

• Concentrations: Only 11 of the 100 Tier 1 and Tier 2 sites were eventually classified as requiring "no further action" because 2,3,7,8-TCDD soil concentrations were very low, < 1 ppb, or not detected (with detection limits generally at 1.00 ppb). Where it was detected, a general trend was to find very high concentrations where 2,4,5-TCP production wastes were stored or disposed of, with much lower concentrations at soils near these particular areas. At hot spots, concentrations were as high as 2,000 parts per million, but generally soil concentrations were in the parts per billion. It was this parts per billion generalization that led to the assignment of a 1 ppb soil concentration for the demonstration of the off-site source category in Chapter 5. There were findings in the low ppb range for Tiers 3, 5, and 6, but at much lower frequency and no findings higher than the tens of ppb range. For exposure assessments, the characterization of soil concentrations in a site containing hot spots has to be carefully considered. For site evaluations and proposed options for remediation, an areally weighted average might be considered, although this could dilute loss estimates depending on the area chosen - choosing a large area corresponding to property lines might, for example, lead to an "average" concentration orders of magnitude lower than concentrations found in hot spots. One approach which should be considered is a "hot spot" impact compared to an areally averaged impact. It should also be remembered that removal of highly contaminated soils is a common practice and another option for evaluation would be a concentration assuming hot spots are removed.

• Potential for Ground Water Contamination: PCBs have been found in ground water in sites associated with dielectric fluids of transformers. Oils can migrate through soils as a separate immiscible phase and reach ground water, which has been the common explanation for PCB impacts to ground water. Ground water contamination by 2,3,7,8-TCDD has very rarely been found in ground water, although it has been released to the environment in an oil matrix. The Times Beach area of Missouri is the principal example of this release, where waste oils containing 2,3,7,8-TCDD were used for dust control. Ground water sampling did occur in many of the Tier 1 and 2 National Dioxin Study sites, but the results were mostly non-detects. One occurrence at 0.18 ppt was noted for the Hyde Park site of Hooker Chemical in Niagara, NY, and a high of 1.8 ppb was found in an on-site monitoring well at National Industrial Environmental Services in Furley, KS. There were, however, numerous high occurrences in sub-soil samples in hot spot areas, in bottom sediments of evaporation lagoons, and so on, in the hundreds of ppb range.

There have been some limited experimentation showing different patterns of 2,3,7,8-TCDD migration in soils in the presence of solvents or in an oily matrix. Palusky, et al (1986) studied the mobility of 2,3,7,8-TCDD in soils associated with each of 6 solvents. Migration was found to be higher with aromatic solvents and chloroform in comparison to saturated hydrocarbons and methanol. They speculated that the extent of migration related to the solubility of 2,3,7,8-TCDD in the solvent. Puri, et al. (1989) studied the migration potential of 2,3,7,8-TCDD in soil, water, and waste oil mixtures. Over time, they observed a reversible sorption pattern of TCDD, and concluded that a carrier medium with a significant amount of waste oil would play a dominant role in the movement of TCDD through soils.

• Site-specific Characterization: In the case of landfills or sludge land application sites, the assignment of a soil concentration and an area can be made with some reasonableness. Such is not the case with the industrial contamination sites such as those studied in the National Dioxin Study, as briefly discussed above in the concentration bullet. Most of the sites studied in the National Dioxin Study were in the order of tens of hectares to below ten hectares. On the other hand, the Dow Chemical site in Midland, Michigan is described as a site 607 ha in size (Nestrick, et al, 1986). That area corresponds to the size of the property, and the many soil sampling sites within that area were described as "background". Several of the pesticide formulator sites studied in Tier 3 were 2 hectares or less in size. Many of the them were extensively or partially paved with buildings, which complicate fate and transport modeling. Some of the Tier 5 sites of 2,4,5-TCP use were agricultural fields, which are less complicated to describe. However, two sites were described as 2500 acres in size, which again is very large and makes assignment of an average soil concentration non-trivial. Other use sites were described as railyards and railroad rights of way. While estimates of loss into air could be made in complicated sites such as these, use of soil erosion modeling becomes very complicated if not undoable with paved areas, buildings, drainage ditches, roads, and the like.

Contaminants emitted from incinerator stacks are transported in air and deposit on the exposure site, water bodies that may be used for drinking or fishing purposes, and on surrounding land. Chapter 3 describes the application of the COMPDEP (reference) model to obtain vapor-phase air concentrations and deposition rates of particles at a specified distance from an example stack emission source. These quantities are assumed to be given for purposes of discussion in this section; further discussion of the air transport modeling is given in Chapter 3.

Estimating soil concentrations based on particulate depositions follows a similar approach as estimating exposure site soil concentrations resulting from erosion of contaminated soil from off-site areas of contamination. Section 4.5.1. describes how soil concentrations are estimated given total (wet plus dry) deposition rates. Surface water impacts are assumed to result from direct deposition onto surface water bodies as well as erosion from the impacted effective drainage area. This solution is an extension of the solution given in Section 4.3.1. for the on-site source category, and is given in Section 4.5.2. Following now are bullet summaries for similarities and small refinements to algorithms previously discussed:

• Air impacts: The atmospheric transport modeling described in Chapter 3 was comprised of two computer simulations: one which considered that emissions were in a vapor form and were transported as such, and one which considered that emissions were in particle form and likewise were transported as such. The result of the vapor-phase runs was a unitized ambient air concentration at various distances up to 50 km in all directions from the stack. Only the results in the predominant wind direction were used in this demonstration. The result of the particle-phase runs were an ambient reservoir of air-borne contaminants sorbed to particulates (used only for inhalation exposures), and wet and dry deposition unit rates also at various distances up to 50 km. By "unitized", what is meant is that emissions for the vapor or particle runs can be thought of as "1" mass/time (g/sec) emissions. Results for all distances are linear with respect to this emission rate; that is, if the rate of vapor contaminant determined to be emitted is "5", than ambient air concentrations at any location are 5 times what they are when "1" is assumed to be emitted. The same holds true for emissions in the particle phase. Chapter 3 developed a framework for assigning a vapor and a particle fraction for specific dioxin congeners. For example, 2,3,7,8-TCDD was assumed to have a vapor fraction of 0.55 (55% was in vapor form) and a particle fraction of 0.45. The final model results for air concentrations, and dry and wet deposition rates for all congeners, starting from these unit model runs and then incorporating congener-specific emission rates and vapor/particle splits, are given in Tables 3-12 to 3-17. The vapor-phase air concentrations were used to model vapor phase transfers in the vegetative bioconcentration algorithms. They were also used, summed with the simulated reservoir of particle-bound contaminants, to estimate the total reservoir of contaminant available for inhalation exposures.

• Vegetative impacts: The rates of wet and dry deposition modeled by COMPDEL were used to determine vegetative impacts. The model for particle deposition impacts to vegetations is described in Section 4.3.4.2 above. Of course, this above section solves for dry deposition as a reservoir times a dry deposition velocity (for dry deposition), and as a reservoir times rainfall and a washout factor (for wet deposition); such a solution is not required for the stack emission source category since the deposition totals are estimated by the COMPDEP model. Other parameters for the vegetative model - the B_vpa (air-to-leaf vapor transfer factor), the R_w (fraction of wet deposition retained on vegetation surfaces), crop yields and interceptions, and the vegetative washout factor, k_w, are used for the stack emission source category.

• Biota concentrations: The algorithm estimating concentration in fish tissue based on bottom sediment concentrations is the same as in previous source categories. Modeled rates of contaminant deposition on particles onto the exposure site are used to estimate a "tilled" and an "untilled" soil concentration, as described below in Section 4.5.1. Underground vegetable concentrations are a function of tilled soil concentrations. The soil concentration used for cattle soil ingestion is untilled. Beef and milk concentrations are again a function of vegetative and soil concentrations, diet fractions, and bioconcentration and bioavailability factors as described in Section 4.3.4.3.

Chapter 3 describes the use of the COMPDEP Model to estimate the particulate phase deposition rates at the exposure site. This total deposition rate, F, includes both dry and wet deposition, and is used to estimate the steady state soil concentrations. The deposition of contaminated particulates from the air is assumed to be somewhat analogous to the process of eroding contaminated soil from an off-site source depositing on an exposure site. Specifically, the following assumptions are also made: 1) only a thin layer of soil becomes contaminated, 2) this layer is either "untilled" or "tilled", depending on surface activities, and 3) surface residues are assumed to dissipate with a half-life of 10 years corresponding to a first order decay rate of 0.0693 yr^-1. Considerations of upgradient erosion and exposure site soil removal are not made. Depositions occur over the exposure site and surrounding land area on an on-going basis. It might be said that upgradient soil concentrations are similar to exposure site concentrations at all times. Like the soil source categories, a tilled mixing depth of 20 cm is assumed. However, an untilled mixing depth of 1 cm is assumed for this source category, in contrast to the 5 cm assumed for the off-site soil source category. It is felt that the process of erosion assumed to transport contaminated soil in the off-site soil source category to a site of exposure is a more turbulent process. It assumes that contaminated soils mix with "clean" soils that are between the site of contamination and the site of exposure. In contrast, ongoing airborne deposition of particles is felt to be a less turbulent process impacting all watershed soils simultaneously; hence the assumption of a 1-cm mixing depth. The qualitative mass balance statement (similar to the one made above in Section 4.4.1, with _C equalling change in exposure site soil concentrations over time) can now be made as:

(the incremental addition to C resulting from the change in

deposition of stack emitted particulates) -

_C = (the incremental substraction of C resulting from

degradation of residues at the exposure site)

This is mathematically stated as:

where:

C = the exposure site soil concentration, mg/kg

F = deposition rate of contaminant on particles, mg/yr

M = mass of soil at exposure site into which contaminant mixes, kg

k = first order dissipation rate constant, 1/yr.

The solution to this equation is:

which computes C as function of time, t. Similar to the assumption made above in Section 4.4.1., the steady state solution for C is simply F/kM. The deposition rates supplied by the COMPDEP model are in units of g/m²-yr, so a conversion to mg/yr requires a multiplication by the land area of the exposure site and a multiplication of 1000 mg/g. Procedures to estimate M are given above in Section 4.4.1.

4.5.2 Surface Water Impacts

The solution for stack emission impacts to surface water bodies is an extension of the solution for soil contamination described in Section 4.3.1. Stack emissions deposit onto soils within the effective drainage area to result in an average basin-wide soil concentration. Soil erosion then delivers contaminants to surface waters as in Section 4.3.1. Stack emissions also directly deposit onto and impact the surface water body as well. All the assumptions laid out at the beginning of Section 4.3.1 apply here as well. New quantities needed for this solution include: a rate of contaminant deposition onto soils of the effective drainage area used to estimate average soil concentrations (such concentrations are estimated using the approach given in Section 4.5.1. above), a rate of contaminant deposition onto the water body, and a rate of particulate matter deposition onto the water body.

Equations (4-1) through (4-8) are now displayed again with these additions.

where:

C_swb = concentration on soil entering water body, mg/kg

ER_w = total watershed annual soil erosion, kg/yr

DEP_c = total annual direct deposition of contaminant, mg/yr

C_wat = dissolved-phase concentration in water column, mg/L

V_wat = water body annual volume, L/yr

C_ssed = concentration on suspended sediment, mg/kg

M_ssed = mass of suspended sediment introduced per year, kg/yr

C_sed = concentration on sediment settling to bottom, mg/kg

M_sed = mass of bottom sediment introduced per year, kg/yr

Mass balance and equilibrium equations continue:

where:

DEP_p = total annual direct deposition of particulate matter, kg/yr

f_s = fraction of annual erosion remaining as suspended materials, unitless

f_sd = fraction of annual deposition remaining as suspended material, unitless

Kd_ssed = soil-water partition coefficient for contaminant in suspended sediment, L/kg

OC_ssed = fraction organic carbon in suspended sediment, unitless

OC_sed = fraction organic carbon in bottom sediment, unitless

Substituting again as in Equation (4-7):

As before, the bracketed quantity in the right hand side of Equation (4-56) can be termed f , so that C_ssed can be solved as (C_swb ER_w + DEP_c)/f . The numerator in this term can be expanded to describe contaminant contributions by the effective drainage area which has received depositions, the first quantity in the numerator, and to describe direct depositions, the second quantity:

where:

C_swb = concentration on soil entering water body, mg/kg

ER_w = total watershed erosion, kg/yr

DEP_c = annual deposition of contaminant on water body, mg/yr

E = enrichment ratio, unitless

C_w = average soil concentration of dioxin-like compound in effective area of watershed, mg/kg

SL_w = average unit soil loss for land area within watershed, kg/ha-yr

A_w = effective drainage area of watershed, ha

SD_w = sediment delivery ratio for watershed, unitless

RDEP_c = rate of contaminant deposition, g/m²-yr

A_wat = area of water body, m²

1000 = converts g to mg

Again as before, the right hand side of Equation (4-57) can be termed, r , and the concentration in suspended sediment, C_ssed, is equal to r /f . Other water body concentration terms, C_wat and C_sed, can now be solved using Equations (4-54) and (4-55). Guidance on these terms and assignment of values for the demonstration scenarios in Chapter 5 is now given.

• C_swb and ER_w: Equation (4-57) shows all the terms necessary to arrive at an estimate of the annual contaminant entry into the water body via erosion, the C_swb * ER_w term. Section 4.5.1 describes the algorithm to estimate soil concentrations given a deposition rate of contaminant. One deposition rate will be chosen to represent average deposition rates over the effective drainage area of the watershed (the effective drainage area is termed A_w). This rate will be the rate given in COMPDEP modeling at 0.5 kilometers, which implies that the water body and the effective drainage area into the water body are near the stack. Tables 3-15 and 3-16 (Chapter 3) display wet and dry deposition rates for this distance. These rates are added to arrive at total deposition, shown in Table 3-17. Second, a representative mixing depth to characterize average watershed soil concentrations needs to be selected. Previous algorithms used a mixing depth of 20 cm for tillage activities, specifically home gardening, and 1 and 5 cm for non-tilled soil concentrations (1 cm for the stack emission and 5 cm for the off-site soil source category). For the sake of demonstration, it will be assumed that a representative watershed depth will equal 10 cm, which might be interpreted as an average of tilled and untilled lands within the effective drainage area. The values for SL_w (6455 kg/ha-yr), A_w (4000 ha), ER (3), and SD_w (0.15) were all given and discussed in Section 4.3.1. and will not be repeated here.

• DEP_c: The second quantity of Equation (4-57) describes the annual input to the surface water body that comes from direct deposition. This term is RDEP_c * A_wat * 1000, where RDEP_c is the rate of contaminant deposition onto the water body, A_wat is the area of the water body, and 1000 converts g to mg. The rate of contaminant deposition at 0.5 km will also be used to describe direct deposition impact to the surface water body, since for demonstration purposes, there is no justification for saying this distance is further or nearer the point of stack emission. The area of the water body has not been required for any other reason, and one will now be given. First, the effective drainage area of 4000 ha is relatively small and will result in a relatively small stream, at 1.524 x 10⁷ m³/yr flow volume. This volume is also equal to the average cross sectional area of the stream (m²) times stream velocity (m/yr). Assuming a stream velocity of 4.73 * 10⁶ m/yr (15 cm/sec; 1/2 ft/sec), which is reasonable for a small stream, the cross sectional area is solved as 3.22 m². An average 1 meter depth and 3.22 meter width appear reasonable. This width times the stream length would give stream surface area, A_wat. Assuming a rectangular shaped watershed, dimensions of 40 ha wide by 100 ha long (to arrive at the 4000 ha effective drainage) seem reasonable. This length of 100 ha translates to 10000 meters, and the full surface area of the stream is 32200 m². This will be the value assumed for A_wat.

• DEP_p: The rate of particulate deposition onto the lake is required to achieve a mass balance of all annual soil erosion + particle deposition contributions to water body solids. The rate of particulate matter emitting from the stack and arriving at downwind locations was not supplied in Chapter 3. Instead, a literature value of 0.03 g/m²-yr developed by Goeden and Smith (1989) will be used. This value was based on modeling emissions from a resource recovery facility. Total particulate emissions from the stack were projected to be 4.63 g/s, and the deposition rate onto a nearby lake was modeled to be 0.03 g/m²-yr. No further information was supplied. Now, with the surface area as solved for above at 32200 m², the total particle deposition, DEP_p in kg/yr, is 966 g/yr.

• f_s and f_sd: These are the fractions of total erosion and depositing particles remaining as suspended materials within a year. As discussed in the solution for the "on-site source category" in Section 4.3.1, f_s was solved for as: a value for total suspended solid, TSS of 10 mg/L, multiplied by a total flow volume V_wat of 1.524 x 10¹⁰ L/yr, divided by the total erosion into the water body, 3.87 x 10¹² mg/yr. This resulted in an f_s of 0.039. Note that this implies a total suspended load of 152,400 kg/yr. It could be assumed that the minuscule 1 kg/yr of particles directly depositing onto the stream remain in suspension during the year, on the basis of being smaller in size than eroded soil. This assumption will, in fact, be made, but it will be supported as follows.

In a quiescent water body, settling occurs through gravity and can be expressed in terms of Stokes Law:

where:

V_s = Stokes settling velocity, cm/sec

g = acceleration of gravity, 980 cm/sec²

m = absolute viscosity of water, g/cm-sec (poise)

= 0.089 g/cm-sec @ 25 C

r _s = particle density, g/cm³

r = density of water, 1 g/cm³

d = particle diameter, cm

For purposes of this discussion, a reasonable assignment of particle density of is 2.5 g/cm³ for depositing particles or eroding soil. Therefore, making substitutions, the right hand side of Equation (4-58) reduces to 918 d².

Now, assumptions for the particle sizes of eroding soil and depositing particles can be made to arrive at a ratio of settling velocities, V_ssoil/V_spart. The basis for assigning an enrichment ratio for delivery of contaminants via soil erosion was that fine-sized particles were the ones eventually reaching the water body via erosion. Lick (1982) states that a major fraction of the sediments (suspended and bottom) in the Great Lakes are fine grained, silts and clays, and that data from Lake Erie indicates that 90% of the sediments are of this category. Brady (1984) shows USDA's classification of soils according to particle size, and gives a range of 0.0002 to 0.005 cm for silt sized particles and less than 0.0002 for clay size particles. The following assumptions are made to arrive at a representative diameter for particles in eroded soil: eroded soil is comprised of a 50/50 split of these two sized particles, silt-sized particles are, on the average 0.0026 cm in diameter, and clay size particles are 0.0001 cm in diameter. With these assumptions, the average particle size for eroding soil is 0.0014 cm. The settling velocity for a 0.0014 cm particle is 1.8 x 10^-3 cm/sec. In Section 3.4.3, Chapter 3, the argument was developed that 87.5% of the total emission rate of dioxin-like congeners would be associated with particles less than 2 m m. The basis of this argument was a surface area to volume ratio, with smaller particle sizes having significantly larger ratios. This does not mean that 87.5% of the 1 kg/yr depositing particles are of this size. However, for this discussion, the size of depositing particles will be assumed to be 2 m m (2 x 10^-4 cm), since these size particles deliver most of the dioxin-like compounds to the water body (and the ultimate purpose of this exercise is to determine a value for the fraction of depositing particles which remain suspended and impact suspended sediment concentrations). The settling velocity, V_spart, is estimated as 3.7 x 10^-5 cm/sec.

The ratio V_ssoil/V_spart is about 50. Said another way and with all the assumptions and simplifications made above, depositing particles will remain in suspension 50 times longer than eroding soil in a quiescent water body.

Given this high a difference in settling velocities, it seems reasonable to assume f_sb equals 1.0. The fraction of soil erosion remaining in suspension, f_s, will be estimated given TSS, V_wat, etc., as before (see Section 4.3.1), only DEP_p (the total amount of depositing particles, in kg/yr) will comprise a given increment of suspended materials when solving for f_s.

• V_wat, OC_ssed, OC_sed, and Kd_ssed: These have all been discussed in Section 4.3.1. The values for these parameters in the demonstration scenarios in Chapter 5 are: V_wat = 1.524 x 10¹⁰ L/yr, OC_ssed = 0.05, OC_sed = 0.03, and Kd_ssed = OC_sed * Koc, where Koc is the organic partition coefficient of the contaminant.

As discussed in Volume II, Chapter 3, dioxin-like compounds can be released to waterways via various types of effluent discharges such as discharges from municipal waste water treatment facilities and pulp and paper mills using chlorine bleaching. Also discussed is the fact that these emissions have declined substantially in recent years, especially from pulp and paper mills. Since the procedures for considering point source discharges to waterways are somewhat different than those associated with the nonpoint source procedures for soil contamination and stack emissions, they are covered separately in this section. This source category is also different from others in that effluent discharges into surface water bodies are assumed only to impact fish and water.

The approach used in this report is an extension of the "simple dilution" model described in the Superfund Exposure Assessment Manual (EPA, 1988c). Other models are available which offer more spatial and temporal resolution than the model described here. One such model is the EXposure Analysis Modeling System, or EXAMS (Burns, et al., 1982, and Burns and Cline, 1985). The EXAMS and a simple dilution model were both applied in an assessment of effluent discharges from pulp and paper mills (EPA, 1990d). In this assessment, 98 of the 104 pulp and paper mills were modeled with both models using site-specific information (water body flow rates from STORET for all but 6 of the mills, effluent flow rates and contaminant discharges, etc.). Three key quantities - one model result and two model parameters - led to a range of exposure conditions for humans consuming fish impacted by discharges from these pulp and paper mills: a water column concentration, a bioconcentration factor (BCF) applied to the water column concentration to get fish tissue concentration, and a fish ingestion rate. The simple dilution model was used to estimate total water concentrations - i.e., mg TCDD total/L water. The EXAMS model was used to estimate dissolved phase water column concentration - i.e., mg TCDD dissolved in water column/L water. Then, with each set of water concentrations, two sets of exposure estimates (a low and a high estimate, in one sense) were generated - one with a BCF of 5,000 and a fish ingestion rate of 6.5 g/day, and one with a BCF of 50,000 and a fish ingestion rate of 30 g/day. Note that in deriving the range of results in that exercise, the BCF was applied to both a total and a dissolved phase water concentrations. EPA (1993) discusses several bioconcentration/bioaccumulation empirical parameters for 2,3,7,8-TCDD, and makes the clear distinction for those which are to be applied to a total water concentration versus those applied to a concentration in the dissolved phase. The dilution and EXAMS model study indicated that the simple dilution model generally estimated higher water column contaminant concentrations compared to the EXAMS model, although this trend was not consistent among all water bodies modeled. The results from both models were comparable when the receiving water body had relatively low suspended solids concentration.

One key limitation of the EXAMS and the simple dilution model for use with dioxin-like compounds in aquatic systems is that they do not account for sediment transport processes. The EXAMS model was designed to determine the fate of transport of contaminants in the dissolved phase. Another spatially and temporally resolved model for this source category is the Water Analysis Simulation Package, the most up-to-date version termed WASP4 (Ambrose, et al., 1988). This model does include sediment processes and has been applied in a comprehensive evaluation of 2,3,7,8-TCDD bioaccumulation in Lake Ontario (EPA, 1990b). It requires extensive site-specific parameterization, but should be considered for more detailed site-specific evaluations of strongly hydrophobic and bioaccumulating contaminants such as the dioxin-like compounds.

The dilution model described below will be demonstrated in Chapter 5 with a set of data developed using site-specific data from the 104 pulp and paper mills of the 104-mill study. As will be discussed below, a hypothetical effluent discharge will have characteristics developed as the average of key characteristics from the 104 mill study. These key data include: flow rates of the receiving water bodies, suspended solids concentration in these receiving water bodies, effluent discharge flow rates, suspended solids in the effluent discharges, organic carbon content of solids in the effluent stream, and discharges of 2,3,7,8-TCDD.

The principal assumption for the simple dilution model is that contaminants released into a water body uniformly mix and equilibrate with the surrounding water in an area near the effluent discharge point. This area is commonly referred to as a "mixing zone". For application of this model with dioxin-like compounds, what is desired is a concentration on the suspended solids in this mixing zone. Multiplication of the organic carbon normalized concentration on suspended solids and a Biota Suspended Solids Accumulation Factor, or BSSAF, will result in a concentration of contaminant in fish lipids. This is defined similarly to the BSAF used for other source categories of this assessment, except that the organic carbon normalized concentration is that of suspended solids rather than of bottom sediments.

The BSSAF is one of several empirical factors discussed for estimating the impact to fish in water bodies impacted by 2,3,7,8-TCDD (EPA, 1993). Others include the BSAF, total and dissolved phase bioconcentration factors (BCFs), and total and dissolved phase bioaccumulation factors (BAFs). BAFs and similar to BSAFs and BSSAFs in that all three reflect total exposure of fish to contaminant, including water column, sediment, and food chain exposures. The BCFs reflect water column exposures only. EPA (1993) states that there is currently no data available on organic carbon normalized concentrations of dioxin-like compounds on suspended solids, hence no basis to compare BSAF and BSSAF. This assessment assumes a similar numerical assignment of BSSAFs and BSAFs.

The total water concentration in a simple dilution model is:

where:

C_tot = total water concentration, mg/L

MASS_c = mass of contaminant in discharge, mg/hr

Q_u = flow at a point just upstream of effluent discharge, L/hr

Q_e = effluent flow, L/hr

Dissolved phase and suspended sediment concentrations are then estimated using an approach developed by Mills, et al. (1985) and others:

where:

C_wat = dissolved-phase water concentration of contaminant, mg/L

C_tot = total water column concentration, sorbed + dissolved, mg/kg (note: mg/kg is essentially equal to mg/L since 1 L » 1 kg)

Kd_mix = suspended sediment-water partition coefficient for contaminant in mixing zone, L/kg

TSS_mix = total suspended solids in water column in mixing zone, mg/L

C_ssed = concentration of dioxin-like compounds on suspended sediments, mg/kg

10^-6 = converts mg/L to kg/L

The total suspended solids concentration in the mixing zone is a function of the suspended solids just upstream of the discharge point and the suspended solids introduced in the effluent stream:

where:

TSS_mix = adjusted total suspended solids concentration, mg/L

TSS_u = total suspended solids concentration at a point just upstream of effluent discharge, mg/L

TSS_e = total suspended solids concentration in effluent discharge, mg/L

Q_u,Q_e = upstream and effluent discharge flow rates, L/hr

The suspended solids partition coefficient in the mixing zone is a function of the organic carbon partition coefficient of the contaminant and the organic carbon fraction of suspended solids:

where:

Kd_mix = suspended sediment-water partition coefficient in the mixing zone, L/kg

Koc = compound specific organic carbon partition coefficient, L/kg

OC_mix = organic carbon content of suspended sediments in the mixing zone, unitless

This organic carbon content can be solved as the weighted average concentrations of the organic carbon contents of the suspended solids in the effluent discharge and the suspended solids of the receiving water body:

where:

OC_mix = organic carbon content of suspended solids in mixing zone, unitless

TSS_u = total suspended solids concentration at a point just upstream of effluent discharge, mg/L

TSS_e = total suspended solids concentration in effluent discharge, mg/L

Q_u,Q_e = upstream and effluent discharge flow rates, L/hr

OC_u, OC_e = organic carbon contents of suspended solids upstream of the discharge point and within effluent discharge stream

Fish lipid concentrations for this solution are then given as:

where:

C_lipid = fish lipid concentration, mg/kg

BSSAF = biota suspended solids accumulation factor, unitless

C_ssed = concentration of dioxin-like compounds on suspended sediments, mg/kg

OC_mix = organic carbon content of suspended sediments, unitless

Finally, whole fish concentrations are simply this lipid concentrations times a fraction of fish lipid, or C_lipid * f_lipid.

The key model parameter is the BSSAF. A value of 0.09 for 2,3,7,8-TCDD was assumed for BSAF based on data from Lake Ontario. One important difference between the Lake Ontario ecosystem and the effluent discharge source category is that the impact to Lake Ontario is thought to be principally historical (EPA, 1990b), while for the effluent source category, the impact is, by definition, ongoing. This difference may translate to differences in assignment of BSSAF as compared to BSAF. Consider two aquatic settings where bottom sediments are found to have equal concentrations of dioxin-like compounds - one in which contamination is ongoing and one in which contamination is primarily in the past. For the aquatic setting where contamination occurred in the past, water column and suspended sediment concentrations would be lower as compared to the aquatic setting where contamination is ongoing, because water column impacts are only a function of depuration of bottom sediments for the historically impacted water body. It is certainly arguable that exposure of aquatic organisms is greater in the ecosystem where impacts are ongoing, as compared to a system where impacts are historical, when bottom sediment concentrations are equal in the two systems. Now recall the assumption made for the soil contamination and stack emission source categories (in both cases the water body impact is ongoing) concerning the relationship between suspended and bottom sediments - that the organic carbon normalized concentrations are equal. If this is a valid assumption for a system with ongoing impacts, and if in fact fish are relatively more exposed when impacts are ongoing rather than historical, then this argues that a BSSAF for an ongoing contamination setting should be greater in numerical value than a BSAF for a setting where contamination was historical.

However, no data could be found to support such a hypothesis, and there would be no numerical basis for an assumed difference between BSAF and BSSAF. For this reason, the values assumed for BSSAF and BSAF are equal for this assessment. It should be noted that all bioconcentration or biotransfer parameters, such as the BSSAF, are qualified as second order defaults for purposes of general use. Section 6.2. of Chapter 6 discusses the use of parameter values selected for the demonstration scenarios, including a categorization of parameters. Second order defaults are defined there as parameters which are theoretical and not site specific, but whose values are uncertain in the published literature. The parameter values in this category should be considered carefully by users of the methodology.

The effluent discharge solution algorithm was evaluated using data and information from the 104 pulp and paper mill study (EPA, 1990c), which measured discharges of 2,3,7,8-TCDD from 104 mills in 1988, and from the National Study of Chemical Residues in Fish (NSCRF; EPA, 1992a), which measured fish tissue concentrations of 2,3,7,8-TCDD at points downstream from several of these mils. A third modeling study (EPA, 1990d) collected critical data for this modeling evaluation, such as harmonic mean flows downstream of the mills. Finally, the National Council for Air and Stream Improvement (NCASI) provided details on their assessment of this data, which was used here. Importantly, this information included linking specific fish samples to specific mills. A full description of this modeling evaluation is in Chapter 7, Section 7.2.3.6.

There was a dichotomy of model performance as a function of the size of the receiving water body. For most of the mills, the receiving water bodies had harmonic mean flows around 10⁸ L/hr, with a range of 10⁷ to 10⁹ L/hr. A small number of mills, however, discharged into more substantial receiving water bodies which had an average flow of 5 x 10¹⁰ L/hr. Comparing model predictions of fish tissue concentrations for mills discharging into the smaller water bodies, it was found that the model tended to underpredict fish tissue concentrations - the average predicted whole fish concentration was near 7 ppt, whereas the average observed whole fish concentration was near 15 ppt. The same was not true for the large receiving water bodies. In that case, the average whole fish tissue concentration observed was an order of magnitude or more higher than predicted whole fish concentration. No precise explanation could be given for this result. The most likely explanation is that, for these large water bodies, there were other sources of dioxin releases. This comparative exercise did assume inherently that the effluent discharge was the sole source of fish tissue concentrations of 2,3,7,8-TCDD.

It was noted that, for the smaller receiving water bodies, an increase in the assumed BSSAF of 0.09 (which was the value of BSAF assumed otherwise in this assessment) to 0.20 resulted in an average model prediction of fish tissue concentration of near 15 ppt, essentially the same as the observed fish concentration. This could be some empirical evidence for the argument developed above - that the BSSAF for a system with ongoing impacts should be greater in numerical value than a BSAF developed from data on an ecosystem where impacts were primarily historical.

In any case, parameters for the demonstration scenario in Chapter 5 for this source category were derived from 104-mill data. Data from only 77 of the mills was used for the following parameter developments. Mills not included are: 1) the ten mills discharging into the largest water bodies, 2) 9 mills for which EPA (1990d) was unable to derive harmonic mean flows from STORET data, and 3) 8 mills for which data on total suspended solids content in the effluent stream was unavailable from EPA (1990c; actually 11 mills did not suspended solids data, but three were in other categories deleted).

Values of model parameters for the demonstration are now summarized:

• TSS_u, TSS_e: The average upstream total suspended solids term from the 77 mills, TSS_u, was 9.5 mg/L. The average suspended solids concentration within the effluent streams from the 77 mills was 70 mg/L.

• OC_u, OC_e: No information was available on the organic carbon content of the suspended solids upstream of the effluent discharge point. A value of 0.05 was assigned, which was the value assigned for other source categories. No data as well could be found for the organic carbon content of the effluent solids. However, such solids are essentially biosolids from biological treatments of mill sludges. The organic carbon content of such solids is expected to be much higher than 0.05. The value recommended for OC_e was 0.36 (Steven Hinton, PhD., P.E., National Council of the Paper Industry for Air and Stream Improvement, Inc.; Department of Civil Engineering, Tufts University, Medford, MA 02155). This was based on an average proportion of carbon in algal biomass of 0.36 given in Morel (1983).

• Q_u, Q_e: Flow values for the receiving water and effluent stream were summarized in EPA (1990d). The average effluent flow rate, Q_e, for the 77 mills was 4.10 * 10⁶ L/hr, and for the receiving water body, Q_u, was 4.65 * 10⁸ L/hr.

• Koc, BSSAF, f_lipid: Values of Koc and f_lipid are the same ones which have been used for the other source categories. As discussed in the introduction to this section, the Biota Suspended Solids Accumulation Factor, BSSAF, will be assumed to be the same as the Biota Sediment Accumulation, BSAF. This value is 0.09 for 2,3,7,8-TCDD.

• MASS_c: The mass of 2,3,7,8-TCDD exiting from the 77 mills averaged 0.197 mg/hr. However, this data was pertinent for 1988. Since then, pulp and paper mills have reduced the discharge of dioxin-like compounds in their effluents by altering the pulp bleaching processes. Gillespie (1992) reports that data on effluent quality from all 104 mills demonstrate reductions in discharges of 2,3,7,8-TCDD of 84%. On this basis, the value of MASS_c for all three example compounds will be 0.0315 mg/hr (16% of 0.197 mg/hr).

Using these parameters in the simple dilution model for 2,3,7,8-TCDD results in the following:

1) If the mass loadings of 2,3,7,8-TCDD are assumed to be fully sorbed to solids in the effluent discharge, and not to exist in the soluble phase in the discharge, than the concentration of 2,3,7,8-TCDD on discharging effluent solids is 1.1*10^-4 mg/kg, or 110 ppt.

2) The total suspended solids concentration in the mixing zone, TSS_mix, equals 10.0 mg/L. The organic carbon content of suspended solids in the mixing zone, OC_mix, is estimated as 0.069. It is seen how the effluent discharge influences these two key quantities: the unadjusted TSS_u was given as 9.5 mg/L, and the unadjusted OC_u was 0.05.

3) The overall suspended solids concentration of 2,3,7,8-TCDD in the mixing zone after mixing and equilibrating with surrounding water, C_ssed, was 4.5 ppt. This compares to the concentration that might be on the effluent solids of 110 ppt, indicating more than an order of magnitude reduction in concentration by mixing with solids of the receiving water body, and partitioning into the water column.

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