CONTINUED: 4.3.1. Surface Water and Sediment Contamination

. Rainfall/erosivity index, R:

The R term represents the influence of precipitation on erosion, and is derived from data on the frequency and intensity of storms. This value is typically derived on a storm-by-storm basis, but it has been compiled regionally for the development of average annual values (EPA, 1977). Annual values range from < 50 for the arid western United States to > 300 for the Southeast. The value used in this assessment will be 160, which is typical of rainfall patterns seen in much of the midwestern United States.

. Soil erodibility, K:

The soil erodibility factor reflects the influence of soil properties on erosion, with values ranging from <0.05 for non-erodible sandy soils to >0.50 for highly erodible silty soils. The value used in this assessment will be 0.30, which is typical of, for example, sandy or silty loam soils with 2% - 4% organic matter contents.

. Length-slope factor, LS:

The topographic factor reflects the influence of slope steepness and length of the field in the direction of the erosion. Steeper slopes and longer lengths lead to higher LS values, with a range of 0.1 for slopes <1.0% and lengths <100 ft to >2.0 for slopes generally >10%. The two key considerations for its assignment, therefore, are the size of the field for which erosion estimates are being made, and the slope of that field. The example scenarios in Chapter 5 had field sizes of 0.4 ha (1 ac) for a rural residence, 4 ha (10 ac) for a small rural farm, and 10 ha (25 ac) for an off-site contamination site.

Guidance for use of the Universal Soil Loss Equation stops short of defining appropriate sizes of field for which unit estimates are to be derived, except that the USLE was developed for agricultural "fields" where cover, slope, soil type, etc. are assumed to be uniform. For purposes of estimating erosion losses in this assessment, a field of 4 ha for estimating the LS factor will be used. In a rural watershed with agricultural and non-agricultural settings, this would be a reasonable average area of uniformity.

If square shaped, a 4 ha area translates to a side length of 200 m. For purposes of assignment of the LS factor, it will be assumed that the contaminated site has a 2% slope. EPA (1977) (and other references as noted above) show nomagraphs giving the LS factor as a function of slope length and slope. With a 200 meter slope length and a 2% slope, the LS factor is approximately 0.20. This factor will be used for all soil loss estimates required in this estimates.

. Support practice factor, P:

The support practice factor reflects the use of surface conditioning, dikes, or other methods to control runoff/erosion. P can be no greater than 1.0. However, values less than 1.0 should only be assigned when specific practices are employed which are designed to reduce erosion. For the example scenarios in Chapter 5, it will be assumed that no such practices are in place at the site of concern or throughout the watershed to control erosion. Therefore, a value of 1.0 will be assumed.

. Management practice factor, C:

The final term in the USLE is the cover and management practice factor, C, which primarily reflects how vegetative cover and cropping practices, such as planting across slope rather than up and down slope, influences erosion. C values can be no greater than 1.0, with this value appropriate for bare soils. A C value of 1.0 is an appropriate choice for active landfills or sites of high soil contamination (like Superfund sites) mostly devoid of vegetation.

For an inactive landfill with grass cover or any area with dense vegetative cover such as grass, a value of 0.1 or less is appropriate. Values greater than 0.1 but less than 0.7 are appropriate for agricultural row crops, which offer more protection than bare soil, but not as much protection as dense vegetation. Three erosion estimates are required for scenarios demonstrated in Chapter 5. One is for areas of high soil contamination, or the scenario demonstrating the "off-site" source category. It will be assumed that the off-site contaminated site is largely devoid of vegetation in this case, and a value of 1.0 will be assumed.

A second erosion estimate is needed to characterize average unit soil loss throughout a watershed draining into a surface water body. The example scenarios are based on a rural setting which has agricultural and non-agricultural (i.e., rural residences) areas. The C value in this circumstance will be assumed to be 0.3. Finally, an soil erosion estimate is needed in the algorithm transporting contaminated soil from an area of high soil contamination to a nearby site of contamination, as part of the algorithms developed for the "off-site" source category. In this case, the land between a site of soil contamination and the nearby site of exposure will be assumed to be covered with dense vegetation, such as grass. In this case, the C value will be 0.1.

As just described, three unit soil loss estimates are required for this estimates and the difference between the three will be expressed in the C term. Multiplication of the five USLE terms gives unit soil loss estimates of 9.60 (with C = 1.0), 0.96 (with C = 0.1), and 2.88 (with C = 0.3) t/ac-yr. The value of SLs and SLw for the demonstration of the on-site scenario in Chapter 5 is 2.88 t/ac-yr. Since Equation (4-8) and other uses of unit soil loss estimates are needed in kg/ha-yr, these unit losses are easily converted to 21515, 2152, and 6455 kg/ha-yr.

. As and Aw:

These are the area terms, including the area of the contaminated site, and the effective drainage area of the watershed, both in ha. The scenarios demonstrated in Chapter 5 have assumed 0.4 ha (1 acre roughly) for exposure sites described as rural residences, 4 ha (10 acres) for farms, and 10 ha (25 acres) for an off-site area of soil contamination. If the area of contamination is at the site of exposure, as in the "on-site" source category, then As should be assigned an area equalling the site of exposure (and the concentration term, Cs, should equal the average soil concentration over this site of exposure). If the area of contamination is away from the site of exposure, as in the "off-site" source category, As should equal the total area of contamination (and again Cs should equal the average soil concentration over this area).

The total area impacting a river system has been termed a watershed. For purposes of this assessment, an "effective" drainage area will almost always be less than the total area of a watershed. A "watershed" includes all the land area which contributes water to a river system. For large river systems, this area is in the order of thousands of square miles and includes several tributaries and smaller streams feeding into the main branch of the river. Each stream and tributary has its own sub-basin, whose sediment originates from a land area much smaller than thousands of square miles.

If the contaminated site lies within that sub-basin, that it would be appropriate to include only the area within that sub-basin as the effective drainage area. This is one circumstance where an "effective drainage area" would be less than a total watershed area. Another consideration for determining the effective drainage area is the positioning of the contaminated site with respect to the point where water is extracted for drinking and fish are caught for consumption. If these points are significantly upstream in the river system in relation to the contaminated site, there is no reason to conclude that sediments or water near where the water is extracted are impacted by the contaminated site.

If these withdrawal points are downgradient of the contaminated site, then there is reason to believe that sediments and water are impacted. However, if they are downgradient from the contaminated site but not at the bottom of the watershed, then sediment and water quality further downgradient from the withdrawal points is not of concern and land draining into these downgradient portions would not be part of the "effective drainage area". One further possible consideration is how far upgradient in the watershed one should go when determining the size of the effective drainage area. While sediments introduced at the furthest points may eventually work their way down to the mouth of the watershed, this may take geologic time and not recent historic time. Therefore, sediment quality near a site of contamination need not consider these far reaches.

For a standing water body such as a lake or a pond substantially fed by ground water recharge, an assumption that probably should be made using the simple framework of this assessment is that all sediments within the lake/pond are completely mixed. Therefore, the effective area should equals all area around the lake/pond contributing sediment, and, as in the above discussion on river systems, a part of the land area contributing sediments to streams or rivers which may feed the standing water body. From this discussion, it is clear that determination of an effective drainage area depends on site specific considerations, but it will likely be less than the total watershed area.

For purposes of demonstration, the effective drainage area, Aw, will be assumed to be 4,000 hectares (10,000 acres, 15.6 mi2). Furthermore, it will be assumed that the water body in question is part of a river system, which mainly impacts the assignment of the total suspended solids parameter, TSS (as discussed below). This assignment is not based on any specific sites that have been studied. It is only justified as being a reasonable size for dilution of contaminated soils which originate from contaminated site. Given the other area terms discussed above, 0.4, 4, and 10 ha, then the assignment of 4000 ha would appear to add a substantial amount of clean soil for mixing considerations.

A useful data source for this term and the suspended sediment term below, for specific sites in the United States, is Appendix F in Mills, et al. (1985). This appendix includes a compilation of data from river and reservoir sediment deposition surveys, including total drainage area, water body volumes, and rates of sediment deposition (mass/area-time). A caution in using this and similar data bases when evaluating specific sites is that, again, these total drainage areas are just that, total areas. Water bodies in this data base are located in the 48 conterminous states. An estimate of suspended sediment concentrations can be made using the water volume and the sediment deposition rates from this data, and an assumption on sediment deposition velocity. The specific weight of sediments in the water body, also supplied in this appendix, can be used to estimate sediment deposition velocity.

. SDs and SDw:

These are the sediment delivery ratios applied to the exposure site and the watershed as a whole. Such a ratio is required because not all the soil which erodes from an area deposits into the receiving water body. The following delivery ratio was proposed for construction sites (EPA, 1977):

Equation V3 4-10

Note that the sediment delivery empirical equation simplifies all land features pertinent to erosion to a function only of length. The equation was developed to estimate sediment loads from construction sites to nearby surface water bodies, and from distances up to 250 m (800 ft, roughly). Without specific information on the sites from which it was developed, it is assumed that the land area between the construction sites and the receiving water body is "average" and this relationship can be used for applications other than construction sites.

As noted in previous bullets, the example scenarios demonstrating the on-site source category assumed Cs = Cw, and SLs = SLw. The impacted water body was assumed to be 150 meters away from the site of contamination, also the site of exposure for the on-site source category. This distance translates to a delivery ratio of 0.26. Site-specific conditions could result in a larger (steeper slope, e.g.) or smaller proportion of the eroded soil being delivered to the water body than would be estimated with this equation.

Figure 4-5 shows a watershed delivery ratio as a function of watershed size (figure from Vanoni, 1975). As seen, the ratio decreases as land area increases. The total watershed size assumed for the example scenarios in Chapter 5 was 4,000 hectares, or 40 km2. From Figure 4-5, this translates to a watershed delivery ratio, SDw, of 0.15.

. fs:

As soil erodes into the water body, it will settle onto the bottom to become bottom sediment. Part of the settled material will become resuspended because of turbulent flow. The finest materials in eroded soil may not settle for a long time, and essentially always be in suspension. One way to arrive at the fraction of annually eroding material which remains in suspension ("remains in suspension" for purposes of discussion - in reality, little, if any, will remain in suspension, but will rather deposit and resuspend) involves complex modeling. A wealth of such models exist, such as those described in Wang (1989). The approach used here is more simple than those in Wang (1989).

If an average level of suspended material in the water were specified, in units of mg/L, what would be known with otherwise required parameters is the total amount of erosion reaching the water body (as discussed above) as well as the annual water volume (discussed below). ...
table Figure 4-5 Watershed delivery ratio, SDw, as a function of watershed size.
... A required parameter for this assessment will therefore be the level of suspended solids in the water body, TSS. With this parameter and the annual water flow volume, Vwat, the total suspended load equals, TSS (mg/L) * Vwat (L/yr). The assignment of these two terms are 10 mg/L and 1.524*1010 L/yr, leading to a total suspended load of 1.524 * 1011 mg/yr, or 1.524*105 kg/yr. Total erosion into the water body, in similar units, equals, As * SLs * SDs + (Aw - As) * SLw * SDw. With parameter assignments as discussed above, the total annual erosion equals 3.87 x 106 kg/yr. Therefore, the fraction of total load that is suspended is 0.04 (1.524*104/3.87*106).
expand table Figure V3 4-5

Given this formulation, the fs term is not a model input value, but is solved on the basis of the other parameters noted.

. TSS:

This is the total suspended sediment in the water body. This value will be lower for standing water bodies such as ponds or lakes as compared to streams or rivers. The more turbulent flow in rivers will suspend sediments to a greater degree than a relatively calm lake. A complex modeling exercise evaluating the impact of 2,3,7,8-TCDD to Lake Ontario assumed a suspended sediment concentration of 1.2 mg/L (EPA, 1990b). For use in pond or lake settings, an assumption of a suspended sediment concentration of 1-2 mg/L is reasonable. All example scenarios in Chapter 5 assume that the 4,000 ha watershed drains into a river suitable for supporting fish for consumption and water for drinking purposes.

General guidance offered for the potential for pollution problems in rivers and streams as a function of average suspended sediment concentration are: 10 mg/L or less - no problem, 100 mg/L or less - potential problem, and greater than 100 mg/L - probable problem. A cutoff concentration for protection of aquatic life is 80 mg/L (Mills, et al., 1985). The value assumed for TSS for all example scenarios in Chapter 9 is 10 mg/L, indicating no turbidity problems and a river supportive of fish for consumption.

. Vwat:

The stream in the example scenarios will be assumed to derive its annual flow only from the effective drainage area, Aw. This would imply that the scenarios are best described as sub-basins ( see the discussion on effective drainage area, Aw, above). Given the area of drainage, one way to estimate annual flow volume is to multiply total drainage area (in length squared units) times a unit surface water contribution (in length per time). The Water Atlas of the United States (Geraghty et al., 1973) provides maps with isolines of annual average surface-water runoff, which they define as all flow contributions to surface water bodies, including direct runoff, shallow interflow, and ground-water recharge.

The range of values shown include 5-15 in/yr throughout the Midwest cornbelt, 15-30 in/yr in the South and Northeast, 1-5 in/yr in the desert Southwest, and a wide range of 10-40 in/yr in the far West. For this assessment, an assumed 15 in/yr is used to estimate the annual flow volume. Over a 4,000 hectare drainage area, total flow volume equals 1.524 x 1010 L/yr (15 in/yr * 0.0254 m/in * 4,000 ha * 10,000 m2/ha * 1000 L/m3).

. Kdssed:

This adsorption partition coefficient describes the partitioning between suspended sediment and the water column. For numerous applications for organic contaminants, particularly for estimating the partitioning between soil and soil water, this partition coefficient has been estimated as a function of the organic carbon partition coefficient and the fraction organic carbon in the partitioning media:

Equation V3 4-11

The organic carbon partition coefficient, Koc, can be a measured value or it can be estimated. Schroy, et al. (1985) listed an organic solids/water partition coefficient of 468,000 for 2,3,7,8-TCDD. Information in Jackson, et al. (1986), imply that this is a very low partition coefficient for 2,3,7,8-TCDD. They obtained soil samples contaminated with 2,3,7,8-TCDD from 8 sites in the Times Beach area of Missouri, and 2 from industrial sites in New Jersey.

These contaminated soils had 2,3,7,8-TCDD concentrations ranging from 8 to 26,000 m g/kg (ppb), and organic carbon contents ranging from 0.015 to 0.08. They determined soil water partition coefficients, Kds, for these soil samples, and using the organic carbon fraction data, estimated Kocs for 2,3,7,8-TCDD. The mean Koc from these ten samples was roughly 24,500,000. EPA (1990b) evaluated the Koc for sorption of 2,3,7,8-TCDD onto Lake Ontario sediments. They concluded that log Koc was greater than 6.3 (Koc = 2,000,000), but less than 7.3 (Koc = 20,000,000).

In the absence of measured values, the Koc can be estimated from a chemical's octanol water partition coefficient, Kow. Empirical equations relating Kow to Koc are listed in Lyman, et al. (1982). Of six different equations listed in that reference, the following derived by Karickhoff, et al. (1979) is used to estimate the Koc for the example compounds in Chapter 5:

Equation V3 4-12

This equation was empirically developed from a limited number of hydrophobic contaminants (n=10, R2 = 1.00). It implies that Koc is very similar to Kow for strongly sorbed compounds such as the dioxin-like compounds. Using the log Kow of 6.64 given in this assessment for 2,3,7,8-TCDD in Karickhoff's relationship estimates a Koc of roughly 2,700,000.

. OCsed, OCssed:

The organic carbon content of solids and sediments of water bodies are generally higher than organic carbon contents of the surrounding lands. Furthermore, organic carbon contents of suspended organic materials and solids are typically greater than those of bottom sediments. A significant sink for strongly hydrophobic contaminants such as the dioxin-like compounds is thought to be suspended, or non-settling, organic material. In modeling 2,3,7,8-TCDD in Lake Ontario (EPA, 1990b) using the WASP4 model, a compartment separate from suspended solids termed "non-settling organic matter" served as a permanent sink.

For purposes of this assessment, a single reservoir of suspended materials onto which incoming dioxin-like compounds sorb is principally characterized by OCssed, and the values selected for OCsed and OCssed should reflect the relative partitioning behavior of suspended and bottom materials. As noted above, these water body carbon contents are also related to the organic carbon contents of surrounding soils. The model parameter, OCsl, is the soil organic carbon fraction and is required for modeling of soil contamination by dioxin-like compounds.

Foth (1978) lists the organic nitrogen content of several soil types ranging from sand and sandy loam to clay. The range from that list is 0.0002 - 0.0024 on a fractional basis. Assuming a carbon to nitrogen ratio of 10 (Brady, 1984; who notes that C:N ratios vary from 8 to 15, with the typical range of 10 to 12), organic carbon ranges from 0.002 to 0.024. A soil organic carbon fraction, OCsl, is assumed to be 0.01 for all example settings in Chapter 9, which is in the middle of this range. The organic carbon content of bottom sediments, OCsed will be higher at 0.03. Bottom sediments originate as erosion from surrounding land, but also include decay of organic materials within water bodies. The organic carbon content of suspended materials can approach 0.20, but OCssed will be assumed to be 0.05 for the example settings in Chapter 5.

4.3.2. Vapor-Phase Air Concentrations

The algorithms for estimating vapor-phase concentrations of contaminants were presented and derived in Hwang, et al. (1986). These procedures were developed for soil surface and subsurface contamination with polychlorinated biphenyls, PCBs. The models are based on the assumptions that:

1) PCBs move through the soil primarily by vapor phase diffusion, i.e., leaching is not considered,

2) PCB vapor in the soil matrix reaches a local equilibrium with pore air,

3) degradation processes for PCBs were not considered , and

4) the PCB contamination occurs at the surface and extends down infinitely.

These assumptions are similar to the general types of assumptions that have been made for all the algorithms estimating exposure media concentrations in this assessment. The procedures in that PCB assessment were also used for this assessment. Details of the derivation are presented in Hwang, et al. (1986).

The average flux rate over an exposure duration of ED can be estimated as:

Equation V3 4-13

The effective diffusivity, Dea, is solved as a function of contaminant diffusivity in air, and soil pore porosity:

Equation V3 4-14

The soil adsorption partition coefficient, Kds, is given as:

Equation V3 4-15

It is noted in Hwang, et al. (1986) that this procedure would tend to overestimate emissions and resulting exposures in situations involving small spills which would not involve deep contamination. It is also noted that the average flux rate is inversely proportional to the square root of the duration of exposure - the longer the duration of exposure, the lower will be the average flux rate. Whereas this solution assumes an unlimited reservoir of contaminant, it is an unsteady state solution (unlike other solution strategies) and is essentially an average flux rate over an amount of time defined by the exposure duration. Inherent in the solution was the consideration that residues dissipate by volatilization at the surface layers, resulting in contaminants diffusing upwards from deeper soil layers over time. With this longer path of diffusion, volatilized amounts decrease, and hence the average flux over time also decreases.

Vapor-phase concentrations along the center (y=0.0) of an area source can be estimated from (Hwang, 1987):

Equation V3 4-16

This was the model used to estimate on-site vapor-phase concentrations. The dispersion terms, Sz and Sy can be estimated using site-specific wind rose data. In the absence of data, these terms can be estimated assuming the most common stability class, D, as:

Equation V3 4-17a/b

Background on Koc and OCsl were given in Section 4.3.1. above. Guidance for other terms in this algorithm now follow.

. Eslp:

Porosity is defined as the pore space in soils occupied by air and water, and for sandy surface soils show a range of 0.35-0.50. Medium to fine-textured soils (loams, clays, etc.) show a higher range of 0.40-0.60 (Brady, 1984). Soil porosities in the example settings were 0.50.

. H:

Henry's Constants were discussed in Volume 2, Chapter 2. The values of H used for the three example compounds were: for 2,3,7,8-TCDD - 1.65*10-5 atm m3/mole; for 2,3,4,7,8-TCDF - 4.99*10-6 atm m3/mole; and for 2,3,3',4,4',5,5'-HPCB - 1.00*103 atm m3/mole.

. Psoil:

Particle bulk density is defined as the mass of a volume of soil solids. This contrasts the more common parameter, bulk density, which is the mass of a unit of dry soil, which includes both pores and solids. Particle bulk density, Psoil, has a narrow range of 2.60 to 2.75, and for general calculation purposes, Brady (1984) recommends a value of 2.65 for average mineral surface soils, the value used for the example settings.

. ED:

The exposure duration is simply the amount of time individuals are exposed. Two exposure durations were used in the demonstration scenarios, 9 years for "central" and 20 years for "high end" exposures. Used in this algorithm, and as discussed earlier, longer exposure durations translate to lower average volatilization fluxes. This presumes a soil concentration assumed to be uniform over depth starting at time zero, and to become depleted over time. The selected exposure durations of 9 (2.83*108 sec) and 20 years (6.31*108 sec) was used.