Review of the Risk Analysis for the Round Two Biosolids Pollutants (Dioxins, Furans and Co-planar PCBs), dated December 14, 1999

By Ellen Z. Harrison, Director, Cornell Waste Management Institute
This review was performed by Ellen Harrison at the request of Versar, Inc. under a contract with the US EPA.
This review of the Risk Analysis for the Round Two Biosolids Pollutants (the RA) performed by Abt Associates on behalf of the US EPA focuses on areas in which I and some of the colleagues with whom I consulted have expertise.
Overall this is a disappointing document from both the point of view of ease of understanding and also methodology and execution. The level of detail is scant as compared to other related documents prepared for EPA. Working through examples of the actual application of the various algorithms would make it significantly easier to evaluate and this is done in many similar risk analysis documents. Key assumptions are not well explained or justified. Specific examples of these deficiencies are noted below.
Risk Assessment Issues
It is not clear why the RA took the approach of assessing the impact of application of sludges containing 300 ppt TEQ dioxins rather than working from an "acceptable" risk level to derive a concentration in sludge that would correspond to that level of risk (as was done for Round 1 pollutant evaluation). This latter approach lends itself more easily to adjusting calculations if cancer risks are reassessed (as seems to be in the works). The approach taken in the RA seems geared towards rationalizing the acceptability of application of all but the most dioxin-contaminated sludges since 95% of US sludges would meet the 300 ppt level.
Use of different assumptions in the risk assessment process leads to different standards. Using a risk assessment process, the state of Wisconsin evaluated dioxin risks from paper sludge land application and derived a cancer-based standard of 1.2ppt in soil associated with a one-in-a-million risk and a value of 0.19ppt where grazing is allowed (Goldring, 1992). If a higher allowable cancer risk of one-in-ten-thousand is accepted by EPA in the sludge rules, acceptable levels based on the Wisconsin assessment would be 19 ppt in soils where grazing is allowed. Our understanding of the RA is that 30 ppt TEQ in pasture soil was calculated from the application of sludges containing 300 ppt under the assumptions used in the RA (Exhibit 3-7). Thus even based on cancer as an endpoint, using the Wisconsin risk assessment assumptions and methods would result in a lower standard. This makes the point that what assumptions are made is critical to the RA results.
As discussed below, however, cancer may not be the most sensitive end point where risks greater that one-in-a-million are accepted. If, however, the reassessment of cancer risk for dioxins shows it to be 10-100 times greater than assumed in the RA, then cancer would likely be the sensitive end-point based on current information.
Simultaneous exposure to multiple pathways
Realistically, the HEI (which is specified appropriately in the RA as the farm family) would be exposed to multiple pathways simultaneously. Only in the breast feeding scenarios (15a and b) are multiple exposure pathways considered in the RA, and then only for the mother and not the child. For pathway 15a, the HEI breast-feeding farm family mother would be exposed via pathway 2 (home gardener) rather than pathway 1 as analyzed. The farm family is likely to have a garden and to eat from it.
The RA provides no justification for this failure to sum exposures to multiple pathways. Protocols are suggested elsewhere by EPA for the assessment of human health risk from hazardous waste combustion facilities (EPA 1998), where dioxin-like compounds are a primary issue and exposure routes would be similar. There, it is suggested that the appropriate exposure scenario is defined as a combination of exposure pathways. For the HEI, the following pathways are listed (Table 4-1, EPA 1998) as applicable: direct inhalation of vapors and particles; incidental soil ingestion; ingestion of drinking water; and ingestion of homegrown foods (including produce, beef, milk, chicken, eggs and pork); plus potentially the ingestion of fish from on-farm ponds or watercourses.
Assessing the risk to the breast feeding farm child should evaluate more than only intake via breast milk. Inhalation, while representing a relatively low exposure route for adults, may be higher for infants given their greater respiration rate. This should be evaluated. In addition, a more holistic evaluation of the exposure of young children is needed (see below).
Without making adjustments to the risk levels calculated in the RA based on the suggestions that follow, it is difficult to estimate the cancer risk for the HEI based on multiple pathway exposure. However, summing the cancer risks for each of the relevant pathways calculated in the RA (Exhibit 4-1), namely pathways 2 (home garden), 3 (soil ingestion), 4 (animals eating forage), 5 (animals eating soil), 11 (inhalation of dust), 12 (drinking water and ingesting fish), and 13 (respiration) results in a 1.3 x 10-4 cancer risk. This does not include dermal exposure, adult soil ingestion, nor consumption of poultry and eggs. It also does not reflect the reassessment of cancer risks reflecting a 10 to 100 fold greater potency than used in this RA which is currently under discussion at EPA. The rationale for including both pathways 4 and 5 as well as numerous other modifications of the risk calculations are discussed below.

Additional pathways not included in the RA

The dermal exposure pathway is not addressed at all in the RA, nor in the EPA 1998 document. It appears that this path may be among the several more significant exposure routes when soil levels are the concern. It represented 8% of the contribution for farm family exposure as calculated in the useful evaluation of exposure scenarios to only 1 ppt in soil in the EPA Estimating Exposure to Dioxin-like Compounds (Table III-4, US EPA 1994).
Exposure via poultry including particularly eggs is not included in the RA. Recent work suggests a cancer risk of 10-4 for individuals eating 3-4 eggs per week from chickens exposed to soils containing 30-40 ppt TEQ (Harnly et al., 2000). This is not an unlikely pathway for farm family exposure and should be included. Free range poultry eat soil and would bioaccumulate dioxins. Geese were found to consume 8% of their diet as soil and wild turkey 9% (Beyer et al., 1994). Free range poultry are suggested to exhibit similar behavior, with 10% of their diet being soil (US EPA, 1998). Chickens accumulate dioxins and poultry foraging on soils containing low ppt PCDD/PCDF levels are predicted to bioaccumulate these compounds to unacceptable levels (Stephens et al., 1995). The risks to those consuming poultry and eggs should be assessed in the RA and the risks added to those of other relevant pathways.

Uncertainty and variability

The point estimates used for a number of parameters in the RA are questionable due to both variability and uncertainty. These include many of the parameters discussed further below. The RA should at least provide a sensitivity analysis to assess the impact of the deterministic assumptions on the results. Better, a revised risk assessment using probabilistic methods which incorporate the various suggestions in these comments and which then calculates a daily intake that includes background intake from non-sludge sources could be conducted.
Realistic data and scenarios incorporating the variability and uncertainty should be used in a multi-media, multi-exposure probabilistic analysis. EPA itself recognizes the need to represent variability and uncertainty in risk assessments and the value of probabilistic tools such as Monte Carlo analysis. (US EPA, 1997a; US EPA, 1997b, US EPA, 1999).
There are many examples of parameters for which variability can play an important role in quantifying risks. Using probabilistic methods to analyze impact of breast-feeding, Hoover showed that fat content in breast milk is a key variable. If an inappropriate single point value was used, results are skewed. (Hoover, 1999). Dietary intake of key foods such as meat, eggs, poultry and milk varies widely.
Uncertainty also makes the use of single point estimates questionable. Uncertainty is great regarding dioxin cancer potency and toxicity due to very different responses of different species. There are inadequate data on impacts on humans during fetal development and infancy (Hoover, 1999) (also inadequate data on wildlife for these same developmental impacts).
TEFs are continuing to evolve, so their use is important but a source of uncertainty. The recent WHO consultation reevaluted them and recommends revisions that would result in an approximate 10% increase in TEQ calculations compared to using I-TEFs (WHO, 1998). There are a number of other halogenated compounds that could contribute to the total concentration of compounds exhibiting similar toxicity. These include brominated analogues of PCDD and PCDFs (Van den Berg et al., 1998). Polybrominated biphenyls and dioxins seem to pose similar risks to dioxins and PCBs (US EPA, 1994; Hornung et al., 1996; Helleday et al., 1999, Weber and Greim, 1997; Henck et al., 1994). They appear to be carcinogenic (Hoque et al., 1998; Henderson et al., 1995). They are detected in sludges (Hagenmaier et al., 1992a). If they act in similar toxicologic or oncogenic mode to dioxins, they need to be factored into the risk assessment or at least recognized as contributing to the risk. Numerous other PCBs might also be included.
In regard to the groundwater pathway, the role of facilitated transport of dioxins bound to organic matter needs to be investigated. Sorption onto organic matter may give rise to the facilitated transport of these compounds into ground water (Nelson et al., 1998). This is a particular concern as complexation of hydrophobic chemicals with organic matter can also inhibit the ability of microorganisms to degrade these compounds even though they may still be available and therefore toxic to higher organisms (Rinella, 1993 #4).
In assessing the home garden scenario, there does not appear to have been specific assessment of the exposure due to eating of members of the genus Cucurbita. This group has been shown to exhibit unusually high uptake (Hulster et al., 1994). These squash and cucumbers are commonly grown vegetables in the home garden. Whether their contribution to exposure via this pathway is important or not needs to be evaluated.

Unclear assumptions and rationale

A number of the assumptions in the RA need to be more transparent. For example, it is not clear why the application rates in the RA is changed from Round 1 TDS assumptions. The soil ingestion rate assumptions were also e changed. The rate of daily ingestion changed from 200 to 400 mg/day and the assumption regarding what is ingested changed from straight sludge to a sludge soil mixture. The proportion of sludge in this mix and the concentration of contaminants in that mixture is not clear in the RA. Since the RA suggests that this is the highest risk pathway, these assumptions are critical and need to be clear and well justified. Better, a probabilistic look at this pathway would be important given some of the uncertainty surrounding soil concentrations (see below) as well as ingestion rates.
A critical example about which we would seek clarification has to do with how soil concentrations were calculated in the RA (Exhibit 3-7). Various equations are displayed, but the justification for the values used are not clear. It is not clear how much of the dioxins is projected to leach, volatilize and be lost through erosion. While the equations for loss are given, working out of examples is needed for clarity and for comparison.
The RA faces a complex task in trying to simultaneously assess the risks posed by an array of related chemicals that possess certain similarities, but also behave differently from each other. The behavior of the different congeners in regard to partitioning and fluxes, breakdown and metabolism, etc. are significantly different (Douben et al., 1997; Cousins et al., 1997). For some of the parameters, individual values are used for each of the congeners, while for others a single value is used to represent all the congeners. The RA should be more explicit about when each of these different approaches is applied.
In the RA several algorithms include a parameter which is the area receiving sludge. The figures used are not consistent in different sections of the RA. For calculating inhalation exposure, 1074 hectares are assumed to be sludged, while only 28 hectares is used in the surface water and air to plant pathways. No justification is provided for either of these figures nor for why two different figures would be used in the same assessment.
The frequency of application of sludge is assumed to be every other year in all parts of the RA except the pasture, where it is assumed to be applied every 3 years. There is no rationale or explanation for this difference in assumptions. The RA appears to assume zero dioxin ingestion by animals in yrs 2 and 3 when no sludge is assumed to be applied. It is not reasonable to assume that levels fall to zero in pasture and that there is no residual in the years following application.
In the RA, many of the risk assessment assumptions questioned in our analysis of Round 1 (Harrison et al., 1999) also pertain to this assessment. For example, a very low percent of watershed is assumed to receive sludge (0.006%). For small watersheds, this is not a realistic assumption. In our experience on a dairy farm in upstate New York, large acreage can receive sludge and be the majority of the watershed of small, but fishable, receiving streams. Other EPA assessments suggest that on-farm ponds may be a source of farm family fish and should be evaluated (US EPA, 1998). Another variable which we found in previous analysis to underestimate potential exposure is the produce intake of home gardeners. Use of a single point estimate of dietary intake does not seem appropriate for the wide range of consumption patterns.

Cancer is not most sensitive end point if 10-5risk is used
A wide range of non-cancer effects are found (Brouwer et al., 1998). Non-cancer endpoints, especially developmental impacts through fetal and nursing infant exposures, are more sensitive end points than cancer where less than one-in-a-million cancer risk is used as cancer end point (data in US Dept. of Health and Human Services, ATSDR, 1998; WHO, 1998; Dr. Henry Anderson, Wisconsin Department of Health and Social Services, personal communication). The immune system is a sensitive target for toxicity (US Dept. of Health and Human Services, ATSDR, 1998).
Developmental behavioral impacts were found at a does of 0.12 ppt/kg bw/day in rhesus monkeys and developmental effects are found to be among the most sensitive LOAEL in animals leading to a chronic oral MRL based on these effects (US Dept. of Health and Human Services, ATSDR, 1998).

Impact to wildlife and soil organisms not assessed
Data on ecological impacts are limited. However, depending on what level of human cancer risk is determined to be acceptable (and what cancer potency factor is used), wildlife may be a more sensitive endpoint. Ontario has developed a soil concentration limit of 10 ppt TEQ based on protection of wildlife (earthworm eating birds) (Birmingham, 2000). Shrews may be at risk from TCDD and TCDF in land applied paper mill sludges (Abt Assoc., 1994). Wildlife ingest significant quantities of soil (17% of diet for nine banded armadillo, 10% for woodcock, 9% wild turkey) (Beyer et al., 1994). Whether this represents a risk due to sludge application is not clear. Thus risks to predators including raptors higher on the food chain need to be assessed.

Background Exposure
This RA does not assume any background exposure and is only an assessment of incremental exposure from sludge application.
A WHO panel recently completed a reevaluation of PCB/PCDD/F and determined that: 1) a revised TDI of 1 to 4 pg/km body weight is established with 4 considered a maximum for long-term exposure on a provisional basis and 1 as a goal; 2) existing background exposures may be causing subtle effects at current intake levels of 2-6 pg TEQ/kg bw/day; 3) efforts should be made to limit environmental releases to the extent feasible. (WHO, 1998; Brouwer et al., 1998). Thus in an industrialized country like the US, current exposure from background sources may already put us at risk of unknown nature and extent. The RA should include background exposure in the assessment.

Soil concentration
Critical to the assessment of risks from a number of pathways is the soil concentration of dioxins resulting from the application of sludge. The RA calculates this by using models for the loss via vaporization, erosion, leaching and degradation (assumed to be zero for dioxins). Applying the models, a final average concentration of approximately 30 ppt TEQ for agricultural land is calculated for years 26-100 (Exhibit 3-7). It is unclear why this average over years 26-100 was selected. This assumption should be discussed and rationalized. Assessing risks at the end of the 100 year application period would be an equally valid assumption, especially since we will not be able to necessarily identify locations where EQ sludges have been applied.
Soil concentrations resulting from application are a key variable. The approximately 10 fold reduction in concentration predicted in the RA does not seem warranted nor does it seem to agree with much of the literature nor with projections in the Abt Associates document on risks to terrestrial wildlife from papermill sludges (Abt Associates, 1994). Volatilization and leaching are thought by most researchers to be insignificant for dioxins (Fries and Paustenbach, 1990). Even following surface application of sludge, when volatilizations and photodegradation would be greatest, PCDD/Fs were found to be fully persistent (McLachlan et al., 1996). Various field studies show little or no loss of dioxins from soils (Orazio et al., 1992; Hagenmaier et al., 1992 b). Half-life in soil is a critical parameter and for TCDD estimates vary widely and are also very dependent on whether the material is in the surface soil or subsurface (estimates from 9 to 100 years in US Dept. of Health and Human Services, ATSDR, 1998). In contrast, field observations of sludge-borne PCBs show a diminution of concentration, likely due to volatilization (Alcock et al., 1996).
Use of models to predict key transfers of compounds through the agricultural ecosystem, particularly the air to plant transfer, is problematic for a number of reasons (Douben et al., 1997). For some congeners of PCBs, model predictions of volatilization are quite accurate, while for others predicted fluxes underestimated by more than an order of magnitude (Cousins et al., 1997). Since there is significant question about the validity of the models, it has been suggested that use of measured concentrations is a more accurate approach (US EPA, 1994). Such data indicate very low disappearance rates for dioxins. With a half life of over 20 years being realistic (McLachlan, 1997). Research on related chlorinated hydrocarbons further indicates that such compounds become more recalcitrant over time such that the concept of a half life obeying first order kinetics is inappropriate (Linz and Nakels, 1997).
A calculation of soil concentrations of dioxins on agricultural land where no loss through leaching, erosion or volatilization is predicted and where dilution via mixing into the top 15 cm of soil is the only factor diminishing concentration would be an upper bound for soil concentrations. At year 100, given 5 T/ac/yr application of sludges containing 300 ppt TEQ, a concentration of 70 ppt TEQ would result. The average for years 26-100 would be approximately 55 ppt. While not accepting that the average value over years 26-100 is appropriate as the concentration to use in the assessment, the calculated value of 30 ppt in the RA which shows nearly half of the applied contaminants are no longer present seems questionable. This is a place in which the uncertainty of the models is large and the implications for the assessment are also large. While dilution alone may be overly conservative, the models used appear to overestimate losses.

HEI risk
A farm family is the appropriate HEI. Their exposure is potentially higher than others for many of the pathways. They tend to be on-site more, thus having a longer respiratory exposure time (EPA, 1994). They are also likely inhale more particulates and adult soil ingestion might be expected to be relatively high. Farm families also eat significantly more home grown vegetables than other populations (US EPA, 1999).
A key set of parameters is proportion of meat, milk, eggs and poultry ingested which are raised on the farm. Since these are the primary dietary sources of dioxins and since diet is the primary non-occupational exposure route, this is a critical issue for the farm family. The RA states in Exhibit 3-3 that for pathways 4 and 5, the HEI evaluated is the farm household producing a major portion of the animal products they consume and that the animals eat soil and plants grown in soils amended with biosolids. However, the RA does not seem to follow this approach. (This is an example of the lack of transparency in the RA. While Exhibit 3-3 makes the statement above, Exhibit 3-12 shows a very small fraction from sludged sources.)
The first question is how much dairy and meat fat are consumed. No rationale is provided for why the RA uses the mean non-metropolitan intake of beef, lamb and game fat and the 95th percentile value for diary fat. Using figures specific to the HEI farm family would make more sense and be consistent with evaluating risks to the HEI. The Exposure Factor Handbook (US EPA, 1999) shows that farm families consume a mean intake of 2.63g/kg/day of home produced beef and the highest group ingests 8.9 g/kg/day (Table 13-18). At 10% fat content (US EPA, 1999), these levels correspond to a fat consumption of 0.263 g/kg/day (mean) or 0.89g/kg/day (high end), far higher than the 0.0818 g/kg/day in Exhibit 3-12. For dairy, mean consumption for the farm family is 1.7 g/kg/day and the high end is 9.1. With a fat content of 4% this translates into 0.68-0.364 g/kg/day of fat from dairy, significantly lower than the 1.76 g/kg/day used in the RA.
The second question is the percentage of the consumed foods coming from animals exposed to sludge. While Exhibit 3-3 implies that the RA assumed all animals were exposed, this does not seem to be how the RA was actually carried out. Exhibit 3-12 shows that very low estimates for the fraction of consumption derived from sludged soils were used. Contrast the values of 0.097 for beef fat and 0.031 for dairy fat daily ingestion from sludge amended soils used in the RA with 0.44 and 0.40 respectively suggested as home-grown percentages for farm families in the EPA document Estimating Exposure to Dioxin-like Compounds (EPA 1994). These estimates are derived from a USDA farm survey, which while dated (1966), is likely more accurate than figures derived from the general non-metropolitan population (few of whom raise livestock) used in the RA. Data on the proportion of meat and dairy eaten by farm families is provided in tables 13-71 and 13-72 of the Exposure Factor Handbook (US EPA, 1999). The values given are that an average of 32% of all meat, 47% of beef, 25% of all dairy, 21% of eggs and 15% of poultry consumed by farm families are home grown. It is reasonable to assume that all of the home-grown sources would be sludge-exposed.
Not evaluated in the RA is exposure from eggs. Farm families eat a mean of 9 g/kg/day of eggs (US EPA, 1999). At a fat content of 8.4% (US EPA, 1999, this represents 0.75 g/kg/day of fat from eggs, a potentially significant exposure. The farm family also eats an average of 1.54 g/kg/day of poultry (US EPA, 1999) and these may be birds which free-range on sludged lands.
The farm family is also likely to inhale air that contains particulates carrying contaminants as well constituents volatilized from sludge. They may drink water that has received sludge-leached inputs. Their diet may likely include vegetables grown on sludge-amended soils and may include fish caught from ponds or streams impacted by sludge. Thus analysis of simultaneous exposure to multiple pathways is needed.
A revised risk assessment based on farm family health including prenatal and nursing exposure to infants of mothers residing on dairy farms is needed in which higher values are used for the percent of diet from sludged sources as well as simultaneous exposure from multiple pathways.

Risk to children
Given the significance of the soil ingestion pathway which alone represents a cancer risk of 8.3x10-5 according the Exhibit 4-1 in the RA, a more comprehensive analysis of the risks to children is warranted. Since the risks to a child from the soil ingestion pathway alone is nearly one in ten thousand, it is critical that the total risk to which an HEI child might be exposed be evaluated. This should include summing exposure to multiple pathways including: eating home-grown produce, using realistic consumption data for a farm family; soil ingestion; breast feeding for 6-12 months; drinking milk from farm raised cows exposed via pathways 4 and 5; breathing air and inhaling dust; and ingesting water. Dermal exposure and exposure via eating of eggs and poultry should also be evaluated.
The inclusion of an assessment of the exposure of breast-feeding infants is appropriate and even with the limitations in the RA pointed out in this review, that pathway is potentially significant. Average daily intake of dioxins for breast-fed infants on a body weight basis may be almost 1-2 orders of magnitude greater than that of an adult (WHO, 1998). Milk, a food in which bioaccumulated dioxins are shed and thus present in significant quantities is a primary food for infants and children. Over 95% bioavailability of most PCB, dioxin and furans from breast milk has been reported. (McLachlan et al., 1993; McLachlan, 1996; Hoover, 1999). A 90% absorption was assumed in the RA which may thus underestimate the risk. As pointed out above, the risk may also be underestimated since the mother may be exposed via additional pathways not included in path 15 a or b (including the home garden pathway).
Given the significance of the soil ingestion pathway, which alone represents a cancer risk of 8.3x10-5 according the Exhibit 4-1 in the RA, a more comprehensive analysis of the risks to children is warranted. Since the risks to a child from the soil ingestion pathway alone is nearly one in ten thousand, it is critical that the total risk to which an HEI child might be exposed be evaluated. This should include summing exposure to multiple pathways including: eating home-grown produce, using realistic consumption data for a farm family; soil ingestion; breast feeding for 6-12 months; drinking milk from farm raised cows exposed via pathways 4 and 5; breathing air and inhaling dust; and ingesting water. Dermal exposure and exposure via eating of eggs and poultry should also be evaluated.
Current levels of exposure to nursing infants exceeds the 4 pg TEQ/kg level TDI recommended by WHO (WHO, 1998) for nearly the entire population in Canada (Hoover, 1999) ­ and we would expect similar results for infants in the US. Thus background exposures cannot be ignored in assessing risks to children. Most standards are developed based on adult models. Some research has tried to improve on this by incorporating pharmacokinetic adjustments for infants. "The breast-fed infant's intake of organochlorines has been found in general to exceed guidance values, raising the possibility that breast-feeding may pose health risks" (Hoover, 1999, p 528).
The RA should include assessment of non-cancer risks to children, including fetuses. Sensitivity to toxicity of PCDDs may be greater during the fetal/neonatal period than for adults and may have an impact on male reproductive system development (US Dept. of Health and Human Services, ATSDR, 1998). Cognitive functioning in preschool children is negatively impacted by in utero exposure to PCBs and dioxins (Patandin et al., 1999). Birth weight and postnatal growth until 3 months of age were impacted by in utero exposure to PCBs and dioxins (Patandin et al., 1998). Immune suppressive and delayed reproductive effects are also a concern. A probabilistic risk evaluation of organochlorine exposure through breast milk showed that for a significant percentage of the population, PCBs and PCDD/PCDFs provide the greatest concern for non-cancer health effects from chemicals in breast milk (Hoover, 1999).

Exposure through animals ingesting sludge ­ Pathways 4 and 5
The RA underestimates the exposure via pathways 4 and 5. The RA calculations suggest that pathway 5 represents a relatively significant source of exposure, as such it is critical that it be closely analyzed for errors and uncertainties. The assumptions regarding this pathway contain significant flaws.
In assessing the impact of animal diet there are numerous routes of exposure which need to be assessed simultaneously. These include: 1) direct soil ingestion from pasture (the soil ingested should use contaminant concentration in the sludge and not diluted by tillage as discussed below); 2) ingestion of forage which has been grown in sludge-amended pasture and which thus would have contaminant loads due to uptake of volatilized contaminants (the model for this uptake should be based on contaminant concentration in the sludge and not diluted by tillage as discussed below); 3) sludge adhering to the foliage of pasture-forage; 4) ingestion of forage which has been grown on sludge-amended tilled soils and which would thus have contaminant loads due to uptake of volatilized contaminants; and 5) soil contained in harvested forage from tilled soils. These are further discussed below.
As mentioned above in the section on farm family risk, the percentage of dietary intake of meat and dairy coming from on-farm and thus from sludged sources is significantly underestimated. In addition, pathways 4 and 5 for animals ingesting forage and for animals ingesting soil need to be summed since a farmer using sludge would likely use it on pasture and also on field crops (US EPA, 1998). Thus the animals would be exposed both through contaminants in forage (pathway 4) as well as through soil ingested directly (pathway 5). Whether eating forage directly from a pasture or eating harvested forage crops, the animal will be ingesting forage to which sludge should be assumed to have been applied.
In order to appropriately calculate the exposure of these animals, the fraction of forage obtained from harvested crops (pathway 4) and the fraction obtained from eating while in pasture needs to be estimated. Since harvested crops are likely to be grown where sludges have been tilled into the soil, some dilution is expected. For pasture-grown forages, however, the soil is not typically tilled (see discussion below) and thus the plants would be exposed to the full concentration of pollutants in the sludge. This is important because in modeling the uptake of compounds into plants, the concentration of PCBs in the soil surface is an important variable (Cousins et al., 1998). This distinction between harvested and grazed forage consumption is also important in assessing impacts of ingesting sludge which adheres to the plants which is significant for pasture plants (Chaney et al., 1996).
Section 3.3.1 specifically mentions that particulate pollutants are not considered, yet the literature suggests that models used to calculate produce concentrations should include exposure via plants due to small particles . Some research suggests that re-entrainment and deposition of soil particles is a minor contribution (Harrad and Smith, 1997). However, most researchers suggest it is the major pathway for soil-bound PCDD/Fs to aerial plant parts (McLachlan, 1996; Hulster and Marschner, 1995 in McLachlan, 1997; Smith and Jones, 2000).
In rural areas, total dioxin deposition may be 20-40% from re-entrained soil particles (Kao and Venkataraman, 1995 in Smith and Jones, 2000). In rural agricultural areas where soils are tilled and animals disturb the soil, it would seem that particulates are a critical aspect of dioxin exposure. The relative importance of particulate vs. vapor-transfer into plants is different for different congeners (McLachlan, 1997), suggesting that a revised calculation of plant dioxin concentrations needs to take particulate deposition into account using different values for different congeners. The contribution of particulates to total dioxin levels in vegetation was found to be significant, with soil contamination increasingly significant with increasing chlorination levels and with higher levels in soils (Smith and Jones, 2000). For PCBs which are more volatile, this route was less important (Smith and Jones, 2000).
It would also appear that the RA does not take into account the fact that significant amounts of soil are contained in harvested field crops, with several percent soil being estimated in harvested grass silage (Berende, 1990 in McLachlan, 1997). For grasses, it is, and a worst case situation is 100 mg soil g-1 grass DW (Smith and Jones, 2000). This soil which "tags along" with harvested forage needs to be included in the RA pathways 4 and 5. Calculations of the contribution of particulate contamination accounted for 30% of the total PCDD/Fs found in grass containing a conservative background estimate of 20 mg soil g-1 grass DW, while soil was responsible for a majority of the pollutant load under a worst case situation of 100 mg soil g-1 plant DW (Smith and Jones, 2000). PCBs were less related to soil load.
In assessing the impact of soil contamination of plants, it appears that for PCBs and possibly also PCDD/Fs that there is an increasing concentration with decreasing soil particle size. If so, then the use of the bulk soil concentration to calculate the impact of soil in contaminating the vegetation would underestimate soil contribution (Smith and Jones, 2000).
The assumptions made regarding the amount of soil ingested by grazing animals (1.5% of diet is value used in RA for all grazing animals) seems to be based on best management practices. Grazing cattle ingest from 1-18% of their dry matter intake as soils and sheep may ingest as much as 30% depending upon management and the seasonal supply of grass (Fries, 1996; Thornton and Abrahams, 1981). One or 2% intake was a low value obtained when availability of forage was greatest in the spring and an average yearly intake of 6% for cattle was observed when exclusively pasture-fed (Fries, 1996). Other researchers use estimates such as 6% of diet as soil and point out the critical importance of assessing the amount of sludge ingested through material adhering to vegetation ­ and the lack of data on appropriate values for that variable (Wild et al., 1994). A previous EPA document suggests 4% or 0.5kg/day soil ingestion by grazing beef cattle and 3% or 0.4kg/day for dairy cows (US EPA, 1998). Assumptions regarding the percent of time spent in pasture and percent of diet obtained from pasture are important in determining what value to use for soil ingestion. In New York state, dairy experts suggest that cows are in pasture approximately half of the year (Galton, 2000). Use of a probabilistic method for risk assessment would allow a more realistic range of values to be used for these key parameters.
A key parameter is the concentration of pollutant in the soil ingested by the grazing animal. The RA is not clear on what value is used. Again, an example of where the RA needs to be much more transparent about values that are used in calculations. The RA presumably used the values given in Exhibit 3-7 which are the calculated average values for years 26-100 resulting from application of 5T/ac/yr mixed into 15 cm of soil.
Sludge applied to pastures is generally not incorporated into soils yet the risk assessment appears to assume that the sludge is tilled into the soil to the depth of 15 cm. Tillage into pastures is not common practice, hence mixing would be minimal (Fries and Paustenbach, 1990; Wilson et al., 1997). Rather, pastures are top-dressed with sludge and are plowed only every 4 years or less (very rarely in the case of steep, rocky or otherwise marginal lands). This would have a significant bearing on the concentration of sludge-borne contaminants ingested. Since the RA assumes that the animals are ingesting not sludge, but a sludge-soil mix, this is a critical mis-assumption. Even where mixed into soils, the concentration may be greater than that calculated in the RA (see section on soil concentration).
In addition, when sludge is spread on pasture or growing crops, a significant amount ends up adhering to the leaves of plants. This residue is not easily removed by rainfall. Thus it is likely to be an important route of exposure to pastured animals and also to animals eating harvested forage (Wilson et al., 1997). The RA does not seem to account for this exposure. The amount of sludge ingested due not to uptake or soil ingestion but to ingestion of sludge adhering to plants needs to be included in the assessment and this material would contain the full concentration of pollutants contained in the sludge and not be diluted by any soil mixing.
For pathway 4, it is not clear , but it would appear that the RA may have underestimated by a factor of 100. Section 3.3.4 refers to the use of methods presented in section 3.3.1 to determine the concentration of pollutants in the forage. That section makes use of a Vgag, a correction factor of 0.01 which was obtained from the EPA Human Health Risk Assessment Protocol document, Chapter 5 (US EPA 1998). While that document recommends the use of that factor for calculating concentrations in produce, it specifically recommends the use of 1.0, no correction, for pasture grass. This is another example of how the RA is so abbreviated that it is not possible to determine the methods used.

Bioaccumulation through agricultural food chains is a critical aspect of assessing human health risks of PCDD/Fs in sludge application and data are sparse (McLachlan, 1996). In research done on a dairy system where inputs did not include sludge, but only airborne contaminants, it appears that the higher chlorinated compounds become less concentrated as you move up through the agricultural food chain although unfortunately the last link in the chain ­ humans ­ seem to exhibit significant biomagnification (McLachlan, 1996).
The ability to use models to predict final concentrations of dioxin congeners in produce (including meat and milk) varies widely among the individual congeners as shown by a comparison of modeled and observed concentrations, with some of the congeners that comprise the largest share of TEQ contributions in sludge (such as OCDD) found in significantly higher concentrations than predicted (Harrad and Smith, 1997). The bioconcentration factor for various PCB congeners was predicted to vary greatly (by a factor of 32) in a study of the air-milk transfer of PCBs and the predicted values agreed reasonably with measured values (Thomas et al., 1998). Different congeners exhibit very different rates of absorption, metabolism and bioaccumulation. The higher chlorinated congeners are generally less well absorbed (McLachlan and Richter, 1998).
The concentrations of PCDD/Fs and PCBs varies with different plant species exposed at the same sites. For relatively non-volatile compounds, the variability is generally less than a factor of 4 and can be explained by plant surface area/volume, while for more volatile compounds (like PCBs), interspecies variability of 30 fold or more was observed (Bohme et al., 1999). This variation may be important in assessing the exposure through pathways 2 and 4.
In assessing uptake of dioxins into animals, the duration of the experiments is important. For dairy cows, a steady state concentration is predicted due to the excretion of dioxins in milk after 9-11 weeks (Heeschen et al., 1994 in McLachlan and Richter, 1998) while in beef cattle concentrations will increase (although dilution due to increasing biomass is important to consider) (McLachlan, 1997). In short term experiments of less than 6 months duration, equilibrium may not have been reached (Fries, 1996). The amount absorbed by cows is a critical parameter and it seems to be very sensitive to the Kow of the compound and these values are very uncertain (McLachlan, 1997). The suggestion that absorption of PCDD/Fs are inhibited by being in a sludge matrix does not seem to be justified by experimental data (McLachlan and Richter, 1998).

Soil Ingestion
This pathway is identified as the one posing the greatest cancer risk (Exhibit 4-1). As such, it is one that should be closely examined. Unfortunately the RA gives it only a brief consideration. No rationale is provided for some of the assumptions made nor is it clear what value is used for the critical soil concentration parameter Cj.
The RA has modified the soil ingestion pathway analysis from the Round I assessment. The quantity of soil ingested has been raised from 200 to 400 mg/day. At the same time the assumption that the soil ingested is pure sludge has been changed to ingestion of a soil/sludge mixture. The concentration of that mixture is not specified (at least not that I could find). Does the RA assume that the exposure is to a sludge/soil mix containing the ~30 ppt concentration calculated in Exhibit 3-7? If so, that is likely to be an underestimate even for sludges incorporated into agricultural fields (see section on soil concentrations). It might greatly underestimate concentrations for sludge products which might be applied to home gardens (like composts) and not tilled.
It is not only children who ingest soil. Adults ingest soil as well and teenagers may ingest more than adults. As recognized in Estimated Exposure to Dioxin-like Compounds (EPA, 1994), failure to include soil ingestion by older children and adults may underestimate lifetime soil ingestion. Since this pathway represents the highest risk, it would be important to include estimates of lifetime exposure via soil ingestion. While I did not find literature on it, it would be reasonable to suggest that farmers and home gardeners ingest higher quantities of soil than average.

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