D.6.0 ECOLOGICAL RISK ASSESSMENT METHODOLOGY AND RESULTS
D.6.1 INTRODUCTION
This section summarizes the methodology and results of the ecological risk assessment (risks to plants and animals from potential exposure to radioactive and toxic contaminants) for the various TWRS alternatives. Potential ecological risks are evaluated under baseline conditions (i.e., the No Action alternative) with ecological impacts from other alternatives being compared to the baseline impacts. The No Action alternative is a conservative and bounding scenario since it assumes that all of the tank waste would remain in-place and would be available for direct contact and potential migration to groundwater and the Columbia River. Consequently, the No Action alternative represents the greatest potential impacts to ecological receptors (terrestrial and aquatic).
Under baseline conditions, radiological doses and chemical hazards were estimated for potential ecological receptors from 1) direct contact with tank waste; 2) exposure to tank waste contaminants in groundwater that reaches the Columbia River; and 3) exposure to routine contaminant releases to the air. For other alternatives (e.g., in situ and ex situ alternatives described previously), potential ecological risks were estimated from radionuclides and chemicals released to the air during remediation activities.
The ecological risk assessment methodology is conceptually identical to the methodology used to estimate potential human health risks. All chemicals of concern for human health were also considered chemicals of concern for potential ecological receptors. Consequently, the ecological risk assessment used the same source terms and contaminant transport data that were used in the human health risk assessment (Section D.2.0). The URF approach developed for human health risk was followed for terrestrial receptors except that ecological species-specific factors were substituted for the human land use-specific factors and are described in more detail in Section D.6.3.2. Potential radiation doses to aquatic organisms from tank waste contaminants calculated to reach the Columbia River by groundwater migration were evaluated using the CRITRII model (Baker-Soldat 1992).
The ecological risk assessment in this EIS follows the approaches recommended in EPA's Framework for Ecological Risk Assessment (EPA 1992) and the Hanford Site Baseline Risk Assessment Methodology (DOE 1993d). The basic components of this ecological risk assessment are 1) problem formulation; 2) characteriz ation of potential exposures; 3) estimation of potential ecological impacts from radionuclides and toxic chemicals; and 4) summariz ation of the risk assessment results (EPA 1992).
D.6.2 PROBLEM FORMULATION
This section describes the ecosystem potentially at risk, potential ecological effects of the contaminants of concern, endpoints selected for risk assessment, and the conceptual model.
D.6.2.1 Ecosystems Potentially at Risk
The Hanford Site supports a variety of arid terrestrial habitats, a major aquatic habitat in the Columbia River, and a number of threatened, endangered, or candidate species, as described in Volume Five, Appendix I. The primary ecosystems potentially at risk from exposure to tank waste include the shrub-steppe habitat in and immediately adjacent to the Central Plateau; mobile organisms that may enter the area (for example, birds and deer); and aquatic wildlife in the Columbia River.
D.6.2.2 Ecological Effects
To date, no specific ecological effects of exposure to tank waste have been documented. The waste is in tanks buried in the ground (i.e., 4.6 m [15 ft] below the ground surface), which limits potential contact with any leaking waste to deep-rooted plants and burrowing animals. The areas adjacent to the tanks are highly disturbed, kept clear of vegetation, and represent low quality habitat thereby further limiting organisms' access to the waste. No current ecological risk exists since there is no complete exposure pathway for the tank waste. Any potential ecological effects would occur in the future following the loss of institutional controls. Natural succession and potential failures of tanks could increase the likelihood of contact with the waste. The direct ecological effects of concern at this time would be radiation and toxic chemical exposures that could lead to individual mortality, reproductive and developmental effects, and a variety of potential indirect effects on other ecological variables. Examples of potential indirect effects include decreased biodiversity, habitat loss or alteration, and impacts on productivity and nutrient turnover. As described in the following sections, this screening level assessment focuses on radiation doses and chemical intakes in individual indicator organisms.
D.6.2.3 Endpoint Selection
Human health risk assessment typically focuses on two well-defined endpoints associated with the health of individual humans, cancer incidence, and the noncancer effects of hazardous chemicals. However, ecological risk assessment is concerned with many species and attributes of ecosystems other than their species composition, such as nutrient turnover rates, energy flow, and food web complexity. Particular endpoints must therefore be chosen for each new ecological risk assessment.
D.6.2.3.1 Assessment Endpoints
Assessment endpoints are the specific ecological characteristics to be protected (EPA 1992, Suter 1993). For purposes of this EIS, the primary assessment endpoint for the effects of radionuclides and hazardous chemicals is prevention of the adverse effects of these substances on any ecological receptors. A second, more specific, endpoint is prevention of adverse effects on Federal or Washington State species of concern that may occur in the TWRS area. These species, described in Section 4.4 and Volume Five, Appendix I, include Piper's daisy (Erigeron piperianus), the sage sparrow (Amphispiza belli), Swainson's hawk (Buteo swainsoni), and the loggerhead shrike (Lanius ludovicianus).
D.6.2.3.2 Measurement Endpoints
Measurement endpoints are characteristics that are subject to measurement and correspond in some way to the assessment endpoints. The measurement endpoints chosen to correspond to the assessment endpoints are 1) estimated radiation doses to terrestrial organisms compared with the 0.1 rad/day expected to have no adverse effects (IAEA 1992) (this screening radiological value is intended to be protective of chronic reproductive and developmental effects for a wide range of terrestrial species and is not specific for any one species); 2) the ratio of estimated hazardous chemical intake by terrestrial organisms to the intake expected to have no adverse effect (HI value greater than 1.0 indicates a potential for adverse effects); and 3) estimated radiation doses to aquatic organisms compared with the 1.0 rad/day expected to have no adverse effects (NCRP 1991).
D.6.2.4 Conceptual Model
The primary objective of the conceptual model is to develop a series of working hypotheses about how contamination may impact the ecological components of the natural environment (EPA 1992). For purposes of this EIS, these hypotheses center on potential exposures of individual organisms to radiation and hazardous chemicals.
The conceptual model for terrestrial organisms is a flow diagram illustrating potential complete pathways for movement of tank waste or radiation to a selected suite of representative species (Figure D.6.2.1). The representative species included a generic plant, the great basin pocket mouse (Perognathus parvus), the coyote (Canis latrans), the mule deer (Odocoileus hemionus), the red-tailed hawk (Buteo jamaicensus), and the loggerhead shrike (Lanius ludovicianus). The exposure pathways considered were food, soil, and water ingestion; inhalation; and direct radiation. This model is designed to assess effects at several trophic levels such as the primary producer, herbivore, and mammalian and avian carnivores while being simple enough to efficiently assess potential effects at the waste sites within the scope of this EIS. The species chosen are all known to occur on the Hanford Site, and all of them could potentially be exposed to tank waste constituents at some future time.
Figure D.6.2.1 Conceptual Model of Potential Exposure of Ecological Receptors to Hanford Site Waste
As illustrated in Figure D.6.2.1, the pocket mouse serves as a vector for contaminant movement through the food chain from plants to mammalian and avian carnivores. Because the mouse has no requirement for drinking water and obtains all its water from food, it would be subject to impacts from radiological and nonradiological chemicals in soil and food and to direct radiation while in burrows. Its small home range would cause it to spend all its time within a contaminated area and obtain all its food there (Table D.6.2.1).
The mule deer has a wider home range than the mouse, requires water, and consumes small amounts of soil while grazing, allowing some direct exposure to contaminants unmodified by plant uptake (Table D.6.2.1; Arthur-Alldredge 1979). The fraction of contaminated plants consumed was set equal to the ratio of the grid cell area to the home range (100 hectare [ha]/1,240 ha = 0.008).
The coyote is a mammalian predator, requires water, and was assumed to consume only pocket mice as prey for purposes of this assessment. The fraction of contaminated prey consumed was set equal to the ratio of the grid area to the home range (100 ha/302 ha = 0.33).
The red-tailed hawk is an avian predator with a wide home range, requires water, and is assumed to consume only pocket mice as prey for purposes of this assessment. The fraction of contaminated prey consumed was set equal to the ratio of the grid cell area to the home range (100 ha/218 ha = 0.46). Potential effects on the red-tailed hawk also serve as measurement endpoints for effects on other raptors of concern such as the Swainson's hawk, for which relevant data are not available.
The loggerhead shrike is a passerine (songbird) bird species that is much smaller than the red-tailed hawk and has a smaller home range. The shrike feeds on insects, small mammals, and other birds (Fitzner-Rickard 1975). For purposes of this EIS, the shrike was assumed to consume only pocket mice as prey. Its small home range would cause it to spend all its time within a contaminated area and obtain all its food there (Table D.6.2.1).
The CRITRII model was used to estimate radiation doses to aquatic organisms (Baker-Soldat 1992). That model uses a simple food chain and bioaccumulation factors to estimate internal and external radiation doses to algae, fish, crustaceans, mollusks, and muskrats, raccoons, herons, and ducks feeding on aquatic organisms.
D.6.3 ANALYSIS
The analysis phase of an ecological risk assessment consists of technically evaluating data for potential exposures to and effects of the contaminants (EPA 1992). This section describes how the exposures were estimated for each representative receptor of concern.
D.6.3.1 Source Terms and Direct Exposure
The source terms were the same as those used for the human health risk assessment. Constituent concentrations for direct exposure to tank waste were estimated from waste inventory data and volumes (WHC 1995g and Jacobs 1996), assuming an average density of 1.5 kg/L. Air concentrations for the No Action and remediation alternatives were estimated from average annual routine emissions and the minimum and maximum onsite Chi/Q values. Because ecological receptors would not have access to groundwater unless it reached the surface, water concentrations used were the minimum and maximum calculated (i.e., modeled) concentrations in groundwater reaching the Columbia River at 300, 500, 2,500, 5,000, and 10,000 years. Use of the maximum modeled concentrations provides conservative, upper-bound estimates of exposure point concentrations and potential exposures.
D.6.3.2 Characterization of Exposure
This section describes the general methods used to estimate the intake of hazardous chemicals, the associated HIs, and radiation doses resulting from radionuclide intake by terrestrial organisms. The section first describes the equations used as they are typically presented in the risk assessment literature and then describes how the equations were modified to calculate URFs to simplify computation. Strictly speaking, the "URFs" as applied to ecological receptors are unit dose or HI factors, in that the result is an estimated radiation dose or chemical HI, rather than a probability of some adverse effect. However, the term URF is maintained here for purposes of consistency with the methodology used for the human health risk assessment.
D.6.3.2.1 Estimation of Hazardous Chemical Intake
Uptake of contaminants from soil by a generic plant was estimated by multiplying the soil concentration by the soil-to-plant concentration factors used in the GENII model at the Hanford Site (Table D.6.3.1 and D.6.3.2).
Table D.6.3.1 Transfer Factors Used to Estimate Radiation Doses to Ecological Receptors
Table D.6.3.2 Properties of Chemicals Used to Estimate Hazard Quotients
The equation is:
(1) Cvi = (Csi)(Bvi)(0.4)
| Where: | |||
| Cvi | = | Contaminant concentration in plant, mg kg-1 wet weight | |
| Csi | = | Contaminant concentration in soil, mg kg-1 dry weight | |
| Bvi | = | Soil-to-plant concentration factor (unitless) (The factor for grain concentration was used for the pocket mouse, which is assumed to consume seeds. The vegetative portion values were used for the mule deer.) | |
| 0.4 | = | Dry weight/wet weight conversion (DOE 1994) |
The intake rate of hazardous chemicals for a herbivore via consumption of plants is typically calculated as:
(2) Ii = (Cvi)(IR)(FI)/(BW)
| Where: | |||
| Ii | = | Intake rate of the ith contaminant, mg kg-1 day-1 | |
| Cvi | = | Contaminant concentration in plant, mg kg-1 | |
| IR | = | Ingestion rate of food, kg day-1 wet weight | |
| FI | = | Fraction ingested from contaminated source, unitless | |
| BW | = | Body weight, kg wet weight |
Consumption rates by carnivores are calculated similarly, substituting the contaminant concentrations in the herbivore for the concentrations in plants. Contaminant concentrations in herbivore muscle are typically estimated using the equation:
(3) Cmi = (Cvi)(IR)(FI)(Bmi)
| Where: | |||
| Cmi | = | Contaminant concentration in muscle, mg kg-1 wet weight | |
| Cvi | = | Contaminant concentration in plant, mg kg-1 wet weight | |
| IR | = | Ingestion rate of plants by herbivore, kg day-1 | |
| FI | = | Fraction ingested from a contaminated source, unitless | |
| Bmi | = | Plant-to-muscle transfer factor, day kg-1 |
However, as described in the following text, radionuclide body burdens were estimated from element-specific fractions retained, biological half-lives, and radiological half-lives. Therefore, for purposes of consistency, nonradiological body burdens were estimated in the same way, assuming an infinite radiological half-life. The resulting equation is:
(4) Cmi = [(Cvi)(IR)(FI)(FR)(Bi)]/BW
| Where: | |||
| Cmi | = | Contaminant concentration in muscle, mg kg-1 wet weight | |
| Cvi | = | Contaminant concentration in plant, mg kg-1 wet weight | |
| IR | = | Ingestion rate of plants by herbivore, kg day-1 | |
| FI | = | Fraction ingested from a contaminated source, unitless | |
| FR | = | Fraction retained (Baker-Soldat 1992) | |
| Bi | = | Effective half-life (days), calculated as described in Baker and Soldat (Baker-Soldat 1992); assuming radiological half-life to be infinite reduces it to the biological half-life. |
This equation assumes that the body burden is at steady state following chronic intake by a secondary receptor.
Food ingestion rates and body weights used in estimating exposures for this EIS are listed in Table D.6.2.1. Intakes via inhalation and water ingestion were estimated following procedures recommended in EPA (EPA 1993) when species-specific values were not available (Table D.6.2.1).
D.6.3.2.2 Calculation of Hazard Indices
The HI, the ratio of estimated intake to that expected to have no adverse effect, is typically calculated as:
HI = I/NOAEL
Where I is calculated as described in equation (2), and the No Observed Adverse Effect Level (NOAEL) is obtained from the literature as described in the following text. Both are expressed as mg per kg body weight per day.
An HI greater than 1.0 for a given chemical indicates that the estimated intake exceeds the threshold level and adverse health effects may occur. An HI less than 1.0 is indicative of no adverse impacts. For sites with multiple chemicals, the HIs may be summed, making the assumption that the modes of action and target organs of the chemicals are similar. Thus, a site may be said to present a hazard if the sum of the HIs exceeds 1.0, even if the individual chemical HIs are less than 1.0. URFs were estimated to allow calculation of the HIs directly from media concentrations, without the necessity of separate calculations of uptake at each trophic level. This consists of simply combining all the variables except the medium concentration for each constituent of concern for each organism. URFs for food ingestion and water ingestion are summarized in Tables D.6.3.3 and D.6.3.4, respectively.
Table D.6.3.3 Food Ingestion Unit Risk Factors, Chemicals
Tables D.6.3.4 Water Ingestion Unit Risk Factors, Chemicals
NOAELs were obtained from a variety of sources, with Opresko et al. (Opresko et al. 1994) as the primary source (Table D.6.3.5). Wildlife NOAELs for test species other than those of interest here were scaled to the body weight of the organism using the equation:
(5) NOAELy = (NOAELx)[(bwx)/(bwy)]1/3
| Where: | |||
| NOAELy | = | NOAEL for the organism of interest | |
| NOAELx | = | NOAEL for experimental animal available from the literature | |
| bwy | = | Body weight of the organism of interest | |
| bwx | = | Body weight of experimental animal with the known NOAEL (Table D.6.3.6) |
Table D.6.3.5 Ingestion No Observed Adverse Effect Levels Used to Estimate Hazard Quotients
Scaling factors estimated according to Equation (5) are summarized in Table D.6.3.6. NOAELs for plants (Table D.6.3.5) were obtained as benchmark soil concentrations from Will and Suter (Will-Suter 1994), and the vegetation HIs were calculated as the waste unit soil concentration divided by the NOAEL.
Table D.6.3.6 Scaling Factors for Extrapolating No Observed Adverse Effect Levels Between Species
D.6.3.2.3 Estimation of Radiation Doses
Radiation doses to ecological receptors were calculated using URFs analogous to those for chemicals. The basic equation used to estimate radiation dose to the pocket mouse was as follows:
(6) Dose rate (rad d-1) = [(CS)(PS)(WW)(Qv)(FI)(EF)(ED)(FR)(Bi)(Ei)(1 y/365 d)]/[(BW)(AT)]
| Where: | |||
| CS | = | Radionuclide concentration in soil, Ci/kg | |
| PS | = | Soil-to-plant transfer factor | |
| WW | = | Wet-to-dry weight conversion factor, 0.4 | |
| Qv | = | Ingestion rate, kg/day | |
| FI | = | Fraction ingested from contaminated source | |
| EF | = | Exposure frequency, 365 day/year | |
| ED | = | Exposure duration, 1 year | |
| FR | = | Fraction retained (Baker-Soldat 1992) | |
| Bi | = | Effective decay constant of the radionuclide (days), calculated as described in Baker and Soldat (Baker-Soldat 1992); takes both radioactive decay and biological turnover into account. | |
| Ei | = | Effective energy absorbed, (5.12 · 104 kg rad Ci-1 d-1 MeV-1 dis) (MeV dis-1), using MeVs obtained from Baker and Soldat (Baker-Soldat 1992) | |
| BW | = | Body weight, kg | |
| AT | = | Averaging time, 1 year |
The doses to predators were calculated similarly, substituting the concentration in the mouse for that in the plant. Radionuclide properties and transfer factors used in the calculations are listed in Tables D.6.3.7 and D.6.3.1, respectively. URFs were estimated to allow calculation of doses directly from media concentrations, without the necessity of separate calculations of uptake at each trophic level. URFs for food ingestion are summarized in Table D.6.3.8. Radiation doses were calculated as the product of the URF and the medium concentration.
Table D.6.3.7 Radionuclide Properties Used to Estimate Radiation Doses to Ecological Receptors
Table D.6.3.8 Food Ingestion Unit Risk Factors, Radionuclides
Doses resulting from ingestion of water, ingestion of soil, and inhalation were estimated in the same way, substituting the appropriate intake rates for the food ingestion rate and are summarized in Tables D.6.3.9, D.6.3.10 and D.6.3.11, respectively. As noted in Table D.6.2.1, inhalation and water ingestion rates were estimated using equations from the EPA (EPA 1993) when species-specific values were not available.
Table D.6.3.9 Water Ingestion Unit Risk Factors, Radionuclides
Table D.6.3.10 Soil Ingestion Unit Risk Factors, Radionuclides
Table D.6.3.11 Inhalation Unit Risk Factors, Radionuclides
Doses to pocket mice and plants via direct radiation were calculated using the equation:
(7) Dose rate (rad day-1) = [(24)(2.12)(E)(C)]/p (Jacobs 1996)
| Where: | |||
| 24 | = | h/d | |
| 2.12 | = | Constant to convert units to rad h-1 | |
| = | (U)(V)(W)(X)(Y)(Z), dis-rad-g /Ci-hr-MeV | ||
| Where: | |||
| U | = | 1 Ci/106 Ci | |
| V | = | 3.7 1010 disintegrations/Ci-sec | |
| W | = | 3600 sec/hour | |
| X | = | 106 eV/MeV | |
| Y | = | 1.6 10-12 erg/eV | |
| Z | = | 1 rad-g/100 ergs | |
| E | = | Average gamma energy per disintegration, MeV/dis | |
| C | = | Radionuclide concentration in soil, uCi/cm3 | |
| p | = | Soil density, g/cm3 |
URFs for direct exposure are listed in Table D.6.3.12.
Table D.6.3.12 Direct Radiation Unit Risk Factors, Radionuclides
D.6.4 RESULTS
Results are summarized in Tables D.6.4.1 through D.6.4. 11 . Overall, the results of this screening analysis fall into two extreme classes. Direct contact with waste, which would be unlikely even under the No Action alternative, is estimated to result in radiation doses that would likely be lethal in a short time (Table D.6.4.1). The chemical hazards associated with direct exposure to tank waste, while less dramatic, are still estimated to be up to several orders of magnitude higher than the 1.0 HI benchmark for concern (Table D.6.4.4). Any direct effects on individual organisms exposed to stored waste could lead to a variety of indirect effects on the ecosystem, including decreased biodiversity, habitat loss or alteration, and impacts on productivity and nutrient turnover. Exposure to routine air emissions under the No Action alternative is estimated to result in a radiation exposure far below background levels (Table D.6.4.2). Exposure to contaminated groundwater reaching the Columbia River is not estimated to result in radiation doses approaching the 0.1 rad/day benchmark for terrestrial organisms (IAEA 1992) (Table D.6.4.3). Likewise, maximum radiation doses to aquatic organisms in the Columbia River, 300 or 500 years in the future, are well below the 1.0 rad/day benchmark for aquatic organisms (NCRP 1991) (Table D.6.4.6). Because the direct impacts of air and groundwater exposure are expected to be small, any associated indirect impacts on the ecosystem would be correspondingly minor.
Table D.6.4.8 Estimated Radiation Doses from Inhalation of Routine Releases, In Situ Alternatives
Table D.6.4.5 presents the maximum HIs associated with ingestion of groundwater calculated to reach the Columbia River under the No Action alternative. For concentrations of contaminants calculated to reach the Columbia River 300, 500, 2,500, 5,000, and 10,000 years in the future, the maximum HIs for the coyote, mule deer, red-tailed hawk and loggerhead shrike were all well below the HI criterion of 1.0. The ecological hazards were based on a conservative, bounding scenario involving consumption of groundwater contaminants at the point where groundwater daylights on the Columbia River bank (e.g., springs or seeps) and assumes no dilution of the groundwater contaminants by the river before the receptors have access to it. Based on the conservative nature of the exposure scenarios, the estimated hazards for the representative species indicate that no adverse effects would be expected for terrestrial receptor s consuming groundwater in the future. Consequently, no indirect ecosystem impacts would be anticipated from future groundwater consumption.
The only radiation or chemical exposures evaluated for ecological receptors during remediation were radiation doses associated with routine releases during tank waste remediation. No estimated radiation doses resulting from routine releases during the in situ or in situ/ex situ combination alternatives exceeded the 0.1 rad/day benchmark suggested by IAEA (IAEA 1992) for ecological impacts (Tables D.6.4.8 and D.6.4.11 ). For the Ex Situ Intermediate Separations and Phased Implementation alternatives, the maximum estimated radiation doses resulting from routine releases exceeded this benchmark because of C-14, Cs-137, I-129, and Sr-90 releases (Table D.6.4.7). Exposures exceeding 0.1 rad/day also would be expected under the Ex Situ No Separations and Ex Situ Extensive Separations alternatives (Table D.6.4.9). However, exceeding the 0.1 rad/day benchmark assumes long-term exposure at the location of the maximum Chi/Q. It is unlikely that any ecological receptor would spend all of its lifetime at this location of highest exposure. The exposure at the location of the minimum Chi/Q would be approximately 100,000 times lower. It is therefore considered unlikely that ecological receptors would be exposed to harmful levels of airborne radiation resulting from routine releases under any alternative. Corresponding indirect impacts on the ecosystem would be similarly unlikely.
D.6.5 UNCERTAINTY
The greatest uncertainty in calculating both the HIs and radiation doses was associated with the source data. Source terms are based on estimated inventories and, for radionuclides, subsequent decay. Additional or better source data could either increase or decrease the estimated hazards. Secondary contributors to uncertainty are the transfer factors used to estimate plant uptake and assimilation in the mouse. Additional data on these factors could either increase or decrease the estimated hazards. Additional likely secondary contributors to uncertainty are the NOAELs for chemical hazard and the water ingestion and inhalation rates. The CRITRII model (Baker-Soldat 1992) was used only for estimating maximum radiation doses to aquatic organisms exposed to groundwater entering the Columbia River at 300 to 500 years. These estimates were all lower than one millionth of a rad per day, the benchmark recommended by NCRP (NCRP 1991) as protective of aquatic organisms. It is unlikely that detailed uncertainty analysis would alter the conclusion that groundwater risks are very low. Additional discussion of the uncertainties in the ecological risk assessment is provided in Volume Five, Appendix K.
D.6.6 DERIVATION OF ECOLOGICAL NO OBSERVED ADVERSE EFFECT LEVELS
This section describes the derivation of those NOAELs not taken directly from Opresko et al. (Opresko et al. 1994) or DOE (DOE 1994). Table D.6.3.5 lists all the NOAELs used in this document.
D.6.6.1 Boron in Birds
According to Smith and Anders (Smith-Anders 1989), 30 mg/kg of boron in the diet substantially reduced weight gain in ducklings. Control ducklings weighed 36.2 g, (N = 23, SD = 0.7). 30 mg/kg of boron on a fresh weight basis was 35 mg/kg on a dry weight basis. Consider this portion of the study a subchronic study because it was less than 10 weeks, although the adult feeding portion included reproduction. Consider the 30 mg/kg in diet to be a subchronic lowest observed effect level (LOEL). Feeding rates of adult mallards were 222, 184, and 209 g food/day in feeding trials. The mean equals 205 g/day. Male adults weigh approximately 1.3 kg and females approximately 1.1 kg (Table 4 in Smith-Anders 1989). The mean duck weight is thus 1.2 kg. If a 1.2-kg adult consumes 205 g/day, a 36-g duckling is assumed to consume (36/1200) · 205 = 6.15 g/day.
[(30 mg B/ kg food) · (6.15 g food/day) · (1 kg/1000 g)]/0.036 kg body weight)] = 5.125 mg/kg/day as a subchronic LOEL.
5.125 mg/kg/day · 0.1 = 0.5125 mg/kg/day as a chronic LOEL for a 36-g mallard duckling, following the extrapolation suggested by Opresko et al. (Opresko et al. 1994).
D.6.6.2 Boron in Mammals
Table 9 in Eisler ( Eisler 1990) states that rats fed 350 or 525 mg B/kg diet as borax or boric acid for 2 years had no observable effects on fertility, lactation, litter size, weight, or appearance. Using the rat weight of 0.35 kg from Opresko et al. (Opresko et al. 1994), estimate the food intake rate from EPA (EPA 1993) Equation 8 as 0.017 kg/day. Assume that 350 mg B/kg diet dry weight is a chronic NOAEL. Then, [(350 mg/kg food) · (0.017 kg/day)]/0.35 kg body weight = 17 mg/kg/day as a chronic NOAEL for a rat.
D.6.6.3 Cerium
The Hazardous Substance Data Bank (HSDB), May 1995 NTOX entry, reproduced in the following text, states that cerium compounds are nontoxic when ingested.
| 1 - HSDB
NAME - DICERIUM TRIOXIDE RN - 1345-13-7 NTOX - INSOL CERIUM COMPD, SUCH AS THE OXIDES, ARE NONTOXIC WHEN INGESTED ORALLY... /CERIUM OXIDES/ [VENUGOPAL. METAL TOX IN MAMMALS 2 1978 , p. 151] **PEER REVIEWED** |
D.6.6.4 Chromium in Birds
Rosomer et al. (Rosomer et al. 1961), as cited in Driver (Driver 1994) state that "chickens appear to be resistant to hexavalent chromium since exposure to 100 ppm in the diet did not cause any adverse effects." The title of the Rosomer article indicates a "growing chick." Assume that 100 ppm CrVI is a reasonable subchronic NOAEL for chicks (not adults). Using the body weight (BW) and food consumption rate (IR) from Opresko et al. (Opresko et al. 1994), Page A-24: BW = 0.534 kg, IR = 0.044 kg/d, 100 ppm = 100 mg/kg, then [(100 mg/kg food) · (0.044 kg food/day)]/0.534 kg BW = 8.24 mg/kg/day as a subchronic NOAEL.
It can be extrapolated from the subchronic value suggested by Opresko et al. (Opresko et al. 1994) to arrive at (8.24) · (0.1) = 0.824 mg/kg/day as a chronic avian NOAEL.
D.6.6.5 Molybdenum
Table 4 of Eisler (1989) states that female mule deer had no effects after 33 days on a diet of up to 200 mg Mo/kg in their feed. Assume this value is an acceptable subchronic NOAEL. Then, in a like manner, one can extrapolate a chronic NOAEL as follows:
| (200 mg/kg food) · (37 kg food/day) | · (0.1) = 1.3 mg/kg/day |
| 57 kg Bw |
D.6.6.6 Nitrite
The reference for this HSDB entry from May 1995 is reproduced in the following text. The test species is a rat with a body weight of 0.35 kg (EPA 1988) and a water ingestion rate of 0.046 L/day (EPA 1988, Table 1-4). The study duration is three generations and resulted in a 100 mg/kg/day NOAEL, considered chronic due to the length of the study. The nitrite portion of sodium nitrite = (100-33.32) = 66.68 percent 100 · 0.6668 = 67 mg/kg/day. The final NOAEL is thus 67 mg/kg/day.
| 1 - HSDB
NAME OF SUBSTANCE SODIUM NITRITE CAS REGISTRY NUMBER 7632-00-0 |
| NONHUMAN TOXICITY EXCERPTS
... RATS RECEIVED SODIUM NITRITE AT 100 MG/KG IN DRINKING WATER DAILY DURING THEIR ENTIRE LIFE SPAN OVER THREE GENERATION; NO EVIDENCE OF CHRONIC TOXICITY, CARCINOGENICITY, OR TERATOGENICITY ... FOUND. [NRC. DRINKING WATER & HEALTH 1977 , p. 420] **PEER REVIEWED** |
D.6.6.7 Silver
The reference for this IRIS entry from May 1995 is reproduced in the following text. The test species is a human with a body weight of 70 kg (EPA 1989). The study duration was more than 2 years. The effect endpoint was argyria (skin discoloration) and the exposure route was oral and injection in medication at various dosages. Consider the reported NOAEL of 0.014 mg/kg/day to be a chronic NOAEL. Then the final NOAEL is 0.014 mg/kg/day.
| 1 - IRIS
NAME OF SUBSTANCE Silver CAS REGISTRY NUMBER 7440-22-4 ------------------------------------------------------------------------------- REFERENCE DOSE FOR ORAL EXPOSURE | |||||
| ORAL RFD SUMMARY | |||||
| Critical Effect -------------------- Argyria |
Experimental Doses* ----------------------- NOEL: None |
UF ----- 3 |
MF --- 1 |
RfD --------- 5E-3 mg/kg/day |
|
| 2- to 9-Year | LOAEL: 1 g (total dose); Human i.v. Study converted to an oral dose of 0.014 mg/kg/day (Gaul-Staud 1935). | ||||
| ---------------------------------------------------------------------------- | |||||
| *Conversion Factors: Based on conversion from the total i.v. dose to a total oral dose of 25 g (i.v. dose of 1 g divided by 0.04, assumed oral retention factor; see Furchner et al., 1968 in Additional Comments section) and dividing by 70 kg (154 lb) (adult body weight) and 25,500 days (a lifetime, or 70 years). | |||||
D.6.6.8 Tungsten
The reference is an HSDB entry from May 1995, reproduced in the following text. The test species is a rat with a body weight of 0.35 kg (EPA 1988) and a food ingestion rate of 0.017 kg/day, calculated using Equation 3-8, for rodents (EPA 1993). The study duration was 70 days. Two percent in the diet is considered a subchronic NOAEL, with the effect endpoint being growth rate and the exposure route being ingestion. The calculations are as follows:
([0.02] [0.017 kg/day][106 mg/kg])/0.35 kg = 971 mg/kg/day for a subchronic NOAEL. Multiply by 0.1 to get a chronic NOAEL of 97 mg/kg/day (Opresko et al. 1994).
| NTOX - TUNGSTEN METAL POWDER FED 70 DAYS TO WEANLING RATS...@ LEVELS 2, 5, & 10 PERCENT OF DIET...RESULTED IN NO EFFECT ON GROWTH RATE OF MALE RATS BUT CAUSED 15 PERCENT REDN IN WT GAIN IN FEMALES FROM THAT OF CONTROLS. PARTICLE SIZE...NOT REPORTED. [PATTY. INDUS HYG & TOX 2ND ED VOL2 1963 , p. 1162] **PEER REVIEWED** |
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