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Dioxin-like Compounds: Definitions, Sources, Exposures, Effects, and TEFs

1 Dioxins/Furans-An Overview

1.1 What are Dioxins/Furans?

Polychlorinated dibenzo-p-dioxins (PCDDs) and -furans (PCDFs) are two related classes of chlorinated organic compounds. They have similar core structures that can be visualized as two 6-sided benzene rings connected by two oxygen bridges in PCDDs and one in PCDFs; the second bridge in PCDFs is a carbon-carbon bond (Figure 1). There are 8 different positions on a PCDD molecule and 10 on a PCDF molecule, that can be occupied bya chlorine atom or other substituent. This makes possible the existence of 75 individual variations or "congeners" of PCDDs and 135 of PCDFs. The only difference between these various congeners of PCDDs and PCDFs is the specific number and location of the chlorine atoms in each. Different congeners that share the same number of chlorine atoms, but at different locations, are referred to as isomers. Groups of isomers that contain 1, 2, 3, 4, 5, 6, 7, or 8 chlorine atoms are called mono-, di-, tri-, tetra-, penta-, hexa-, hepta-and octa-chlorinated dioxins/furans, respectively (ATSDR, 1998).

Figure 1. Core Structures of Dioxins and Furans

The relative toxicity or potency of various PCDDs and PCDFs is strongly influenced by the number and position of the chlorine atoms in the molecule. The most toxic dioxin, and the most extensively studied, is 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). More highly chlorinated (i.e., penta- through octa-) PCDDs/PCDFs that also have chlorine atoms at the (lateral) 2, 3, 7 and 8 positions (among others) are often described as "dioxin-like compounds" in recognition of the possibility that they may share, to some extent, the established toxicities of 2,3,7,8-TCDD (ATSDR, 1998).

1.2 What are the Sources of Dioxin Contamination and Exposure?

PCDDs and PCDFs are not produced deliberately; they are unwanted byproducts that can, under special conditions, be formed during combustion and certain industrial processes. Dioxins are formed as a contaminant during the manufacture of certain chlorinated organic chemicals (e.g., pentachlorophenol [PCP] and certain herbicides). However, due to significant refinements in the manufacturing process, emissions from the chemical industry no longer represent a major source of dioxins nationwide. Today, in the U.S., older medical and municipal incinerators constitute one of the major remaining sources of dioxins still being released into the environment. Very small amounts of dioxins are also formed during the chlorine bleaching process and released into pulp and paper mill wastewater. Today, however, these emissions constitute a minor source of human exposure to dioxin, accounting on average for about 2% of daily intake (ATSDR 1998).

Estimates of average daily background exposures in the U.S. general population are 0.3 - 0.6 pg TCDD/kg/day and about 1 pg /kg/day of dioxin-like PCDDs and PCDFs (USEPA, 1994a; ATSDR 1998). (A picogram or pg is 0.000000001 or one trillionth of a gram.) Over 90% of that exposure comes from eating meat, fish, and dairy products contaminated with residues that first entered the food chain many years ago (ATSDR, 1998). Dioxins are very stable, highly lipophillic ("fat loving") compounds that, depending on the congener and the species, may also be relatively resistant to metabolism. As a result, dioxins have a strong tendency to bioaccumulate in fat, bind strongly to soils and sediments high in organic content, and persist in the environment for many years. Nevertheless, average concentrations of dioxins in biological and environmental samples have been declining since the 1970s and continue to do so (USEPA 2000a,c). Due to the existence of natural sources, however, dioxins will never disappear completely from the environment.

Before analytical techniques were sensitive enough to demonstrate otherwise, it was commonly thought that dioxins were produced exclusively as a man-made byproduct of industrial activities. It is now known, however, that dioxins pre-date not only the industrial revolution but the human race itself; they have recently been detected in 30 million-year-old clay deposits (Hayward et al. 1999; Winters, 2000). Dioxins are produced by natural, as well as anthropogenic, combustion processes, including forest fires. They are generated in very small amounts during the combustion of almost any organic material. The concentration of chlorine (which is always present in excess) is not a limiting factor (Rigo et al., 1996; Lemieux, 2000). Thus, contrary to popular belief, the burning of materials containing polyvinyl chloride (PVC) will not necessarily produce any more dioxin than does the burning of other organic materials. The rural burning of backyard trash, for example, actually produces more dioxin nationwide than does PVC manufacture (USEPA 1998;Winters, 2000).

2 Potential Adverse Health Effects of Dioxins

Some natural substances like botulinum toxin are more toxic than dioxin, but TCDD produces adverse health effects in laboratory animals at lower concentrations than any other man-made chemical. Less than one millionth of a gram per kilogram body weight (1 ug/kg) can slowly kill a guinea pig or give cancer to a rat. However, even these tiny doses are up to a million times higher than those that typically occur outside the laboratory, today. Average background doses in the U.S. and Europe are in the range of 1 trillionth of a gram per kilogram of body weight per day (0.000000000001 grams/kg/day or 1 pg/kg/day). During the Vietnam War, the herbicide formulation Agent Orange was contaminated with parts per million of dioxin (TCDD). Today, concentrations of dioxin in highly contaminated soil are measured in the low parts per billion (ppb). In food, dioxins occur in parts per trillion (ppt), in water, in parts per quadrillion (ppq), and in air, in parts per quintillion (ppqt). Each of these units of measure is 1,000 times smaller than the previous one. (See Appendix G for the definitions of standard units of concentration.) A millionth (10-6) of a gram is called a microgram (µg), a billionth (10-9) of a gram is a nanogram (ng), a trillionth (10-12) is a picogram (pg), a quadrillionth (10-15) is a femtogram , and a quintillionth (10-18) of a gram is called an attogram.. There are 28.35 grams in one ounce. (In subsequent sections, doses will generally be converted to pg/kg/day to facilitate comparison to an average human background exposure of 1 pg/kg/day.)

2.1 Animal Effects

Relatively little is known about the adverse health effects of non-TCDD dioxins, but the most toxic congener, 2,3,7,8-TCDD, is one of the most extensively studied of all known environmental toxins. Wherever sufficiently high doses of TCDD have been administered, a variety of effects have been observed in almost every animal species tested. Observed effects in animals include death, weight loss, liver toxicity, immune suppression, reproductive impairments, birth defects, and cancer (ATSDR 1998). The doses of dioxin required to produce these adverse health effects in animals vary enormously with species, as well as with strain, sex, tissue, and duration of exposure. For example, reported LD50 values for TCDD - an LD50 is the dose of a substance required to kill 50% of the exposed test animals - vary from 0.6 ug/kg (600,000 pg/kg) in male Hartley guinea pigs to 5,051 ug/kg (5,000,000,000 pg/kg) in Syrian hamsters. This represents more than an 8,000-fold difference between two species of rodent that are much more closely related to one another than either is to humans (ATSDR 1998).

Virtually all known chronic, intermediate, and acute effects levels for TCDD range upward from 100, 1,000 and 100,000 picograms per kilogram body weight per day (pg/kg/day), respectively. For non-TCDD dioxins, known effect levels in animals exceed a million pg/kg/day or 1 ug/kg/day (ATSDR 1998) In absolute terms, these effect levels are among the lowest recorded for any man-made toxic substance, which accounts for dioxin's reputation as "the most toxic man-made substance known." However, relative to the current potential for environmental exposure in humans, these levels are actually quite high.

ATSDR's chronic Minimum Risk Levels (MRL's) are estimates of daily doses that would not be associated with any detrimental non-cancer effects over a lifetime of exposure. Most are based on animal effects and the application of conservative safety factors. For example, ATSDR's chronic MRL for TCDD of 1 pg/kg/day, which approximates average background exposures in the U.S. several years ago, is based on less serious effects on social behavior in monkey offspring and a safety factor of 90 (ATSDR, 1998).

2.2 Human Effects

Humans appear to be 10 to 100 times less sensitive to the effects of dioxin than are laboratory rats and mice (Kimbrough, 1992; Aylward et al., 1998). It is generally assumed by regulatory agencies that sufficiently high dioxin exposures would produce in humans most, if not all, of the adverse health effects seen in laboratory animals. The validity of this assumption, however, has not been confirmed by the highest human exposures recorded to date, and, due to the elimination of most major sources of dioxin emissions, it is extremely unlikely that even higher exposures will ever occur in the future. No human fatality directly attributable to dioxin exposure has ever been recorded. More than 100 epidemiological studies have looked at the potential effects of dioxin and dioxin-contaminated herbicides in humans, including highly exposed workers in the chemical industry, soldiers exposed to Agent Orange in Vietnam, and persons accidentally exposed to TCDD at Seveso, Italy and at Times Beach, Missouri (Institute of Medicine 1994 ). A weight of evidence review of all of these studies indicates that chloracne is the only adverse health effect for which there is unequivocal evidence of a causal link with dioxin exposure in humans (Gotts, 1993, pg. 176; DeVito et al., 1995). Although not life-threatening, chloracne is a serious, potentially disfiguring skin eruption associated with unusually high exposures to dioxin, especially those that occurred prior to the 1980s as a result of occupational or environmental accidents.

The International Agency for Research on Cancer (IARC) recently reclassified TCDD from a "possible" to a "known" human carcinogen (IARC 1997). That classification, which ordinarily requires "sufficient" evidence in humans, was based in this case on "limited" evidence in humans, "sufficient" evidence in animals, and "supporting information" interpreted as suggesting a common mechanism of action for TCDD in various species, including humans. However, the actual mechanism by which TCDD induces adverse effects is still essentially unknown (USEPA, 1989; USEPA 2000a,b). The assumption of a common mechanism based on binding to a common cellular macromolecule (the Ah-receptor) is the first among several assumptions and inferences of uncertain scientific validity that serve as the logical foundation of the interim toxicity equivalency factor (TEF) approach (USEPA, 1989; USEPA 1994a ; USEPA 2000a,b). (See Section 3 of this appendix.)

EPA has recently announced its own intention to re-classify TCDD as "carcinogenic to humans" and other dioxin-like compounds as "likely" human carcinogens (USEPA 2000a). Like IARC, EPA bases its re-characterization of TCDD as "carcinogenic to humans" on extrapolations from animal data, hypotheses concerning the role of the Ah receptor in dioxin's mode of action, and purportedly equivalent "body burdens" (on a TEQ basis) in animal and human populations with cancer (USEPA 2000b, pg 87). At the same time, however, EPA acknowledges that the data from epidemiological studies of cancer in exposed humans "do not confirm a causal relationship between exposure to dioxin and increased cancer incidence" (EPA 2000b, pg 86).

Although some data are suggestive of an association between dioxins and cancer in humans, the studies do not support any firm conclusions. Even the best studies of the most heavily exposed occupational cohorts of the past have provided only inconsistent and inconclusive evidence that TCDD might cause cancer in humans (Zober et al., 1990; Manz et al., 1991; Fingerhut et al., 1991; Steenland et al., 1999). Generally, the observed excess risks for all cancers combined and specific types of cancer were relatively small - standard mortality ratios were generally less than 2.0 - and were seen only in the most highly exposed group with estimated 2,3,7,8-TCDD exposures 100-1000 times those seen in the general population. In addition, affected individuals tended to have exposure to other carcinogenic substances besides TCDD. In the absence of controls for such confounding exposures, the observed cancer excesses could not with any confidence be attributed to TCDD alone. In one of the largest and best-conducted of these occupational studies, observed increases in lung cancer and all cancers combined (42% and 15%, respectively) became statistically insignificant when attempts were made to control for smoking (Fingerhut et al., 1991).

The highest short-term TCDD exposures ever recorded occurred in 1976 at Seveso, Italy (Zone A), where mean exposures in children were estimated to be over 3 million pg/kg (Gough, 1994, pg 251). These children were, therefore, exposed in a single day to much more TCDD than an average U.S. citizen would be exposed to in a lifetime. The highest blood level of TCDD ever recorded (56,000 ppt) was measured in a female child at Seveso just days after the accident. Yet, with the sole exception of chloracne, follow-up studies of the Seveso cohort have firmly established no consistent or unusual pattern, either for the frequency or type of outcome, attributable to TCDD exposure (Bertazzi and Domenico 1994). Ten to fifteen years after the 1976 accident, there was an increase in some types of cancer, a deficit in others, and an overall decrease in expected cancer rates, especially among those with the highest exposure (Bertazzi et al., 1989, 1997). Thus, if dioxin is a human carcinogen, it must be a very weak one at doses that are realistically achievable outside the laboratory. In any case, the cancer risks that might be associated with dioxin exposures today would be undetectable and, hence, unverifiable.

A number of studies have reported measured differences in various developmental outcomes (e.g., levels of thyroxin, certain hormones, liver enzymes, and vitamin K; neurological endpoints; white blood cell counts and other immunological markers) in the breast-fed infants of mothers whose milk contained elevated levels of dioxin and dioxin-like compounds (Koppe, et al., 1991;Pluim, et al., 1993, 94a, 94b; Koopman-Esseboom et al., 1994; Huisman et al., 1995; Weisglas-Kuperus et al., 1995). However, the detected differences generally involve subtle and inconsistent biochemical changes that were well within the range of normal variation and are of no known clinical significance. A caveat common to virtually all such studies is that any hypothetical risks that may be associated with the consumption of dioxin- or PCB-contaminated breast milk are minor compared to the well-established medical benefits associated with breast-feeding (USEPA 2000b,c).

3 The TEF/TEQ Concept

The overlapping nature of some of the effects produced in animals by some congeners of PCDDs and PCDFs makes it highly desirable to include all such "dioxin-like" compounds in risk assessments for dioxin-contaminated sites. However, the relative paucity of relevant toxicological data on non-TCDD congeners originally made such a task virtually impossible. In the early 1980s, regulatory agencies like the EPA temporarily solved this dilemma by developing the TEF approach, and recommending its adoption as a "interim science policy position for use in assessing risks associated with CDD/CDF mixtures, until more definitive scientific data are available" (EPA, 1989).

3.1 Definitions

Toxicity Equivalency Factors (TEFs) are order of magnitude (factor of ten) estimates of relative toxicity of dioxin-like compounds, based on the available animal data for non-TCDD congeners and some in vitro results of limited relevance to whole animal toxicity. (The relative toxicity of 2,3,7,8-TCDD, the most toxic congener, is set equal to one.) These TEFs are often based on just one or two endpoints in a single species, usually Ah-receptor binding or enzyme induction in rodents (USEPA 1989). The Ah-receptor occurs in most vertebrate species, including humans. Binding to this receptor represents the first step in a process by which certain compounds, including PAHs and dioxins, can induce the synthesis of enzymes involved in their own metabolism. As normal physiological functions, neither Ah-binding nor the enzyme induction that may follow, are considered to be adverse health effects, per se. However, for the purposes of TEF development, they are interpreted as potential precursors of adverse effects. In the TEF approach, the measured concentrations of all 2,3,7,8-PCDDs and -PCDFs are converted to "equivalent" concentrations of TCDD by multiplying the concentration of each congener in the mixture by its TEF, thereby expressing each individual concentration in terms of hypothetical "toxicity equivalents" or TEQs. The individual TEQ values are then added together, to yield a single TEQ value for all of the PCDDs and PCDFs detected at the site (USEPA 1989; ATSDR 1998). This approach is expressed mathematically as follows:


Table 1.

Toxic Equivalency Factors (TEFs), 1989-1998
Dioxin Group TEF Furan Group TEF
          other TCDDs

          other TCDFs

          other PeCDDs

          other PeCDFs

          other HxCDDs

          2,3,7,8-HxCDF a
          other HxCDFs

          other HpCDDs



Octachlorodibenzo--dioxin (OCDD) 0.001
Octachlorodibenzofuran (OCDF) 0.001
a any isomer that contains chlorine in the 2,3,7,8-position.
b 1998 revised TEF

The TEFs used to calculate the TEQs were developed by the U.S. Environmental Protection Agency (EPA) and the international community in 1989 and revised in 1998 by the World Health Organization (WHO), parent organization of IARC (USEPA 1989; Van den Berg et al., 1998). Based on new studies of tumor promotion in rats and CYP1A1/A2 induction in mice, WHO raised the 1989 TEF for 1,2,3,7,8-PentaCDD from 0.5 to 1.0 and decreased the TEFs for OctaCDD and OctaCDF from 0.001 to 0.0001, respectively (Van den Berg et al., 1998). Alternative sets of TEFs do exist. Some states (California, for example) have developed their own versions, but all are strongly influenced by the International TEFs. The 1989/1998 International TEFs (I-TEFs) are listed in Table 1. (The 1998 revisions are in parentheses.)

3.2 Use of TEFs

The TEF method was originally developed as a means of arriving at approximations of the toxicity of mixtures of dioxin-like compounds that could then be incorporated into risk assessments. Risk assessments provide regulatory officials with a systematic basis for making risk management decisions and numerically ranking contaminated sites for cleanup. EPA originally considered the TEF method to be an interim approach only, one which will be replaced as soon as research yields a more accurate, practical alternative (USEPA 1989).

ATSDR uses total TEQs for screening purposes only. It is currently ATSDR's policy to use TCDD-specific comparison values, expressed in TEQs, to screen all dioxin-like compounds. However, once dioxin-like compounds have been selected as "chemicals of concern" at a site, further evaluation must follow, utilizing "the best medical and toxicologic information available" (ATSDR, 1992). TEQs alone lack the information needed for the determination of probable public health implications of exposures to dioxin-like compounds. If no congener-specific data are available, no further evaluation is possible because, when health assessors know neither the identities nor the concentrations of the site-specific contaminants of concern, they are unable to avail themselves of the chemical-specific information summarized in ATSDR's Toxicological Profiles for PCDDs, PCDFs, and PCBs (ATSDR, 1994, 1997, 1998).

Unfortunately, the term "toxicity equivalent" has been taken literally in many quarters and the TEF approach is now widely used by many people outside the regulatory establishment in ways for which it was never intended and is completely unsuited. EPA has indicated repeatedly that toxicity equivalency factors are not meant to be used precisely, even in a regulatory context (USEPA 1989, 1994a). It goes without saying that they cannot be used to accurately predict adverse health effects in humans.

3.3 Limitations of TEFs

The TEF approach was adopted by EPA as an interim science policy measure designed to facilitate risk management decisions (USEPA 1989). From the beginning, scientific and regulatory communities around the world cited the shortcomings in the science base supporting the TEF concept, and the latter has always been subject to revision as new experimental data became available (USEPA 1989). Some of the inconsistencies in the TEF method are summarized below.

TEFs are not precise measures of relative toxicity, and they were not designed for the prediction of adverse health effects (USEPA 1989, 1994a). EPA considers the TEF/TEQ approach to be "a useful, but uncertain, procedure." Although EPA has expressed increased confidence in TEFs and has expanded its use of the TEF/TEQ approach over the last 10 years, that agency still identifies the latter as and "interim methodology"for which "the need for research to explore alternative approaches is widely accepted" (USEPA 2000b, Chapter 9, pg 9-31). Today, TEQs based on those TEFs, though useful as screening devices, still cannot be used in the Public Health Assessment process as surrogates for actual, congener-specific data.

TEFs are order of magnitude estimates of average, relative toxicity based on highly variable data sets. TEFs are derived, by consensus, from limited experimental results that typically range over 1-to-2 orders of magnitude. These data most often come from in vitro studies and/or short-term animal studies of Ah receptor binding or P450 enzyme induction which do not represent true measures of whole-body toxicity (USEPA 1989, 1994a).

Although TEFs are defined and used as constants, the numerical value of an experimentally determined TEF will, in fact, vary significantly with a number of parameters, including: dose, duration of exposure, species, sex, strain, target tissue, and biological endpoint (DeVito and Birnbaum, 1995; Safe, 1990; USEPA 1994b). In some cases, e.g., the levels of glucose and certain blood cell types, TCDD may even have opposite effects in different species (USEPA 1994b).

TEFs can be expected to become increasingly imprecise as the conditions under which they are applied become further removed from the conditions under which they were derived. (Putzrath, 1996). TEFs are generally derived from relatively high-dose data obtained in vitro or from short-term rodent assays, but they are most often applied in the context of low-level environmental exposures in humans.

TEFs do not take into account many important biological factors. For example, the health implications of exposure to a mixture of dioxin-like compounds will depend on, among other things, the bioavailability of specific congeners and the potential for antagonistic interactions between them. TEFs, however, do not account for either (USEPA 1989, 1994a).

Finally, the assumption that all adverse effects of dioxin and dioxin-like compounds share a common mechanism, involving, as its first step, binding to the Ah receptor, is just that -- an assumption. Whether or not it is actually true, this assumption is critical to the TEF approach because it provides the theoretical basis for the simplifying inference that TEQs are additive. However, except for the chain of events leading to the induction of certain enzymes (e.g., cytochrome P-450IA1), which is not per se an adverse effect, clear evidence for such an assumption is still lacking (USEPA 1989, 1994c, 2000b, Chap 2, pg 2-19). The actual mechanism of action is still unknown for virtually all dioxin-induced adverse effects, and the available data indicate that binding to the Ah receptor is not sufficient in itself for the mediation of toxicity (USEPA 1994c; USEPA 2000b, Chap2). For example, the species most sensitive to TCDD-induced wasting syndrome (i.e., guinea pigs) is the least sensitive to CYP1A1 induction, a common measure of Ah-binding and the basis of many TEFs.


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Zober et al. (1990). Thirty-four-year mortality follow-up of BASF employees exposed to 2,3,7,8-TCDD after the 1953 accident. Int Arch Occup Environ Health 62: 139-157.


The Agency for Toxic Substances and Disease Registry (ATSDR) issued a draft for public comment petitioned public health assessment on September 30, 1999 for the Koppers Company, Inc. (Oroville Plant) site. Between September 30, 1999, and November 29, 1999, the public had the opportunity to provide comments on the draft public health assessment. During that time, ATSDR received written comments and questions from the California Department of Health Services (CDHS), the California Department of Food and Agriculture (CDFA), Koppers Industries, and the U.S. Environmental Protection Agency (EPA). These comments/questions are presented below. Each comment is followed by a response from ATSDR.

California Department of Health Services

Comment: Our Branch, the Environmental Health Investigations Branch of the California Department of Health Services (CDHS) appreciates that ATSDR has completed the public comment document. However, we are concerned that ATSDR has mischaracterized data collected by CDHS, has not reviewed the full body of available scientific literature in assessing whether exposure levels are likely to cause adverse health effects, and a complete review of available data has not been summarized. Below we provide detailed comments of our concerns.

Response: ATSDR thanks CDHS for their comments. ATSDR has responded to the specific comments presented below and has made changes to the public health assessment as it deemed necessary in response to CDHS's concerns.

Comment: Page 1: "may have resulted in reversible adverse health effects;" Page 8 "ATSDR understands that the reported symptoms have since subsided and no evidence exists that any chronic effects have occurred;" Page 13 "ATSDR does not have sufficient data...but believes that if any outcomes were triggered by exposure to domestic water supplies, the outcomes were quickly reversible. ATSDR understands that the reported symptoms have since subsided and no evidence exists that any chronic (noncancer) effects have occurred as a result of past exposures;" and Page 26 "However, ATSDR understands that the reported symptoms have since subsided and no evidence exists that any chronic (long-term) effects have occurred." These statements require a description of the date on which these statements are based and discussion of the limitations of any such evidence. No data is presented in the document to support ATSDR's "belief" and "understanding." It is unclear what "evidence" was reviewed to determine that "no evidence exists." With the exception of a review of cancer statistics, which as described on page 25 is extremely limited, ATSDR has not conducted follow-up studies or, to CDHS knowledge, conducted studies of chronic effects among the residents or conducted follow up interviews with residents. If "no evidence" refers solely to a literature review and a review of exposure data, this requires clarification and greater detail and also should not be contained in the same sentence as "reported symptoms." These sentences also summarize CDHS studies of symptoms. CDHS opted to call the symptoms "acute." We consider "acute" more appropriate terminology than "reversible." "Reversible" and "subsided," terms which ATSDR uses throughout the document, implies that CDHS actually studied and observed the reported symptoms to cease. CDHS did not do this.

Response: ATSDR understands that no formal followup studies have been conducted to evaluate the current health status in the community. ATSDR's reference to the apparent absence of observed longer-term health effects is based on its understanding that few, if any, complaints have been voiced by area residents in recent years. ATSDR acquired this information via communications with the EPA and CDHS throughout the public health assessment process. ATSDR revised the text to help clarify this point. It now reads "To the best of ATSDR's knowledge, no recent health complaints have been received from people living/working in the vicinity of the Koppers site. No specific studies or surveys, however, have been conducted to collect health status information from area residents." (See page 14.)

In choosing the term "reversible" instead of acute or short-term, ATSDR's intent was to help the public better understand that any symptoms workers/residents may have experienced in the past should not have led to long-lasting health effects. Because of limited exposure data, ATSDR could not draw any firm conclusions regarding the association between the type of effects reported (e.g., skin irritations, headaches, etc.) and possible site-related exposures. ATSDR wanted to emphasize, however, that these types of acute responses are not expected to lead to longer term (chronic) conditions-that is, the outcomes went away, or are "reversible," when the exposure ended. In this case, the term "reversible," alone, is more informative than is the term "acute", alone (CDHS's preference). The term "acute" really only tells the reader how quickly the effect manifested itself after exposure; it says nothing about how long the effect might have lasted. Similarly, the term "reversible" suggests that the effect did not persist for long, but says nothing about how long it took for the effect to appear in the first place. Neither term is exclusive of the other, however, and ATSDR agrees that both terms should probably be used. Therefore, ATSDR has modified certain sections of the PHA (using both terms as necessary) to provide greater clarity.

Comment: Page 11: "Monitoring well data indicate that groundwater conditions have continued to improve since treatment began in 1993. Contaminant concentrations and the size of the plume continue to decrease." Some description of the supporting data is required here. The above two paragraphs cite a 1988 document.

Response: Throughout the health assessment process, ATSDR consulted with CDHS and EPA to ensure that we had a full understanding of the nature and extent of groundwater contamination beneath and downgradient of the site (on and off property), both presently and in the past. To evaluate groundwater conditions subsequent to 1993, ATSDR relied on data provided by EPA which included PCP data for domestic wells and monitoring wells on and off site, compiled through May/June1998. Figure 3 in Appendix A (newly added) delineates the areas where pentachlorophenol (PCP) concentrations still exceed the cleanup goal of 2.2 parts per billion.

ATSDR has added the following text to further describe post-1993 conditions:

"As can be seen in Figure 3, the "plume"of PCP contamination in groundwater has receded. Residual PCP contamination still exists to the south of the Koppers property boundary, but not near any domestic wells currently in use (Dames & Moore 1996; EPA1997a, 1997b,1998, 1999a, 1999b, 2000; HSIGeoTrans 1999)."

Comment: Page 12: "The other eight wells will be continually tested." Page 26: "Groundwater not meeting health-based cleanup goals should not be used for potable purposes." In both these instances, follow-up activities with residents should be recommended and an agency identified who will carry them out to ensure that residents are aware that water should not be used for drinking purposes.

Response: As noted in the Background section of the public health assessment (page 5), Beazer Materials and Services, Inc., (Beazer) is conducting cleanup actions under the terms of an agreement (known as a consent decree) with EPA and the conditions set forth in the "Record of Decision" (ROD) for this site. These actions include the treatment and monitoring of groundwater, both on and off property. EPA is overseeing these actions, which include the continued sampling of monitoring and domestic wells where low levels of PCP contamination are still being detected. As of December 1995, EPA determined that the area of PCP-contaminated groundwater had receded to the extent that the operation of the off-property pump and treat system could be discontinued. The residual off-property PCP contamination is being treated by use of enhanced in-situ bioremediation.

A monitoring plan is in place for on-property and off-property groundwater which generally requires quarterly sampling of area groundwater. Monitoring requirements vary depending on the concentrations of constituents detected in individual wells. If contaminants are detected at levels exceeding the ROD goals in a given well, then quarterly sampling is required. If contamination is not detected in a given well for four consecutive sampling periods (quarters), then sampling of that well will be required every 6 months. After 4 years of non-detects, a 2-year sampling of that well will be required. If at any time contamination is detected, quarterly sampling will be resumed in that well.

ATSDR has revised the text on page 12 to be more explicit about the monitoring program that is in place and overseen by EPA.

Comment: Page 13: "A review of available health effects literature revealed that virtually all studies looking at toxic effects of PCP are animal studies." There are human studies of the toxicity of pentachlorophenol (PCP) that are notable. CDHS reviewed the occupational studies and child case reports that are in the literature for the cited June 1997 CDHS Biological Monitoring and Health Interview Study (Table 16). Some of these literature citations document human death due to PCP exposure. In addition, there is a recent study of PCP workers (Hryhorczuk DO, et al. Environmental Health Perspectives106:401-408, 1998) which presents new findings and summarizes the available human studies on PCP.

Response: ATSDR agrees that this statement requires clarification. This statement was made in the context of exposures to PCP in drinking water, where ingestion is the primary exposure route of concern. Multiple human studies pertaining to PCP exposures certainly exist, but most relate to worker exposures to technical-grade PCP via inhalation or skin contact, as well as studies looking at residential exposures to PCP-treated wood. Many of these studies have no direct relevance, however, to drinking water exposures in the vicinity of the Koppers site. Few reports exist in the literature concerning adverse effects in humans ingesting PCP, the primary route of exposure.

In assessing potential public health impacts of PCP-contaminated groundwater at this site, ATSDR health assessors reviewed the literature on PCP, including the human data to which CDHS refers to in its comment. ATSDR reviewed its Toxicological Profile for Pentachlorophenol (ATSDR 1999) and also conducted a comprehensive literature search. In doing so, ATSDR sought to identify studies (human and animal) that could provide public health perspective given the information at hand regarding possible residential exposures to elevated levels of PCP in the drinking water before alternate supplies were made available.

As noted above, identified studies look primarily at worker exposures via inhalation or contact with "high" levels of technical-grade PCP, as well as studies evaluating indoor exposures in homes using PCP-treated wood. Certain factors, however, limit the usefulness of these studies: exposure and/or dose data are limited in many cases (that is, investigators do not know the PCP levels to which workers/residents were exposed or the doses responsible for the observed effects), and information on exposure to other chemicals or other risk factors are lacking (for example, it is not known whether individuals with observed health effect[s] were exposed to other chemicals or are susceptible to the observed effect[s] for other reasons [e.g., genetic susceptibility, smoking, etc.]). For these reasons, the data from these studies provided limited information on which to assess residential exposure to relatively low levels of PCP in their drinking water.

ATSDR, therefore, to a large extent, relied on data from animal studies in which the doses associated with observed adverse health effects are documented and can be used for comparison purposes in the evaluation of drinking water exposures at the Koppers site. As noted in the public health assessment, the estimated doses in area residents associated with measured levels of PCP are lower than those shown to result in adverse health effects in available studies. For added perspective, and as is described in the text (page 23), ATSDR also reviewed the limited data available on the PCP levels measured in the urine of a subgroup of area residents. These data showed that PCP levels the urine of study participants were lower than those known to be associated with health problems.

The discussion on page 13 has been modified to read as follows:

"While numerous studies exist that have identified adverse health effects associated with PCP exposures in the work setting or in homes with PCP-treated wood, little human data exist describing the effects of ingesting PCP in drinking water. Therefore, in evaluating potential health effects associated with drinking water exposures, ATSDR relied on data from experimental animal studies to provide additional perspective. Available animal studies show effects at doses above 1 milligram per kilogram of body weight per day..."

Comment: Page 14: "No firm evidence linking PCP exposure in water (drinking or skin contact) and the types of symptoms reported. Most available studies looked at exposure to PCP solutions versus diluted levels in potable water." These statements read as if ATSDR is making a distinction between drinking water and PCP solution other than the obvious difference in concentrations. To CDHS, such a distinction is not a health protective approach. There is also a case study of a hospitalized child who was exposed solely to PCP contaminated bathwater (Chapman, JB and Robson P. Lancet 1:1266-1267, 1965). Could ATSDR explain why they attribute importance to the distinction?

Response: The distinction that ATSDR was making in this statement is that it is difficult to assess the potential harm caused by diluted concentrations of PCP in water based on studies that look at observed effects associated with exposures to more highly concentrated solutions. The literature presents documentation of various adverse effects associated with these high level exposures. In some cases, effects are cited, but the concentration of the PCP solution is not provided. While these types of reports certainly tell us that harmful effects can result from dermal exposures to PCP, they do not offer the dose perspective that is key to understanding and communicating the significance of possible exposures in the vicinity of the Koppers site.

The text on page 14 has been modified to read: Studies are not available that evaluate the health effects associated with the PCP levels detected in the vicinity of the Koppers site.

The article cited by CDHS (Chapman et al. 1965) itself illustrates why ATSDR "attributes importance to the distinction" between PCP solutions and trace contamination of groundwater. The situation addressed by that case study is briefly described as follows. Thirteen days before a family moved in to it, a house and its roof timbers were sprayed with an oil-based insecticide formulation containing pentachlorophenol, beta-naphthol, and dieldrin. In the process, the cold water storage tank was contaminated, so much so that plumbers (who subsequently discovered the contamination) were called in to investigate an oily scum on the bath water. After being bathed in this water almost daily for 13 days, a little girl approximately 45 months of age was admitted to the hospital with fever, acidosis, and ketonuria. Urinary PCP levels were 6 mg/100 ml or 60,000 ppb. Nevertheless, the child recovered quickly and completely.

From the toxicological maxim, "the dose is the poison," it follows that some exposure scenarios will be more relevant than others to the site-specific conditions of exposure which are the focus of every ATSDR health assessment. The unusually high exposures described by Chapman et al (1965), for example, are not comparable to those that would result from the contact with or ingestion of groundwater containing trace amounts of PCP at the Koppers site.

Comment: Page 13-14, 1st paragraph: "inadequate evidence exists to show that PCP causes cancer in humans" and "estimated doses would be...which are 3000 times lower than the levels resulting in cancer in laboratory animals." The caveats and uncertainties of the data presented require presentation. A public health protective approach assumes that agents that cause cancer in animals may cause cancer in humans. Chemicals for which there are studies in both humans and animals demonstrate the validity of this approach. Animal studies are also limited in number and the true dose resulting in cancer in animals, and humans, is unknown. CDHS is supportive of the statement "area residents were unlikely to be exposed to site-related contaminants at high enough concentrations or for enough time to suffer long-term health effects (page 1)." However, the uncertainties in this statement are far greater and margin of safety far less than the paragraph on page 14 indicates.

Response: ATSDR's primary goal in its public health assessments is to put possible exposures to environmental contaminants into meaningful perspective for the public. In doing so, ATSDR strives to explain the likelihood that exposure to the detected level of a particular chemical may cause harm. ATSDR recognizes and completely agrees that a prudent public health approach is to assume that agents that cause cancer in animals may cause cancer in humans. It is also important, however, to present the weight of evidence that a given agent (in this case PCP) is associated with cancer in humans and, based on this, assess the likelihood that cancer could result from exposure to detected levels of PCP at the site. In doing so, ATSDR considers all relevant scientific data. We look at 1) the evidence of carcinogenicity in animals and humans, 2) how similarly or differently the chemical acts in humans versus animals (i.e., pharmacokinetics/dynamics, mechanism of action), 3) susceptibility and sensitivity in certain persons (e.g., children), 4) cancer registry data, and 5) any uncertainties in the available data sets.

For PCP exposures via drinking water in the vicinity of the Koppers site, ATSDR concluded that residents were unlikely to be exposed to PCP at high enough concentrations or for enough time to result in increased cancer risk. The rationale for ATSDR's conclusions is as follows: First, ATSDR recognized that PCP is classified as a "probable" human carcinogen. PCP has been shown to cause cancer in laboratory animals via ingestion, but at doses significantly higher than those expected in people drinking water at the measured levels of PCP. Furthermore, it is uncertain whether PCP even causes cancer in humans. Case reports for workers exposed to PCP that show possible associations are largely inconclusive and several epidemiologic studies have shown no association (ATSDR 1999). Very little human data exist for ingestion exposures. In presenting this information, ATSDR's goal is to help the public better understand the uncertainties and, again, enable them to put possible exposures into perspective.

ATSDR has revised the discussion on page 14 to better illustrate these points:

"ATSDR also looked at potential cancer threats posed by PCP at measured levels. ATSDR does not believe that PCP levels were high enough or exposure was long enough to cause cancer. PCP has been classified as a "probable human carcinogen." Although inadequate data exist to show a link between PCP and cancer in people, PCP has been shown to cause cancer in laboratory animals under experimental conditions where high doses were administered. Because possible effects resulting from lower level exposures are difficult to study, scientists are uncertain about the effects of such exposures, especially in humans. That is why screening values (e.g., ATSDR's Cancer Risk Evaluation Guide) are set very low to help ensure that people are not exposed to contaminants levels even remotely close to effects levels seen in experimental studies. It is important to note that even if a 150 pound (70 kg) person were exposed to the average PCP concentration of 200 ppb detected in area groundwater (drinking 2 liters per day, every day over the course of a lifetime), estimated doses would be 0.006 mg/kg/day, which are 3,000 times lower than the lowest levels resulting in cancer in laboratory studies (18 mg/kg/day) (NTP 1989). Lastly, cancer statistics compiled for the census tracts in the vicinity of Koppers through 1989 have not shown increased cancer rates (see Health Outcome section below for further discussion on available cancer statistics, including their limitations).

As far as the "uncertainties" being "far greater" than ATSDR indicates, EPA's own assessment describes the "uncertainties" associated with cancer risk estimation as follows: "the true risks are unknown and may be as low as zero." It is easy to confuse the objective of regulatory "risk" assessment with that of a "public health" information service. EPA's cancer classification of PCP is based solely on "sufficient" evidence from animal studies, supported by inconsistent evidence that PCP may be weakly genotoxic. EPA characterizes the human evidence of carcinogenicity as "inadequate." Given the dose levels and durations of exposure at the Koppers site, PCP-induced cancer is not a realistic health concern.

Comment: Page 14: "Cancer statistics compiled for census tracts...have not shown increased cancer rates." As noted above, making this statement without noting the limitations of such an analysis is misleading.

Response: ATSDR agrees that discussing the limitations of the available cancer data is critical. The limitations are discussed in the discussion of Cancer Registry Data on page 26 and ATSDR has added a reference to these limitations on page 14. See also the response to the previous comment.

Unsupportable speculation has its place, but not in a public health assessment. The latter is designed to offer relevant answers to specific questions, using the best medical and toxicological information available (ATSDR 1992c).

Comment: Page 18: "In looking at the data collected in 1987 and 1988, ATSDR found that approximately 98% of the TEQ in eggs consists of the hepta-, hexa-, and penta-forms of dioxins/furans. All known effects, ..." CDHS considers ATSDR's approach of examining the toxicity of the individual congeners unnecessary. The described TEQ approach is an internationally accepted approach for equating the other forms of dioxins/furans to TCDD and assumes that the other forms have much lower toxicity.

Response: ATSDR recognizes that the TEQ approach is an acceptable and convenient screening approach for assessing potential dioxin/furan hazards. However, as stated on page 18 and in Appendix E (page E-5), because of uncertainties regarding differences in congener toxicity and the mechanisms of toxicity in humans, TEQs cannot be used as a tool for predicting adverse health effects in humans. When screening values are exceeded, further evaluation is needed. Such an evaluation must look at all relevant toxicologic and medical information. In this case, that means 1) congener-specific exposure data, 2) congener-specific effects data, and 3) established principles of toxicology and medicine. That is why, in its assessment of dioxins in home-raised eggs, ATSDR also reviewed the congener-specific data. The purpose was to describe the uncertainties to the public and to provide additional perspective regarding possible dioxin exposures.

Perhaps some quotes from an EPA assessment of TEFs may help clarify the scientific validity of equating forms of dioxin/furan to TCDD and predicting the health effects of those mixtures in humans. The following quotes are all from EPA (1989), "Interim procedures for estimating the risks associated with exposures to mixtures of chlorinated dibenzo-p-dioxins and -dibenzofurans (CDDs and CDFs) and 1989 update," EPA/625/3-89/016.

"...many of the short-term results seen in murine systems are not observed in rat systems. Also, the connection between the enzyme induction response, which supports several of the TEF values, and several of the toxic endpoints manifested by CDDs/CDFs, is unclear. Other mechanisms of action, e.g., effects on vitamin A synthesis and estrogen-like activity, have been suggested as playing an important role in the toxicity of CDDs/CDFs. These continuing elements of uncertainty in the TEF approach highlight the need to treat the approach as ‘‘interim,’‘ that is, one that needs to be further buttressed by experimental data and eventually replaced with a more direct biological assay." (Page 8)

"...these estimates [i.e., TEFs for Hp- and Octa- CDDs/Fs] ignore the issue of relative bioavailability of the CDD/CDF congeners." (Page 10)

"With the exception of 2,3,7,8-TCDD, the 2,3,7,8-HxCDDs, and 2,3,7,8-TCDF, the TEFs are not based on the results of major animal (reproductive, carcinogenic) studies. Generally, TEFs are based on estimates of the relative toxicity in in vitro tests whose relationship to the chronic effects of concern is largely presumptive." (Page 13)

"...research holds the promise of removing the need for any TEF scheme. This is particularly important in the light of the emerging data showing that some of the CDDs/CDFs and related compounds can exhibit antagonistic effects (Safe, 1987), a possibility that is ignored in current TEF approaches."

Although the use of the TEF approach has expanded considerably since 1989, the scientific limitations of the TEF method remain essentially the same more than 10 years later. The concept of "toxicity equivalence" was not developed for the realistic prediction of human health risks from chemical exposures and should not be interpreted literally; it was identified by EPA in its 1994 risk assessment of dioxin as one of its seven "key assumptions and inferences" (HAD 1994, Vol III, pg. 9-73).

ATSDR believes that an examination of individual congeners is of critical importance. Without it, "further evaluation" (which, according to ATSDR guidance, must follow whenever maximum, site-specific concentrations of a contaminant exceed the relevant comparison values) would be impossible. Only by taking into account all of the relevant chemical, biological and pharmacologic factors that are not accounted for in the TEF approach is it possible to produce a health assessment that goes beyond regulatory assumptions in an effort to ascertain (as nearly as possible) the realistic exposures and health risks in people exposed under site-specific conditions.

Comment: Page 18: "Humans are thought to be 10 times less dioxin-sensitive than most laboratory animals," and Page 24: "Furthermore, humans are thought to be at least 10 to 100 times less sensitive to dioxins than are rodents." These statements are in disagreement with a recent review article which states "humans and animals have similar degrees of sensitivity." (Grassman JA, et al. Environmental Health Perspectives: 106(2):761-775, 1998).


The main reason for the appearance of a disagreement is that the first statement that "humans are thought to be 10 times less dioxin-sensitive" was a quantitative conclusion based on toxicological principles and research while the comment attributed to Grassman et al. was a qualitative statement based largely on regulatory assumptions and methodologies. Grassman et al. do assert that humans respond "similarly" and have "similar" degrees of sensitivity to dioxin-induced effects. However, those assertions were not actually supported in the body of that article. In the review article by Grassman et al., the authors briefly describe a variety of real and presumed effects (adverse and otherwise) of dioxin exposure in animals and humans, but make no real attempt to either establish causality in humans or directly compare the doses that produce the same effect in humans and animals. Nor do the authors ever clarify their use of the term "similar." Such clarification is available, however, from the reference (# 17) cited by Grassman et al. as the source of their statement regarding the alleged similarities between human and animal responses to dioxin and dioxin-like compounds. In that cited article, DeVito, Birnbaum, Farland, et al. (1995) stated that, as they used the term, the word "similar" meant "within a factor of ten."

Unlike Grassman et al.(1998), DeVito, Birnbaum, Farland, et al. (1995) did compare effective doses in an effort to support their conclusions. However, those efforts depended heavily on a number of assumptions and practices that are not generally employed by toxicologists outside of the regulatory community. These include:: (1) the definition of "similar" as "within a factor of ten," (2) a focus on non-adverse effects, (3) the use of "body burden" (i.e., the concentration in fat) as the toxicologically-relevant measure of dioxin dose, (4) a literal interpretation of TEQs as toxicity "equivalents," and (5) certain assumptions about the essentially unknown mechanism of action of dioxin, assumptions which are crucial to the TEF approach to risk assessment but remain largely unproven. Thus, the conclusions of DeVito et al. were speculative in nature. This did not constitute a problem with the DeVito et al. paper, because the authors themselves acknowledged many of the limitations of the assumptions on which their conclusions were based. More importantly, those conclusions were not presented as scientific findings that clearly established either the existence of or the mechanism behind dioxin-related adverse health effects in humans. Those authors’‘ primary purpose was to present plausible, theoretical support for the validity of the TEF concept, for regulatory purposes. This they accomplished. In so doing, however, they openly acknowledged that "chloracne is the only toxic effect induced by dioxins for which there is unequivocal evidence linking exposure to effect in humans" (DeVito, Birnbaum, Farland, et al. 1995, middle of pg 1). (Additional clarification of this and other dioxin-related issues can be found in an Appendix E.)

Comment: Page 18, next to last paragraph, and Page 24, last two paragraphs: These paragraphs discuss the toxicity of dioxins and furans and the concentrations found in eggs. These paragraphs should include the recent human cancer studies and the International Agency for Research on Cancer evaluation of TCDD as a human carcinogen. The cited CDHS reference (Goldman LR, et al., Environmental Health Perspectives: 108:1-7, 2000) compares levels observed in one of these studies to those found among the egg and beef consumers. These cancer studies estimated blood levels among a group with a documented elevated rate of cancer to be approximately equal to those found among the egg and beef consumers. Also pertinent is that researchers have estimated that "some individuals may respond to dioxin exposures with cancer and noncancer effects at body burdens within one to two orders of magnitude of those in the general population" Devito MJ et al., Environmental Health Perspectives: 103(9):820-831, 1995).

Response: Appendix E containing general information on dioxin and dioxin-like compounds has been inserted into the final version of ATSDR's Koppers Public Health Assessment. It addresses, among other things, the issues of 1) human cancer studies, 2) IARC's re-classification of 2,3,7,8-TCDD as a human carcinogen, and 3) the limitations of using "body burden" as a dose metric to establish toxicological relationships. In Appendix E, ATSDR endeavors to make clear the distinction between established, objectively-demonstrable, toxicologic and human health observations, on the one hand, and prudent, politically acceptable, public health policy on the other. Nowhere is this distinction more important than in discussions of dose-response relationships. The reader is referred to Appendix E for details. The rest of the following response specifically addresses each of the 3 separate parts of the above comment concerning 1) IARC's re-classification of TCDD, 2) The CDHS reference, Goldman et al. (2000), and 3) the body burden statement from DeVito (1995).

The epidemiological evidence to date remains inadequate to establish a cause-and-effect relationship between TCDD exposure and cancer in humans. Both IARC and EPA continue to regard the epidemiological evidence for TCDD-induced cancer in humans as "limited." Thus, IARC's re-classification (like EPA's proposed re-classification) is based on "limited" evidence in humans, and "sufficient" evidence in animals. Historically, these two criteria have been used to justify designating substances as "probable" human carcinogens. In the case of TCDD, the classification as a "known" human carcinogen depended on additional considerations, which IARC referred to as "supporting evidence." These considerations included 1) evidence interpreted as implying the existence of a common, Ah receptor-dependent mechanism of carcinogenic action in both humans and animals, and 2) the observation of "similar" tissue concentrations of TCDD in animals and human populations exhibiting elevated rates of cancer. (The actual blood levels compared, in this case, ranged from 301 to 32,000 in 4 epidemiological studies and from 1,500 to 10,000 ppt in rats (IARC, 1997, pg342-3).) The inferences drawn from both lines of this "supporting evidence" depended heavily on assumptions central to the TEF approach to the assessment of risk from exposure to dioxin-like compounds. These assumptions and inferences, while useful for the regulatory purposes of screening and risk assessment, lack both the precision and the necessary science base to make them valid tools for predicting effects across species. (See response to previous question.)

For example, the regulatory assumption that most, if not all, of dioxin's adverse effects are mediated by binding to the Ah receptor, the actual mechanism of virtually all of those adverse effects remains essentially unknown. Also, while concentrations in body fat or blood lipids represent excellent biomarkers of exposure to TCDD, they do not constitute reliable biomarkers of effect. (Grassman et al., 1998, which was cited in the previous comment, specifically warns that "cross-species comparisons of tissue concentrations should be approached with caution.") Notwithstanding the regulatory assumption that dioxin concentrations in adipose tissue (used as a surrogate for "body burden") are directly relevant to adverse health effects, there is no evidence that those effects are actually caused by the dioxin that has been sequestered in fat. On the other hand, there is substantial evidence that sequestration in fat actually protects the organism from dioxin-related adverse effects by drawing dioxin away from target tissues like the liver (Kimbrough 1992). There is even a good linear correlation between the LD50 of dioxin in many different species and the log percentage total body fat in those species (Geyer et al. 1990, Toxicology 65:97-107). Based on this relationship, the estimated 30-day LD50 in humans would be about 6,000,000,000 pg/kg. Children exposed during the 1976 accident at Seveso, Italy (Zone A) had the highest blood levels of TCDD ever recorded (up to 56,000 ppt compared to a U.S. average at the time of perhaps 7 ppt) and, at Zone A, were exposed to an estimated mean of 3,125,000 pg TCDD/kg. Nevertheless, after more than 15 years of medical follow-up, chloracne (which resolved in most cases within a few years) remains the only adverse health effect suffered by these individuals that can be attributed to their TCDD exposure. In the context of such information, speculations based on body burden comparisons between animals and humans (or between humans with and without elevated cancer/mortality rates) would have no impact on ATSDR's assessment of the potential for dioxin-related health effects in individuals exposed at the Koppers site.

The limitations of the "body burden" approach to dioxin risk assessment also apply to the quoted comment from DeVito et al.(1995). The statement to the effect that TEQ body burdens within one to two orders of magnitude of those in the general population may be causing cancer and noncancer effects in some people is essentially speculative. Such speculations are commonly necessary to support risk assessments and to justify regulatory decisions. However, they are not particularly relevant to the assessment of the public health effects that are likely to occur in people potentially exposed to site-related contaminants under site-specific conditions.

Finally, it should be noted that Goldman et al. did not actually claim that estimated blood levels among a group with a documented elevated rate of cancer were "approximately equal" to those found among the egg and beef consumers. They observed, rather, that the difference between the means of the high exposure and comparison groups (48 pg ITEQ/g) in their own study was "approximately equal" to the difference between background and the lowest exposed group (40 pg ITEQ/g) with significantly elevated cancer rates in the Hamburg occupational mortality study by (Dieter) Flesch-Janys et al. (1995). (Note that the first author's last name was left off in the cited reference in Goldman et al, 2000.) In the latter case, quantitative estimates of PCDD/F exposure in the whole cohort and specific exposure groups were based on extrapolation and back-calculation from levels measured in blood lipids and adipose tissue in a subgroup of workers (N=190) several years after exposure, and the referents were gas workers. In the Goldman et al. study, the mean exposure in the high exposure group (i.e., egg and beef consumers) was derived from blood samples from only 5 individuals, and PCDD/F levels in the rural comparison group were generally lower than levels in the urban U.S. population. Thus, neither the exposure nor the comparison groups used in these two very different studies are comparable to one another. Also, regarding the implication of significantly elevated cancer rates in the lowest exposure group in the Hamburg occupational mortality study, it should be noted that the relative risks for all cancers combined - the authors did not look at specific types of cancer - remained non-significant or borderline significant until the fifth quintile (highest exposure group) which included estimated exposures ranging from 545 to 4,362 ppt TEQ (Flesch-Janys et al., 1995). In any case, for all of the reasons mentioned previously, group average concentrations of dioxins in fat are no more predictive of adverse health effects than are individual concentrations in fat.

As ATSDR has stated elsewhere in this Public Health Assessment, exposures to products potentially contaminated with elevated levels of dioxins should be limited wherever reasonably possible. (See pages 2, 19, 26, and 27). However, that recommendation is predicated on prudent public health policy and not on any scientific data or speculative estimates suggesting that adverse health effects are likely to occur now or in the future at the levels of exposure documented at Koppers/Oroville. Such distinctions are a critical component of effective risk communication and, hence, of ATSDR's Public Health Assessments.

Comment: Page 24: "Blood levels in the highest consumers group (beef and egg) were less than 9% higher than the high end of this range." This comparison is unnecessary and should be deleted. The cited CDHS article (Goldman LR, et al., Environmental Health Perspectives: 108:1-7, 2000) compares the beef and egg consumers, not only to appropriate age and gender matched rural comparison subjects but also to an urban population, with an approximately equal age and gender distribution. It is unclear whether the ATSDR cited reference groups are age and gender appropriate. Further, means should be compared; not the "high end" as the ATSDR cited reference groups may be much larger and more likely to have outliers than the smaller egg and beef study group.

Response: The primary function of a public health assessment is to effectively communicate to potentially exposed residents a meaningful and useful perspective on the probable health implications of their site-specific exposures. The point of ATSDR's comparison was that the highest blood levels detected at Koppers were not very high, relative to the range of background exposures in the United States and were still quite low compared to levels known to produce effects in human beings. And, a congener-specific analysis demonstrated that the more highly-chlorinated (and, therefore, less toxic) congeners predominated. ATSDR concurs with CDHS's cautionary recommendation that exposures be limited wherever possible. However, for effective risk communication, it must also be acknowledged that the documented levels of exposure in Oroville residents would not be expected to result in adverse effects and, therefore, need not occasion alarm. Any such alarm would be much more likely to lead to detectable adverse health effects (e.g., stress-related, psychogenic illness) than would any of the site-specific dioxin exposures. ATSDR considers that the cited statement provided the appropriate perspective.

Comment: Page 24: "To the best of ATSDR's knowledge, however, no adverse health effects have been reported in the study group, to date." This statement refers to a very small group of residents from whom biological samples were collected. This statement should be deleted as it implies that the study group has been actively followed over time. The study group has not been followed up by any government agency, including ATSDR. Follow up of this small study group would be inappropriate as the study was designed to study exposures, not health effects. Although the biological levels were elevated and suggest an elevated cancer risk, the study group is too small to expect to be able to observe that elevated cancer risk.

Response: ATSDR agrees that it should be clearly stated that the group has not been followed up medically and that the small study group would limit the observation of elevated cancer rates. Such a statement has been added to this discussion.

California Department of Food and Agriculture

Comments related to the Petitioned Public Health Assessment for Koppers Company, Incorporated (Oroville Plant) are limited to two areas of concern, namely methodology and recommended action:


Comment: The report has successfully related risk estimates to human health quantitatively, however it acknowledges that duration of exposure has not been incorporated into the exposure assessment. However important exposure duration is for humans, it is also important to interpret International Toxicity Equivalents (ITEQ) in light of the fact that the average life span of backyard poultry may approach a duration of up to ten years and may be of greater risk.

Response: ATSDR was not attempting to relate risk estimates to human health quantitatively. Apart from being convenient as screening values, "risk" estimates generated using regulatory methodology have little or no relevance to the process of human health assessment. They do not necessarily give a realistic prediction of the true risk, which is unknown and may be as low as zero (EPA 1986). In addition, they are often confused by the public with estimates of actuarial risk which, unlike quantitative cancer risk estimates, are based on actual incidence data in humans. TEQs, too, are of limited relevance to the public health assessment, apart from their use in preliminary screening values. The TEF method does not consider many of those pharmacologic factors (e.g., bioavailability, potential for antagonistic interactions, and congener-specific metabolism, half-lives, and effects) that are of paramount importance in any scientifically valid toxicologic evaluation of the probable health implications of site-specific exposures. Also, ATSDR advising against the sale and continued consumption of home-grown eggs at the two index homes would effectively avert the "greater risk" referred to in this comment.

Comment: Normal (control) background levels of dioxin congeners for Butte County have not been presented and so it is difficult to assess the scientific significance of the report findings in light of this important basis of comparison.

Response: Should such background data become available, ATSDR will include it in an addendum to the PHA. However, such data are not pertinent to the central question of whether or not the identified site-specific exposures are likely to have an adverse impact on public health. ATSDR considers that the available information was sufficient to support a reasonable answer to that question, within the limits of current knowledge and technology.

Comment: A site map with the locations of the two index backyard sites and the remaining 23 sites that were sampled would be very useful to correlate with the demographic, water, and soil maps provided.

Response: ATSDR has not been able to identify a map that clearly depicts these locations. However, ATSDR has created a map in Appendix A using a 1989 CDHS reference (Foraging Farm Animals as Biomarkers for Dioxin Contamination) for contaminated egg locations; although the map in this reference is not clear, a general idea of egg sampling locations is can be determined.

Comment: The report states that exposure assessment considered the following routes of human uptake: Ingestion, Inhalation, and Dermal. In addition, the report states that "chemicals detected at comparison values simply require a more detailed evaluation of site-specific exposure conditions." If such a site-specific assessment was done, it would be important to correlate such results with the levels detected in a geographic origin of animals, animal products and animal by-products noted in the report in the interest of selecting specific risk management strategies appropriate at each site.

Response: This comment references ATSDR's assessment methodology, which outlines in general terms how ATSDR evaluates environmental and health outcome data. Section B of the Discussion section presents the findings of the site-specific analysis which focuses on the following five potential exposure pathways (as outlined on page 8):

  • Ingestion and skin contact with PCP.
  • Exposure to smoke during past fire episodes.
  • Ingestion of homegrown farm products.
  • Contact with on-site soils.
  • Contact with surface water and sediment in streams along the site perimeter.

Site-specific exposure information, including exposure assumptions, are included in these discussions (e.g, drinking water assumptions and egg consumption patterns for area residents). It should be emphasized, however, that the primary purpose of a public health assessment is to identify site-specific exposures and determine their probable implications for the health of local residents. The ultimate origin of the substances to which those residents are exposed is of secondary interest only. Such knowledge is not necessary for the determination of probable health effects, but it could be very useful in the effort to determine the most appropriate measures for limiting exposures.

Comment: No mention is made in the report of sampling insects which are a staple food of backyard poultry.

Response: To ATSDR's knowledge, no efforts were made to sample insects, as the latter were not considered to be potential sources of significant human exposure to PCP or dioxin. It is expected that poultry at Oroville/Koppers would most likely be exposed to these contaminants while pecking at contaminated soils. The detected levels in off-site soil, though only modestly elevated, would have been bio-accumulated in poultry foraging on such soils.

Comment: Human disease rates are not adjusted for age, sex, or race.

Response: ATSDR recognized this limitation in interpreting the available cancer registry data. The potential limitations associated with not adjusting the cancer rates for age, sex, and race are briefly described in the public health assessment on page 26.

Comment: Sample size for animal products was not defined and as a consequence the statistical significance of the report findings is not apparent.

Response: The purpose of the public health assessment is to put past, current, and potential future exposures to contaminants potentially related to the Koppers site into perspective for the public. It is a largely qualitative exercise based on data collected from various sources. ATSDR agrees that the representativeness of the available sampling data is an important factor for interpreting results. (Table 5 summarizes the animal product data.) However, it should be emphasized that ATSDR typically bases its health calls on the maximum concentrations detected and worst-case exposure scenarios, both of which are intentionally unrepresentative.


Comment: Education of backyard producers in the use of proper management of livestock and poultry is a major emphasis of current and future risk mitigation recommendations. Government agencies with expertise in poultry management and diseases such as the CDFA are excellent sources of such assistance.

Response: ATSDR acknowledges the assistance that CDFA and other agencies can offer. ATSDR has learned that CDFA is currently participating in a Task Force, led by the CDHS, to develop specific guidance on the proper management of backyard livestock and poultry and to develop a comprehensive educational program to reach all potentially affected populations. Therefore, ATSDR has modified the recommendations/public health action plan in the public health assessment to describe these activities.

Comment: Recommendation #5: "No untested poultry products from these backyard farms should be sold." The report is unclear as to whether or not sales of poultry products from backyard flocks are occurring. If so, it does not explain what mechanisms are in place to prevent untested poultry products from affected backyard poultry from being sold.

Response: The advisories issued by CDHS have advised backyard producers not to sell or consume backyard products. As stated on page 19 of the public health assessment, in light of the uncertainties pertaining to dioxin toxicity and exposure levels, it is prudent public health practice to limit exposures to the extent possible and ATSDR, therefore, recommends that residents continue to adhere to the CDHS advisory. ATSDR recognizes that no mechanism is in place to strictly regulate or require testing of these products and encourages the type of educational programs described in the previous response.

ATSDR's evaluation of possible exposures to the levels of dioxins/furans detected in backyard poultry in the vicinity of the Koppers site presented in the public health assessment attempts to provide readers with some perspective on the likelihood of adverse health effects and what is and is not known about low-level exposures to dioxins/furans.

Koppers Industries

Comment: Page 1, Paragraph 3: It is unclear if "contaminants detected on-site..." refers to all or selected modalities. With respect to breathing zone air, sampling data from industrial hygiene air monitoring studies for the site would not support the presence of benzene as an airborne contaminant."

Response: This paragraph summarizes what was found in on-site groundwater, soil and/or air. Text discussions and accompanying tables clearly indicate which contaminants were detected in which medium. Benzene was detected in on-site groundwater only, based on the sampling data that ATSDR reviewed (see Table 2).

Comment: Page 14, On-site Air: Industrial hygiene air monitoring data collected on a cross section of plant employees during routine operations/activities revealed no elevated levels of BTS/Naph, CTPVs, Arsenic, Copper, Chromium, and PCP.

Response: ATSDR does not typically review industrial hygiene data because the protection of workers falls under the jurisdiction of the Occupational Safety and Health Administration (OSHA). If air data are available in the vicinity of the aeration lagoon, the two wastewater spray field, and spill areas, these data would be helpful in helping to assess possible past exposures to air on and in the vicinity of the site. However, the greatest community concern related to air exposures was possible health effects resulting from releases during fire episodes, for which no air data are reportedly available.

Comment: Page 15, Paragraph 2: Extensive wipe sampling and limited air sampling was conducted on-site immediately following the fire and during associated cleanup activities.

Response: Wipe sampling would provide some additional information regarding residual levels of contaminants following the fire, but would not fill the primary data gap-that is, levels of potential contaminants in air to which area residents and workers may have been exposed at the time of the fires.

Comment: In general, the report appears to be factual and technically sound. However, one area that is not clearly described involves the relationship of dioxins/furans in backyard-raised eggs and livestock to the subject plant site. A transport pathway for these compounds from the Oroville Plant to the subject receptors has not been established. In fact, data summarized on page 16 of the report (from CDHS 1997, 1999; Goldman et al. 1999) indicates that dioxins were detected in samples collected both before and after the fire incident at the site, which was presumed to be the source of these compounds at the time samples were collected. This suggests that there may not be a relationship between the presence of dioxins/furans in eggs and livestock and activities at the Oroville Plant, and that there may be another source of these compounds that is unrelated to the Oroville site.

Response: ATSDR recognizes the fact that pre and post-fire levels in a single cow sample were at similar levels, suggesting the existence of a pre-fire source of dioxins. ATSDR also recognizes that home-produced eggs and poultry revealed elevated levels of dioxins/furans despite the presence of only trace levels of dioxins/furans in areas soils. The latter was not entirely unexpected, however, considering the inevitable bioconcentration of dioxins in chickens pecking at contaminated soils. Note, also, that ATSDR's primary role in this assessment was to provide a public health evaluation of the reported levels of dioxins/furans in home-produced eggs/livestock in response to community concerns--not to perform an extensive evaluation of the potential dioxin/furan source(s).

U.S. Environmental Protection Agency, Region IX

Comment: General - EPA uses the terminology on-property and off-property vs. on-site and off-site (under CERCLA the site is wherever contaminants are found/migrated to).

Response: ATSDR generally uses "on-site" and "off-site" to designate contaminant location. On-site and off-site designations have been changed to "on-property" and "off-property" as appropriate in the PHA, primarily in groundwater discussions.

Comment: Page 5, third paragraph - As of December 1995, EPA determined that the area of PCP-contaminated groundwater had receded to the extent that the off-property Pump and Treat system operation could be suspended. The residual off-property PCP contamination is presently beginning treated by use of enhanced in-situ bioremediation. (Continued use of P&T off-property would have required the installation of new extraction and re-injection wells piped to the P&T to treat the residual plume. Relocating the off-property P&T remains a remote option if other remediation efforts are not successful.)

Response: In-situ bioremediation has been referenced in the PHA as a result of this comment.

Comment: Page 5, fourth paragraph - Change 100,000 to 107,000 cubic yards.

Response: 100,000 has been changed to 107,000 cubic yards as a result of this comment.

Comment: Page 12, first paragraph - Modify last sentence. The remaining 8 drinking water well owners with PCP contamination will continue to be paid an OWID allotment until it is demonstrated that their wells meet the ROD PCP criteria for 12 consecutive months. (Sampling of these eight wells is not performed continually. They will be sampled prior to ending the OWID subsidy. Five of these deactivated drinking water wells are in the area of the residual plume.)

Response: This sentence on page 12 has been modified in the PHA.

Table of Contents


Community Exposures to the 1965 and 1970 Accidental Tritium Releases



ATSDR Responses follow each comment and indicate any changes that were made to the document as a result of the comment, or explains why changes were not made. The comments, which are presented verbatim, resulted in numerous changes from the public release version of the document (dated May 24, 2002).(18)

Reviewer 1 Comments

1. The assessment addresses all potential exposure pathways with one possible exception: the skin absorption pathway. I assume that skin absorption has been included in the inhalation pathway but no statement to this effect was made in the report. This point should be clarified when the report is revised.

ATSDR Response: A section on skin absorption has been added to Section 3 and a dermal absorption component has been added to the total tritium dose. The dermal absorption component is assumed to be equal to the HTO inhalation component.

2. The general model is appropriate. The level of sophistication and detail is consistent with our current understanding of the environmental transport of tritium following an HT release. Good use has been made of the available experimental data to augment the theory. The use of Monte Carlo analysis to estimate the uncertainty in the predicted doses is commendable, although I have reservations (described in the attachment) about some aspects of its application in this case. The model is sufficiently conservative that it is unlikely that the predicted doses will underestimate the true doses.

Having said this, several well-established environmental tritium codes (UFOTRI, ETMOD, STAR-H3, TRIF) are available for calculating doses from short-term HT releases. These codes are based on the available experimental data and in some cases have undergone extensive testing. Given the existence of these codes, use of the semi-empirical approach adopted in the assessment should be justified. The credibility of the assessment would be enhanced by comparing its predictions with those of one or more of the existing codes.

ATSDR Response: The purpose of this analysis is to determine if the exposures represent a potential health hazard. As the overall model "…is sufficiently conservative that it is unlikely that the predicted doses will underestimate the true doses." comparison with the other referenced models is not necessary from a public health perspective.

3. The RASCAL model is a standard Gaussian dispersion model, the type of model normally used to calculate air concentrations following a short-term release. It gives results that are somewhat higher than those of Canadian standard N288.2, which is also a Gaussian model. However, the meteorological data used in RASCAL for both releases are suspect and consideration should be given to repeating the calculations with more appropriate parameter values (see the attachment for more details). The ISCST code is acceptable for calculating air concentrations from a ground level area source. When combined with experimental data on the loss rate of tritium from soil, the model produces time-integrated HTO concentrations in air that are about a factor of 2 higher with respect to experimental data. The intake of tritium via ingestion is determined from monitoring data, which is a good approach in theory but may underestimate the true intakes in this case. The OBT model is not realistic and results in large overestimates of OBT concentrations in foods. The models used to calculate doses from the environmental concentrations are state-of-the-art.

ATSDR Response: The meteorological data have been revised by use of an atmospheric stability category of "E" (which results in slightly higher concentrations than an "F" category). We have also modified temperature and plume rise factors based on newly available data (Peterson and others, 2002; Quarterman, 1965). The RASCAL results have also been compared with output from the HOTSPOT model, which is another Gaussian plume dispersion model. After the above meteorological adjustments, tritium air concentrations and deposition to soil are similar from the two models, although the RASCAL model does provide a slightly higher estimated concentration. Another adjustment to the model, as suggested in other comments, includes a broader range of HTO emissions from soil. These adjustments are described in the text and referenced as necessary. Tritium intake from ingestion will continue to be based on measured monitoring data. Although such measurements are subject to the uncertainties of sampling and analytical error, there is no a priori rationale for assuming modeled tritium concentrations will more accurately represent actual values.

4. Although I'm not an expert in the area of health effects and cannot comment authoritatively on this aspect of the work, I was able to understand the relationship between exposure and health effects as described in the assessment. A considerable amount of experimental data is presented, all of which is consistent in its implication that health effects are unlikely to occur at thelevel of the predicted doses.

ATSDR Response: Comment noted.

5. The accuracy of the predicted doses is in doubt because of problems with the meteorological data, the use of monitoring data to estimate ingestion doses and the underestimate of the uncertainty in the predicted air concentrations (see the attachment for more details). I believe the calculations of inhalation dose should be repeated using the correct values of the meteorological parameters. In these calculations, unnecessary conservatisms should be dropped and realistic standard deviations assigned to the air concentrations so that the output distributions are meaningful. The use of monitoring data to estimate ingestion doses should be justified or the doses should be calculated using a model to ensure they are not underpredicted. In either case, the OBT model should be modified to more closely reflect reality.

The information is presented clearly and is understandable, with a few exceptions:

  • The discussion of HT deposition, oxidation and re-emission is confusing at times. I have made suggestions for wording changes to improve the clarity.
  • Some parameter values (child breathing rates, dosimetric data for HT) and some intermediate outputs of the calculations (time-integrated HTO concentrations in air) are missing and should be included to allow readers to check the calculations.
  • The assumptions behind the OBT model are not spelled out.
  • The Crystal Ball reports of the output distributions are presented without explanation. It's possible to understand them but they are not user-friendly.

ATSDR Response: The predicted doses are recalculated as described above and in the document text. The discussion of HT deposition and OBT and HT dosimetry has been modified as suggested and parameter values for child breathing rates and HT dosimetric data added to document (in main document or appendices, as appropriate). Results of the Crystal Ball output that are included in the main document are discussed as suggested. Potential uncertainty in the estimated HT and HTO air concentrations has been addressed by conducting separate ISC model runs using mean and 95th percentile HTO emission rates. The resulting mean and 95th percentile HTO air concentrations were used to construct a lognormal probability distribution of HTO air concentrations at breathing height (1 m).

6. If the doses predicted in the assessment are accepted at face value, the conclusions and recommendations are appropriate. These doses are clearly below the level at which any health effects could occur. However, as noted above, there are problems with the accuracy of the dose estimates. Until these are revised, no formal conclusions can be drawn regarding health effects or the need for follow-up actions. Having said this, I fully expect that conclusions based on the revised estimates will be the same as those made in the assessment, namely that the exposures are not likely to produce adverse health effects, that doses are below levels of public health concern and that no specific recommendations are warranted.

ATSDR Response: Comment noted. The revised doses are similar to previously calculated doses such that the conclusions and recommendations are unchanged from the public release version of the PHA.

7. Community members near LLNL requested funds to hire an independent technical consultant to review the assessment. Although ATSDR does not have a process for providing funds to the community, it might consider involving the public in the selection of reviewers and sharing the results of the review with the community.

ATSDR Response: Community members were provided with an opportunity to recommend nominees for the peer review process. The person they recommended was solicited by ATSDR as one of the peer reviewers. However, this person, who was eminently qualified, indicated that he no longer accepts or conducts such peer reviews. The community members subsequently obtained an outside grant to pay for an independent expert review and ATSDR has extended the comment period to allow adequate time for those reviewers to complete their work. ATSDR has also recently presented the results of this public health assessment to the Livermore community (February 18, 2003) along with the major revisions prompted by the review process.

8. Incorrect meteorological data appear to have been used for both releases. Site-specific data for the 1965 release have recently come to light (Peterson et al. 2002). Use of these data would result in lower doses than those presented in the report, so it may not be necessary to re-do the calculations for this release. According to Myers et al. (1973), the stability during the 1970 release was class E or F rather than class B as assumed here. If this is the case, the predicted doses would increase substantially. Problems exist with the air concentrations even if the meteorological information used in the assessment is assumed to be correct (see items 26, 34 and 36 under Specific Comments), so it may be best to repeat all the dispersion calculations.

ATSDR Response: The PHA includes revised dispersion and dose calculations based on the newly available meteorological data for both the 1965 and 1970 releases.

9. There seems to be some confusion over the processes of HT deposition, oxidation and re-emission and over the various conversion and loss rates coming out of the 1987 HT experiment at Chalk River. HT is converted to HTO very soon after it diffuses into the soil. Any tritium remaining in the soil after the exposure is in the form of HTO; any HT that is not converted quickly diffuses back into the atmosphere. It therefore makes no sense to talk about HT concentrations in soil; the only tritium present in the soil is in the form of HTO. The HT deposition velocity accounts for both HT deposition and oxidation to HTO. Brown et al. (1988) discuss an effective oxidation rate of 1.5%/h. This value was calculated by dividing the ratio of HTO to HT air concentrations observed at the end of the release at a downwind distance of 50 m by the travel time from the source to 50 m. As such, it is a very specific parameter and can't be generalized to other situations. Ogram et al. (1988) measured the loss rate from soil and found values between 0.5 and 1%/h.

ATSDR Response: The discussion of HT deposition, etc., has been modified as suggested. A broader range of HTO loss rates have been utilized in the calculations based on additional studies, which are referenced in the document.

10. I did some calculations of my own to verify the accuracy of the results in the assessment. The HT concentrations in air predicted by RASCAL were about a factor 2 higher than concentrations from the Canadian Standard dispersion model N288.2, which is also a Gaussian plume model. The reason for this difference is not obvious but the RASCAL results are conservative. Results from the HT release experiments in France and Canada suggest that the ratio of the time-integrated HTO concentration in air to the time-integrated HT concentration is about 0.1. The ratio in the assessment is 0.2, indicating that the HTO concentrations are conservative by an additional factor of 2, perhaps because the initial concentration on day 1 is overestimated (see item 19 under Specific Comments). Thus the inhalation doses in the assessment appear to be the right order of magnitude, although conservative. However, I believe the ingestion doses, which were calculated from monitoring data in the assessment, could be underestimated. Predictions of a simple model suggest that the HTO ingestion dose could be as high as 20 mrem under assumptions similar to those made in calculating the inhalation dose. On the other hand, the contribution of OBT to the ingestion dose is severely overestimated in the assessment. Even with a much increased ingestion dose, the conclusion holds that the releases posed no health risk to members of the public.

ATSDR Response: The RASCAL results have been compared with results from HOTSPOT, which is another Gaussian plume dispersion model. Although the results are similar, RASCAL results are about 20% higher than HOTSPOT results, which is acceptable for the purpose of initial exposure assessment. The revised inhalation doses are based on maximum 12 hour average HTO concentrations rather than the maximum 1 hour average that was used in the previous estimate.

According to representatives of the California Agricultural Extension Service, very little vegetation is growing in the Livermore area during August. Cattle are being fed from stored forage and any garden crops must be irrigated to survive during this time period. Consequently, the measured tritium concentrations and estimated ingestion doses are probably an accurate estimate of exposure.

11. Although I support the use of stochastic assessments, I'm concerned over some of the distributions used in the analysis. These are by necessity subjective, and have to be carefully thought out and justified if the output distributions are to be meaningful. I believe the standard deviations for the HT and HTO air concentrations are much too small. Since these concentrations underlie the inhalation dose predictions, the uncertainties in the doses are likely much higher than indicated in the report. Moreover, I believe most of the distributions should be lognormal rather than normal, as most environmental parameters are (Hoffman 1979, Sheppard and Evenden 1988, Zach et al. 1989, Ott 1990, Blackwood 1992). Finally, the distributions should not be adjusted to conservative values, as recommended in the last paragraph on page 22. Conservative values should only be used in deterministic assessments. If they are used in stochastic assessments, the output distribution no longer has any meaning in terms of the probability of consequences. Obtaining a realistic estimate of this probability is the main reason for doing a stochastic assessment.

If the distributions are widened as much as I suggest, the 95th percentile doses may become larger than desirable. This is the cost of including conservatisms in a stochastic analysis; the solution is to reduce the conservatisms as much as possible. RASCAL appears to overestimate air concentrations but I'm not sure why. The time-integrated HTO concentration in air seems to be overestimated by an additional factor of 2, perhaps because the initial concentration on day 1 is overestimated, or because all the deposited HT is assumed to be re-emitted. The ingestion doses are not an issue if the monitoring data can be defended but doses predicted from model calculations could be reduced by making more realistic assumptions about the diet and location of the cow, the fraction of food that is contaminated and so on.

ATSDR Response: The revised dose calculations are based on a probability distribution of the HT and HTO air concentrations as described above and in the revised document. As indicated, the revised doses have a broader range of uncertainty, although mean doses are very similar. Output from a Gaussian dispersion model should be assumed to be normally distributed, as that is an underlying assumption of the analytical method. It should be further noted that many lognormal distributions of environmental parameters represent inappropriate sampling methods; the data are skewed because of poor spatial representation of the underlying population. A normal probability distribution should be used if the expected values are as likely to be greater than the mean as they are to be less than the mean. We agree that realistic values should be the basis of the probability assessment and have reviewed the bases for the required probability distributions.

12. Page (iii), paragraph 2, line 6 and elsewhere: The 1965 release occurred on January 20, not January 21.

ATSDR Response: The appropriate changes have been made.

13. Page (iii), paragraph 3, line 3: Replace "such as" with "known generically as".

ATSDR Response: This change has been made.

14. Page 8, paragraph 1, line 9: Replace "into HTO, or soil moisture" with "into HTO (tritiated water), which becomes associated with soil moisture".

ATSDR Response: This change has been made.

15. Page 8, paragraph 1, lines 9-10: Replace "Following transformation into HTO, a significant portion of the soil moisture is re-emitted" with "A significant portion of the tritiated soil water may be re-emitted".

ATSDR Response: This change has been made.

16. Page 8, paragraph 3: The references in lines 5 and 7 to the "soil HT concentration" should be changed to "soil HTO concentration". Also in line 5, replace "transformed to HTO" with "emitted to the air". In the same line, I'm not sure what "additional factors" refer to here. Either spell them out or delete the phrase in parentheses. Finally, uptake by plants and animals should be included specifically in the list.

ATSDR Response: This paragraph has been re-written as suggested.

17. Page 9, Figure 3: Replace the text in the middle right ("HTO dispersion…") with "HTO emission from soil and dispersion in air". Also, replace the last sentence in the text to the right in the box with "The emission rate declines as HTO is lost from the soil".

ATSDR Response: These changes have been made.

18. Page 10, paragraph 1: It's stated here that the maximally exposed individuals are assumed to 0.5 miles from the source in 1965 and 1 mile in 1970. Elsewhere (page (iii), paragraph 2) the distance for the 1965 accident is given as "about 1 mile". This point should be clarified because the distances are invoked later in the assessment (Table 2) to justify the use of 1970 results alone, on the assumption that the 1970 doses were higher than those for 1965. This may not be the case if the MEI in 1965 was indeed at 0.5 miles.

ATSDR Response: This was a typographical error and has been corrected. The MEI for both releases was at the 1 mile location. Also, as a result of changing the atmospheric stability, the maximum HT concentration is also at the 1 to 1.5 mile location and has been so noted in the document.

19. Page 10, paragraph 2, line 4: Replace "the dispersion of HTO from the soil surface into the breathing zone" with "the emission of HTO from the soil and dispersion in the atmosphere".

ATSDR Response: The text has been changed as suggested.

20. Page 10, paragraph 4, lines 1-4: Replace the first two sentences with "The loss of HTO from soil to air results in a continually declining concentration of soil and air HTO".

ATSDR Response: These sentences were shortened and clarified.

21. Page 11, paragraph 2: Myers et al. (1973) note that a surface-based inversion probably prevailed over the Livermore Valley at the time of the 1970 release and that the atmosphere was very stable. This means the stability class was more likely F than B, which makes sense for that time of day. Predicted air concentrations for class F could be very different from those presented here for class B. The doses for the 1970 release should be recalculated using class F stability.

ATSDR Response: The previous assumption of a "B" classification was based on slight or moderate solar insolation. As sunrise tables indicate that the release occurred before sunrise, the release was recalculated using both "E" and "F" classifications. All text, tables, and figures have been appropriately modified.

22. Page 11, paragraph 3: It's artificial and possibly confusing to identify inhalation with the instantaneous concentration and deposition with the cumulative concentration. Both inhalation and deposition are calculated in the same way from the (assumed constant) air concentration during the passage of the plume and the release duration. This same remark applies to paragraph 4 on page 34 in Appendix 1.

ATSDR Response: Both inhalation and deposition must include a time component. In order to include a time component in the HT inhalation exposure the concentration must be based on the instantaneous air concentration in order to integrate the breathing rate over a 30 minute period. As the cumulative concentration is already integrated over time, it would be incorrect to integrate a cumulative air concentration over time. No changes have been made.

23. Page 11, paragraph 4: In line 1, replace "cumulative HT soil concentrations" with "cumulative HTO soil concentrations". In lines 2-3 replace the reference to Brown et al. with Ogram et al. (1988). Both Brown and Ogram presented results from the 1987 HT release at Chalk River but the deposition velocity results are in Ogram. In line 6, replace HT with HTO.

ATSDR Response: The HT-HTO changes have been made as suggested. We have also changed the reference and included additional references to HT deposition velocities.

24. Page 11, paragraph 5, line 1: Replace "deposited onto" with "diffused into".

ATSDR Response: This change has been made.

25. Page 12, lines 1- 2: The first 2 lines on page 12 should be changed from "…by Brown et al. (1988), the transformation and re-emission …" to "… by Ogram et al. (1988), the re-emission…".

ATSDR Response: This paragraph has been re-written, the above reference has been changed, and additional references added.

26. Page 12, paragraph 3, line 4: Replace 1.82E-03 Ci-sec/m3 with 1.82E-03 Ci/m3.

ATSDR Response: The value has been recalculated and the units changed.

27. Page 12, paragraph 5: In line 1, replace Brown with Ogram. In line 2, delete "transformation and". In lines 5, 6, 7, 8, 10 and 11, replace "transformation" with "re-emission".

ATSDR Response: This section has been completely re-written and includes these changes.

28. Page 12, paragraph 5, line 4: My own unpublished analysis of the short-term HT release carried out at Chalk River in 1987 suggests that the time-integrated HTO concentration in air arising from re-emission from soil and plants was about 4% of the HT concentration in air integrated over the release period (30 minutes) at a downwind distance of 400 m. The corresponding value from the 1986 French release (2 minute release, 800 m downwind) was 8%. In both experiments the data on HTO concentrations in air covered a period of only 4 days after the release, so the HTO/HT ratios integrated over all time would be larger and a value of 10% is reasonable. The ratio of the concentrations in the present report is about 20%, indicating that the estimate of the time-integrated HTO concentration in air is conservative by about a factor 2.

ATSDR Response: Although we have not recalculated the HTO/HT concentration ratios, the revised dose calculations are based on 12 hour average concentrations rather than 1 hour values. This change should effectively address the over-estimation of the HTO air concentrations.

29. Page 12, last paragraph: Replace "HT" with "HTO" in lines 1, 3 and 4.

ATSDR Response: These changes have been made.

30. Page 13, paragraph 1: It's not clear from the discussion in this paragraph how the highest one-hour tritium concentration from 1991 or 1993 can substitute for hourly meteorological data over a 12-day period. This is explained better in the second paragraph on page 41 and that explanation should be given here. Having said this, I don't understand why it's necessary to go through this process. It may be true that hourly meteorological data are not available for the release periods but the ISCST code is used only in the first hour, for which meteorological conditions are known. Basing the time-integrated HTO concentrations on the meteorological conditions in effect at the time of the release, rather than on the conditions that lead to the maximum concentration, could reduce the conservatism of the HTO inhalation dose considerably.

ATSDR Response: The available site-specific meteorological data for 1970 (or 1965) are not sufficient to run the ISC model. Because the HTO is emitted from the entire footprint of the HT plume, we felt it was necessary to use ISC or a similar model that could integrate breathing zone concentrations from an areally distributed source. The worst case conditions during a five year period are generally considered to be a reliable proxy for the site and time-specific meteorological data. This paragraph has been re-written to include the paragraph from page 41.

31. Page 13, paragraph 3: In line 1, replace "The HT concentration" with "The HTO concentration", and "Brown et al. (1988)" with "Ogram et al. (1988)". In lines 2-3, replace "The HTO concentration" with "The HTO air concentration". In lines 3-4, delete "of HT and HTO".

ATSDR Response: This paragraph has been deleted and figure 4 replaced with 2 figures showing the estimated HTO emission rates and the probability distribution of the HTO air concentration.

32. Page 13, paragraph 4, line 1: What exactly does "HTO concentrations in plant moisture are in equilibrium with soil moisture" mean? That concentrations in plants and soil are the same or just proportional? Spencer et al. (1988) show that the vegetation/soil (0-2 cm) ratio is about 1/3 after the lag of 12-24 hours.

ATSDR Response: The paragraph has been changed to clarify that the concentrations are proportional.

33. Page 13, paragraph 4, line 6: Replace "underestimates the decline of the HT soil source" with "overestimates the decline of the HTO soil source".

ATSDR Response: This sentence has been rewritten as suggested.

34. Page 14, Figure 4: In the heading, in the key, in the axis label on the left and in the caption, replace "soil HT" with "soil HTO".

ATSDR Response: This figure has been replaced as described above (comment 20).

35. Page 14, paragraphs 1 and 2: In a review of the 1965 release, Peterson et al. (2002) state that the wind was from the southwest (230o - 250o) at 3.6 m/s at the time of the release, and that the stability was class B. This information could be used to generate new predictions for this release, although this is perhaps not absolutely necessary. The present analysis is conservative since the lower wind speed and greater stability will result in higher concentrations than would be obtained using the real data.

ATSDR Response: The newly available information has been used to recalculate the 1965 doses.

36. Page 15, paragraph 2: In line 4, replace "cumulative soil HT loading" with "cumulative air HT concentration". In line 5, replace 3.0e-4 with 4.0e-4. In lines 5-6, replace "the HT to HTO conversion rate" with "the HTO loss rate from soil". The same comments apply to the last sentence in the caption to Table 1.

ATSDR Response: These sentences have been rewritten as suggested.

37. Page 15, paragraph 3: Tables A-2 and A-3 show that the predicted time-integrated HT concentration in air was lower in 1965 than in 1970 close to the source but higher at 0.7 km and beyond. This is to be expected for an elevated release and the release rates, wind speeds and stabilities (class B in 1970 and class C in 1965) assumed in the assessment. Since the HTO emission rates from soil are proportional to the integrated HT concentrations, I expected to find the same pattern in Table 1. But this was not the case, as the table shows higher emission rates for 1970 than for 1965 at all downwind distances. I was able to reproduce the values for 1970 and believe these to be correct. But I wasn't able replicate the 1965 values and suspect an error here.

ATSDR Response: There was an error in entering the 1965 cumulative HT (Table A-3) values. The plume centerline (30o) cumulative HT concentrations should have been 5.6, 17.8, 14.0, 7.3, 4.3, 2.4, 1.2, and 0.7 (Ci-sec/m3) for distances of 0.1, 0.2, 0.3, 0.5, 0.7, 1.0, 1.5, and 2.0 miles (respectively). This error led to another error in calculation of the HTO emission rates in Table 1. Using the correct values, relative to the 1970 values, the 1965 cumulative HT concentrations and HTO emission rates were higher at on-site locations, but lower at off-site locations (1 mile or more from stack). All of these values have been recalculated and the tables re-written using the actual weather data.

38. Page 15, paragraph 3, line 3: Since the atmosphere was more stable (class C) in 1965 than 1970 (class B), there was decreased atmospheric dispersion in 1965, not increased dispersion as stated. The plume would have covered a smaller area of ground in 1965 than in 1970 and a more accurate emission rate would have been obtained with smaller rectangles rather than larger ones.

ATSDR Response: As stated above, the stability classes for the releases have changed and all concentrations, etc., have been recalculated. The text has been re-written to reflect these changes. Also, because of greater certainty regarding meteorological conditions in 1965 and because the cumulative HT concentrations and HTO emission rates are lower for the 1965 release than for the 1970 release, we have not re-calculated the HTO dispersion using ISC for the 1965 release. For the previous calculations, we used the larger footprint areas because of greater uncertainty in the wind directions.

39. Page 16, paragraph 1: Give an idea of the fraction of the time-integrated concentration that is missed by stopping at 12 days rather than at infinity. This will justify stopping at 12 days, which otherwise seems arbitrary.

ATSDR Response: A justification for the 12 day exposure duration is given on page 17 (paragraph 3) and in Figures 4 and 5 (public release version). The 12 day period accounts for >95% of the soil HTO and based on HTO emission rates more than 4 environmental half lives. Inhalation doses for day 13 are less than ˝ of the daily background rate and we believe are accounted for in the chronic doses which are added to the acute doses estimated in this evaluation. A footnote with this justification has been added to this section of the revised PHA.

40. Page 16, paragraph 3: Change the first sentence to read "The 30-minute HT inhalation dose is calculated from the maximum instantaneous HT concentration and the release duration". In lines 3-4, replace "as the HT deposited on the soil surface is transformed and dispersed through the environment" with "as the HTO in the soil is dispersed through the environment".

ATSDR Response: These changes have been made.

41. Page 16, paragraph 4: In item 6, replace "Table 1" with "Table A-1". In item 7, replace "Table 2" with "Tables A-2 and A-3". Replace item 9 with "HTO loss rates from soil on the basis of measured data (Ogram et al. 1988)".

ATSDR Response: These changes have been made.

42. Page 17, paragraph 3, line 9. The loss of HTO from the soil follows first order kinetics. The 1% loss rate applies to the HTO that remains in the soil at a given time, not to the initial concentration at the end of the exposure. After 24 hours, the amount left in the soil is 0.9924 = 0.786 times the initial concentration, not 0.76. This is a small difference after one day but it becomes more important at longer times. For example, if the loss were applied linearly, there would be no HTO left in the soil after 100 hours, when in fact the concentration is still about one-third of the initial concentration.

ATSDR Response: This section has been re-written and no longer includes these sentences. We have also included a specific statement describing the HTO loss rate as an exponential decay parameter.

43. Page 17, Table 2: I tried to reproduce the numbers in this table but found I didn't have enough information to do so. In a report of this kind, it is essential to provide enough information to allow readers to check the calculations if they have a mind to. Dosimetric data for HT should be added to the report, plus breathing rates for children and predicted time-integrated HTO concentrations in air from the ISCST code.

ATSDR Response: HT dosimetric data, child breathing rates and time-integrated HTO air concentrations have been added to the appendices as suggested.

44. Further to Table 2: The HT dose rates at 1 and 2 miles should be in the same ratio as the HT concentrations in air at these distances. The concentration ratio from Table A-2 is 3.27/0.962 = 3.4 whereas the dose ratios range from 1.67 for children to 2.0 for adults. There seems to be something wrong here. Moreover, I couldn't reproduce the HTO adult dose at 1 mile. Taking the initial HTO concentration in air as 2.1 x 10-6 Ci/m3 from Appendix 3, and assuming a 1%/hour decrease, I calculate the integrated HTO air concentration over 12 days to be 0.71 Ci s/m3, or 2.64 x 1010 Bq s/m3. With a breathing rate of 1.41 x 10-4 m3/s, the total activity taken into the body due to inhalation over the 12-day period is 3.71 x 106 Bq. Adding 50% to this to account for skin absorption gives a total intake of 5.57 x 106 Bq. Using this as the "concentration" in the equation on page 40 yields 4.88 x 10-5 Sv or 4.88 mrem. The value given in Table 2 (19 mrem) is 3.9 times higher than this and so is conservative but the difference needs to be accounted for.

ATSDR Response: We believe the differences in the dose ratios may be an artifact of rounding error. Rounding off of these small numbers can provide significant changes in the ratios. As we have completely recalculated the doses and concentrations, if there was an editing or transcription error in the previous table, it has been corrected by insertion of the revised values. With regard to the overall magnitude of the inhalation doses, we believe the high values are largely a result of using 1 hour maxima (for each day). As stated in the response to comment 3, the revised dose estimates are based on use of the 12 hour average values and are more similar to those estimated in this comment.

45. Two final points regarding Table 2. (i) If doses from the 1965 release are indeed higher than those from 1970 at 1 and 2 miles (or if the MEI is 0.5 miles from the source rather than 1 mile), then the numbers in the table may have to be based on the 1965 data. At the very least, the phrase "due to increased dispersion" should be removed from line 5 of the caption since dispersion was less in 1965 than in 1970. (ii) Do the doses include the contribution from skin absorption? If not, they should. If they do, this should be mentioned so it's clear that an important exposure pathway hasn't been missed.

ATSDR Response: The MEI for both releases was located 1 mile from the tritium facility stack. The newly available 1965 meteorological data indicates that the centerline of the 1965 plume was ~60o from the tritium facility stack. Based on this revised wind direction, we have recalculated the number of people and distance to the potentially exposed population for the 1965 release. As with the previous calculations, potential doses from the 1970 release are higher and more people were potentially exposed. Consequently, we are continuing to base our maximum exposure estimations on the 1970 release. The reason that the 1965 doses are lower than the 1970 doses is because of decreased dispersion of the 1965 release. The caption has been modified to clarify that the lower dispersion is for the 1965 release. A skin absorption factor (100% of HTO inhalation) has been added to the total doses.

46. Page 19, paragraph 1: Although I support the use of monitoring data to calculate doses in general, the data must be reliable. More information is needed before you can convince me that this is the case here. Exactly where were the plant and milk samples taken? If they're going to be used to estimate doses, they must be taken on the plume centerline (as determined by measurements and not by wind direction) at the downwind distance of the maximally exposed individual. When were the samples taken? They must be taken when the plant concentrations peak. The fact that the milk concentrations were so much lower than the vegetation concentrations makes me even more suspicious. As a check on the values given in the report, I estimated ingestion dose using a simple model that's described in an Appendix to these comments. I found doses of about 20 mrem for ingestion of fruits and vegetables, two orders of magnitude higher than the values based on the monitoring data. Two conclusions are possible from this: either the models are overpredicting or the monitoring data is underpredicting. I suspect it's a combination of the two, but some recognition should be given in the report to the possibility that the ingestion doses could be higher than indicated from the monitoring data.

ATSDR Response: As stated in the response to comment 3, the actual measurements of vegetation and milk appear to be representative of growing conditions at the time of the 1970 release. Winter conditions, as during the 1965 release, are much wetter such that cattle may be consuming pasture grasses during that time. However, there are a limited number of cold weather garden crops that will be harvested and consumed during that timeframe such that the overall ingestion dose from the 1965 release is unlikely to be significantly higher than that estimated for the 1970 release. Also, newly available 1965 weather data indicates that it rained shortly after the 1965 release such that soil HTO concentrations, and water taken up by plants, would be greatly diluted.

The specific locations and timeframes of the measured samples are uncertain. Although we used those locations with the highest measured values, it is possible that higher values may have been present in the environment. We have acknowledged that uncertainty by assuming that the highest measured value represents the 90th percentile value. This distribution assumes that 10% of the samples will have higher values. On the basis of all of these factors and the human urinanalysis of the people living in the plume area that did not indicate any detectable tritium exposures, we are very confident that the models over-estimate the total tritium doses.

47. Ingestion doses are not calculated separately for the 1965 release on the assumption that air concentrations were higher in 1970. But, as pointed out above, the 1965 concentrations were higher at and beyond 0.7 miles, and the 1965 concentrations at 0.5 miles were higher than the 1970 concentrations at 1 mile. Thus the monitoring data from 1970 would underestimate the ingestion doses in 1965. Separate estimates should be carried out for 1965.

ATSDR Response: As indicated above, the 1965 concentrations were higher at on-site locations, but lower at all areas of potential off-site exposure. Although there may be some difference in the ingestion doses for the two releases, overall the 1970 release would have resulted in higher total tritium doses. As those doses are not likely to result in adverse health effects, there is no need for separate estimates of the 1965 ingestion doses.

48. Page 19, paragraph 2: The monitoring data likely underestimate the ingestion dose from milk as well as from fruits and vegetables. It would be desirable to model this dose, but reliable concentrations in milk are difficult to calculate because of the uncertainty in the amount of tritiated water taken in by the cow with its feed. Also, milk concentrations would not follow the air concentration. They would build up initially as the air concentrations drop off and decrease more slowly than the air concentrations once they reach their peak. But the time-integrated concentration calculated assuming a 1%/h decrease would probably give a result that wasn't far off reality. Using the vegetation concentrations from the Appendix, an ingestion rate of 40 L/d vegetation water by the cow, a fraction of daily intake appearing in 1 kg of animal produce of 0.014 d/L and a milk ingestion rate of 1 L/d, the predicted ingestion dose is about 37 mrem, or about the same as the modeled dose from fruits and vegetables. The possibility that ingestion doses from milk could have been substantially higher than those based on monitoring data should be acknowledged.

ATSDR Response: Language regarding uncertainty of the ingestion doses has been added to the text.

49. Page 19, paragraph 4: From the description of the OBT model given here, it appears that the OBT concentration in plants is assumed equal to the HTO concentration and that the total activity of tritium ingested in the form of OBT equals the activity ingested as HTO. Neither of these assumptions is defensible. OBT concentrations will behave very differently than HTO concentrations, starting out very low and showing a steady increase over the 12-day assessment period rather than a decrease. The intake of tritium in the form of OBT will be much less than the intake of HTO because most plants are primarily water. Using data from the 1987 experiment at Chalk River (Spencer et al. 1988), I calculated the dose from ingestion of OBT in fruits and vegetables as 0.7 mrem over the first 12 days. This is trivial compared to the modeled ingestion dose from HTO. The OBT concentration and the corresponding ingestion dose will stay relatively high over the next month or two but would still contribute relatively little (~ 20%) to the total time-integrated ingestion dose.

ATSDR Response: We have relied on the findings of an expert panel to determine the relative dose contribution from OBT (ATSDR 2002). The revised calculations and dose estimates assume an average increase of 32% of the HTO ingestion dose due to the contribution of OBT. Consequently, about 2/3 of the resulting ingestion doses are due to HTO ingestion and about 1/3 due to OBT ingestion.

50. Page 20, Figure 6: In the second line of the caption, replace "most likely" with "average". The most likely dose occurs at the peak of the distribution.

ATSDR Response: The caption has been changed.

51. Page 21, Table 3: The ratio of mean to 95th percentile doses is the same for adults and children for the inhalation doses but not for the ingestion dose. Why is the 95th percentile so close to the mean for the adult ingestion dose?

ATSDR Response: The ratios of mean to 95th percentile values for adult and child ingestion doses are similar for the revised dose calculations. An exception to this which provides an example of rounding error is in Tables 2 and 3 of the revised document. The 1-mile adult 30 minute HT inhalation dose shows a mean value of 0.1 mrem and 95th percentile value of 0.2 mrem. The ratio of 95th% to mean value is 2, however, the mean value was rounded up from 0.05 and the ratio using unrounded values is 4. The rationale for using rounded numbers is our underlying belief that due to inherent uncertainties, none of the calculated numbers have any significance beyond one, or possibly two, significant digits.

52. Tables 2 and 3: Make sure the numbers in these tables are consistent. For example, the 95th percentile 12-day HTO inhalation dose to a child is given in Table 2 as 130 mrem but in Table 3 as 134. Also, the total doses in mSv in Table 3 are not exactly 0.01 times the doses in mrem.

ATSDR Response: These tables and their values have been corrected for consistency.

53. Page 21, paragraph 3: Why are Myers' estimates of dose from milk ingestion so much higher than the present estimates? His approach and results should be critiqued or readers may wonder if he is right and the present estimates wrong.

ATSDR Response: Myers et al. (1971) estimate of dose from milk was based on a hypothetical cow consuming pasture grasses contaminated at the maximum measured HTO concentration and an assumed moisture content of 4 L/m2. The Myers report states that "a) no other locations reaching this level of contamination were found and b) it is unlikely that any such locations were missed because of the extensiveness of the survey." Our dose estimates are based on milk measurements from the real dairy cows present in the plume area. Statements outlining these differences in approach and the resulting doses have been added to the text.

54. Page 22, paragraph 3: Replace the third sentence with "Plants are a primary conduit for transferring tritium from soil to air, but the air concentration includes the tritium that is actually in the plants."

ATSDR Response: The sentence was changed as follows: Transpiration from plants and evaporation from soil are the primary conduits for transferring tritium from soil to air.

55. Page 22, paragraph 4, lines 9-15: No credit can be claimed for dilution of plant concentrations by uncontaminated water when ingestion doses are calculated from observed plant concentrations. But this argument can be used to maintain (legitimately) that modeled ingestion doses are too high.

ATSDR Response: This is true if the vegetation measurements had included garden or food crops. The Myers report indicates that "The vegetation samples were mostly varieties which remain green during the dry season…". Our statement indicates that garden crops are unlikely to remain green during the dry season. No changes to this sentence have been made.

56. Page 29, paragraph 2, lines 7-8: The sentence beginning "Due to increased dispersion conditions …" needs to be changed to reflect whatever resolution is reached regarding the relative concentrations of the two releases.

ATSDR Response: The phrase "increased dispersion" has been changed to "meteorological conditions".

57. Page 34, paragraph 3, line 7: Delete "decreasing atmospheric dispersion and", since dispersion is independent of release height (in the model at least). Also, the effect of release height for these stabilities would extend only a few hundred meters downwind, not the 3 miles claimed.

ATSDR Response: This section has been rewritten. Because of newly available information on stack parameters, plume rise has been included in the revised RASCAL calculations. However, total dispersion of a plume is highly dependent on release height, which significantly effects the distribution of ground level contaminant concentrations.

58. Page 34, paragraph 4, line 6: Replace "Q (Ci-sec)" with "Q (Ci)". Also, in the last line, replace Brown with Ogram.

ATSDR Response: These changes have been made.

59. Page 36, Table A-4: The values of wind speed, direction, stability class etc. are missing from the bottom of the table.

ATSDR Response: The revised table includes the complete input summary of the RASCAL evaluation.

60. Page 38, paragraph 2, line 3: A southwest wind was assumed for the 1965 release, so the rectangles should cover directions 10º to 50º (as indicated on page 15) rather than 180º to 220º.

ATSDR Response: Wind directions for the 1965 release have been revised on the basis of newly available data. This paragraph has been changed accordingly.

61. Page 38, paragraph 2: Replace lines 6-10 with "HTO emission rates for each rectangular area were derived from the time-integrated HT concentrations in air, the HT deposition velocity and the HTO loss rate from soil, assumed to be 1%/h (Ogram et al. 1988)."

ATSDR Response: This paragraph has been changed as suggested.

62. Page 40, the equation for calculating dose: I'm not an expert in dosimetry so the following comments may not be relevant. But if I can't understand this equation, others may not be able to either. The first term in the equation is defined as a concentration but has units of Bq rather than Bq/m3 or Bq/L. Should this be total intake rather than concentration? The units of the tritium decay energy should be MeV/disintegration to make the equation dimensionally correct. The parameters in the equation other than concentration appear to act like a dose coefficient, but when multiplied out have a value of 8.76 x 10-12 Sv/Bq, a factor 2 lower than the commonly accepted value of 1.8 x 10-11. Has the dose coefficient been revised down recently? Finally, it's not clear how to apply the equation for HT.

ATSDR Response: The equation referenced shows the amount of energy deposited in a defined mass from a defined amount of tritium. The equation is used for the total intake of tritium regardless of the concentration. If the concentration is known, extra coefficients are used to estimate the total intake. These could include liters consumed per day, amount of air inspired during a day based on activity levels, or the amount of food consumed. As for the math, we used the following values and approximated the ICRP value of 1.8E-11 Sv/Bq:

27 picocuries = 1 Bq
1.6 E-13 J/Mev
tritium decay energy (average) = 6
seconds per year = 31.5 E 6

multiplying these numbers out gives 8.2e-10 J/Bq in one year

a Sievert is defined as 1 J/kg and a standard "man" mass is assumed to be 70 kg.

If you divide 8.2 e-10 by 70, then one obtains 1.8E-11 Sv/Bq.

63. Page 41, paragraph 2: Replace the last 3 sentences with "The declining air concentrations are due to the loss of HTO from the soil, which is the source of activity for the air".

ATSDR Response: This paragraph has been changed as suggested.

64. Page 41, paragraph 3, line 2: Replace "dispersion from soil" with "the loss rate from soil".

ATSDR Response: This paragraph has been changed as suggested.

65. Page 41, paragraph 6: It's likely that the uncertainties in environmental HT and HTO concentrations and in the assumptions about the ingestion pathways are much larger than the uncertainties in the dosimetric parameters and would dominate the uncertainty in the dose if they were taken into account. This should be pointed out here.

ATSDR Response: The revised sensitivities show that, as suggested, the environmental parameters account for most of the variance in the dose estimates. This paragraph has been changed accordingly.

66. Page 43, bottom distribution: This should be the "initial soil HTO concentration", not the "initial soil HT concentration".

ATSDR Response: This forecast label has been changed as suggested.

67. Page 44: Add units to the distribution at the top of the page. Also, the description of the lower distribution sounds very much like that of the lower distribution on page 42, but the numerical values are different. What is the difference between these two distributions?

ATSDR Response: This forecast label has been changed as suggested. The cumulative HT concentration forecasts are redundant, however, the "cumulative air act (HT)" forecast is not used in the calculations and has been removed from the report.

68. Page 45, upper distribution: The mean value of the distribution is given as 0.0238 mrem but the average 30-min HT dose for a child in Table 2 is 0.05 mrem. What is the reason for the difference?

ATSDR Response: The Crystal Ball simulations were run iteratively for adult and child intake rates and body weights, and for concentrations at the 1 and 2 mile locations. The page 45 forecast is for an adult exposure. The revised Crystal Ball report has been labeled to clearly identify the specific parameters it is simulating.

69. Pages 46-49: I accept the distributions for the dosimetric parameters since those were set by a panel of experts (but units should be added to the distribution for tritium decay energy.

ATSDR Response: The decay energy units of MeV have been added to the figure.

70. Cumulative HT air concentration/Q: The shape and standard deviation of the distribution must be justified. I believe the standard deviation is much too small. The method of expert elicitation was recently used to estimate as rigorously as possible the uncertainty in the predictions of Gaussian dispersion models (NUREG 1994). Results of this study indicate that the ratio of the 95th to 5th percentiles of the centreline concentration is about a factor of 10 for downwind distances of 500 to 1000 m. A large uncertainty would be expected in the present case from the uncertainty in stability class alone. For the 1970 release, estimates of stability ranged from B to F, which can generate differences of more than a factor of 10 depending on the downwind distance. As a secondary point, the units should be indicated for the distribution.

ATSDR Response: As the revised calculations are based on observed weather conditions, we feel there is considerable justification for use of the E (or F) and B (or C) atmospheric stabilities. In both cases we have used the stabilities that lead to the highest estimated concentrations in the areas of potential maximum exposure. The shape (distribution) of the HT concentrations must be assumed as normal because such normality is the underlying basis for the Gaussian dispersion (concentration distributions are normally distributed within the plume). While the standard deviation of the assumed HT distribution may be small, because it is a normal distribution, the simulation results are based principally on the mean value.

We have compared the results of the RASCAL dispersion with another Gaussian model (HOTSPOT) and found that while the results are similar, RASCAL results are about 20% higher than HOTSPOT results. A validation study of the RASCAL model also found that predicted concentrations over-estimated measured concentrations (Ramsdell, personal communication). Consequently, we are confident that the RASCAL results do not under-predict actual concentrations and that the actual doses will be less than the predicted doses. Unit labels will be added to the Crystal Ball report as suggested.

71. HTO air concentration: The shape and standard deviation of the distribution must be justified. I believe the standard deviation is much too small. The HTO concentrations reflect not only the uncertainty in the HT air concentrations but also the uncertainties in the deposition velocity, the re-emission rate and the dispersion model in ISCST, all of which will be large.

ATSDR Response: The derivation of the HTO concentration distribution has been changed. As you suggest, the revised HTO concentration distribution is based on cumulative HT concentrations (normal distribution), HT deposition velocity (assumed normal distribution), and the HTO re-emission rate (exponential distribution; the HTO re-emission rate has been termed the HTO loss rate in order to differentiate it from the HTO soil emission rate used in the ISC model). Multiplication of these factors produces an apparently lognormally distributed HTO soil emission rate (Ci/sec-m2). The mean (50th%) and 95th% values of that emission rate distribution were used in separate ISC model runs to produce mean and 95th% HTO concentrations. The resulting mean and 95th% concentrations were then used to describe a lognormal HTO concentration distribution which was used to estimate doses. This method has been described in the revised document.

72. HT deposition rate: The literature contains many values of the deposition velocity apart from those reported by Ogram et al. (1988). These should be considered in setting the mean and the standard deviation. Also, I believe the distribution should be lognormal rather than normal.

ATSDR Response: As suggested, we have broadened the references on which the HT deposition velocity is based. However, these additional references, including Sweet and Murphy (1981), Dunstall et al. (1985), Spencer and Vereecken-Sheehan (1994), and a review by Brown et al. (1996) confirm that our assumed distribution of rates is consistent with a variety of measured or modeled rates. Additionally, there does not appear to be any a priori basis for assuming a non-normal distribution of those rates.

73. Vegetation activity: More detail should be given on the number, location and timing of the measurements that form the basis for this distribution, so the reader can judge whether the standard deviation is reasonable or not. I believe the distribution for this parameter should be lognormal rather than normal.

ATSDR Response: To the extent possible, based on the locational descriptions of the samples, we have used those sample results most likely to represent the plume footprint. As previously stated, because of growing conditions at the time of the 1970 release, we believe these measured samples are a reasonable indicator of plant tritium concentrations. More complete information on sample locations and times are not available. If we had chosen to include the many lower concentration and non-detected results, the resulting distribution would appear lognormal. However, as the HT plume must be considered normally distributed, it is reasonable to assume that plant moisture concentrations in the plume footprint are also normally distributed.

74. Milk concentration: A distribution based on just 3 points has no credibility. Since the cows are eating the local vegetation, the distribution for milk should be at least as wide as the distribution for vegetation and should be lognormal rather than normal.

ATSDR Response: There were many more milk analyses. There were, however, only three detectable tritium measurements. As previously stated, these results are based on the cows from dairy farms in this area. As cows would not be consuming pasture grasses during this time period, inclusion of the dairy component is conservative.

75. Adult and child body mass: I have no expertise here but I would have thought body mass would be better described by a lognormal than a normal distribution. Also, the units should be given for these distributions.

ATSDR Response: The revised simulations and dose estimations assume lognormal distributions of body masses and breathing rates.

76. Ingestion Dose Estimates from a Simple Model

Data from Spencer et al. (1988) collected during the short-term HT release experiment at Chalk River in 1987 show that, after a lag time of a few hours, HTO concentrations in vegetation were consistently a factor of about 3 lower than concentrations in the 0 - 2 cm soil layer. This provides a means of predicting concentrations in vegetation, from which ingestion doses can be calculated.

The initial HTO content of the soil is 2.5 x 10-3 Ci/m2 = 9.3 x 107 Bq/m2 (Figure 4 in the assessment). Following an HT release, HTO concentrations in soil drop off exponentially with depth with a scale length of 2.3 cm (Taeschner et al. 1995), so the HTO content of the top 2 cm of soil is 58% of the total, or 5.4 x 107 Bq/m2. Assuming a water content in midsummer of 10% by volume, the total amount of water in the top 2 cm is 2 L/m2. The initial HTO concentration in soil water is therefore 2.7 x 107 Bq/L. Assuming a loss rate of 1%/h, the integrated concentration in soil water over 12 days is 1.1 x 108 Bq d/L. The time-integrated HTO concentration in plant water is a factor 3 lower than this, or 3.5 x 107 Bq d/L. With an intake rate of 0.3 L/d vegetation water, the total tritium activity taken in through ingestion of fruit and vegetables is 1.1 x 107 Bq. Using a dose coefficient of 1.8 x 10-11 Sv/Bq, the ingestion dose becomes 2.0 x 10-4 Sv or 20 mrem. This is substantially larger than the dose estimated from the monitoring data, and comparable to the inhalation dose. It's unlikely that these doses would have been achieved in reality but they were obtained with a model similar to that used for the inhalation dose.

ATSDR Response: We agree that the above calculated doses would be unlikely. Further, we believe that basing the ingestion dose estimates on measured milk and vegetation tritium concentrations provides a more realistic assessment of potential doses. Considering that human biological measurements failed to detect tritium exposure from the 1970 release, it is more important to communicate the conservatism of the inhalation doses rather than to equally over-estimate the ingestion doses.

Reviewer 2 Comments

77. Determining the affected population by estimating the total (55 and 52 persons) and then using census data to estimate numbers of women, children, elderly, etc. doesn't seem reasonable given the small numbers of people that were exposed. Maybe, since the assessment only differentiates child and adult doses, these statistics aren't necessary?

ATSDR Response: Part of ATSDR's task in conducting a public health assessment is to determine how many people may be exposed to the site-related environmental contamination and if there may be some portion of that population that is particularly susceptible to those contaminants. As there are no population data available for the limited areas of these air plumes, the identification of buildings and use of average per household occupancy is a reasonable estimate of the maximum exposed population within the plume areas.

78. There is no mention as to why HT is the only form of tritium released from the stack at LLNL. A paragraph describing why this is the case would be informative.

ATSDR Response: All of the available accident reports and descriptions state that the releases were comprised of HT. As this was the form supposedly contained within the vessels that leaked, there is no basis for assuming that any other form of tritium comprised a significant proportion of the releases. The above sentences have been added to the document.

79. The assessment does not include the person(s) that lived in the area during both releases.

ATSDR Response: The doses are summed to provide annual dose estimates and compared with annual dose limits. As the biological half life of tritium is less than one year, there is no long term dose contribution and there would be no cumulative effect from the different releases. Additionally, the revised document includes newly available weather data that indicates that the two plumes did not affect the same area so it is unlikely that any individual was exposed to both plumes (via a residential exposure scenario).

80. Page 15 (end of 3rd full paragraph) - it is stated that the air concentrations from the two releases are modeled as equal, but this finding isn't stated anywhere else in the document, except in that sentence (it sounds as if the reader should already have known this to be the case).

ATSDR Response: The revised dispersion calculations indicate that doses for the 1970 release were likely to have been greater than those from the 1965 release. The document has been revised to indicate that all dose estimates and evaluation of potential adverse health effects are based on the larger 1970 dose estimates.

81. Page 16 - mention is made of breathing rate and how it was estimated, but there is no differentiation between adult and child breathing rates. This needs explanation.

ATSDR Response: The phrase "adult and child breathing rates" has been added to the text. The assumptions concerning child breathing rates have also been added to the Crystal Ball report in Appendix 4.

82. Page 18 - why didn't the author's provide uncertainty for the fruit consumption variables? (several other variables are not included in the uncertainty analysis).

ATSDR Response: The food ingestion rates (fruit, milk, and vegetables) are based on recommended values from the EPA Exposure Factors Handbook (EPA 2002). We have re-run the Crystal Ball simulations using these values as the mean of a normal distribution and there is no difference in the ingestion dose. Consequently we have not changed the point value ingestion rates into probability distributions.

83. Page 19 - what data supports the use of normal distributions for tritium concentrations in vegetables and milk? Environmental concentrations are most often distributed lognormally.

ATSDR Response: The distributions of tritium concentrations in vegetation and milk are based on use of samples from the plume footprint using the lowest measured values as the 10th percentile and the highest value as a 90th percentile of a normal distribution. A normal distribution was used because there is no a priori reason for assuming that the mean will be skewed between these upper and lower values. Additionally, the initial dispersion of the HT plume is based on a Gaussian (or normal) model which assumes that concentrations in the plume are normally distributed. Consequently, HT distribution is assumed normal. Also, although some environmental and biometric parameters are lognormally distributed, many apparent lognormal distributions are an artifact of inadequate sampling.

84. Page 24 (bottom) - children may have increased breathing rates, but their volumes are probably smaller - what is the minute volume of adults vs children?

ATSDR Response: The minute volumes of adults vs. children are 0.011 m3/minute and 0.008 m3/minute, or 15.3 m3/day and 11.4 m3/day, respectively. The change in relative doses is based on the ratio of the breathing rates and body weights.

85. A normal distribution is used for HT deposition rate, but typically uniform or loguniform distributions are used for these type variables. The use of normal is highly biasing, unless there are data supporting such a distribution; the supporting data should be presented or referenced.

ATSDR Response: Additional references to measured or modeled HT deposition velocities have been added to this section. The distribution we have assumed uses the range of rates included in these references. The use of a uniform distribution requires knowledge of the upper and lower boundary values. Deposition velocities have been found to vary with soil type, soil moisture, vegetation cover, and season. Considering the range of parameters that affect deposition rates, it is unlikely that the measured velocities capture the entire range of variation. Likewise, when integrating the range of velocities over a plume footprint, is seems reasonable to assume that the velocities will approach an average value. The normal distribution assumes that values higher than the mean are just as likely as values lower than the mean.

86. Page 12 (end of 3rd full paragraph) - this is a repeat of what appeared on the previous page.

ATSDR Response: This section has been re-written.

87. Page 12 (4th full paragraph, next to last sentence) - what is "a uniform distribution with a value of 1%/hr"?

ATSDR Response: This uniform distribution is essentially a single point value. On the basis of additional empirical values (as referenced in revised document) the HTO emission rate has been changed to an exponential distribution ranging from 0 to 8%/hour with a mean value of 1%/ hour.

88. 1% per hour does not translate into 24% per day - the 1% refers to the loss rate of an amount of material present at a given time - that value shouldn't be multiplied by 24 hours in a day to get %/day. Carrying this same logic to a weekly value gives us 168%/wk?

ATSDR Response: You are correct in that a 1%/ hour HTO loss rate is not equal to a loss rate of 24% /day. The percentage loss depends on the magnitude of the time step. Using an hourly time step, the percentage loss for the first day would be ~21%. As we are calculating daily doses, we made the simplifying assumption that 1%/ hour is equal to 24%/day for the first day (the percentage loss for the second day is ~18% of initial total and so on). The revised document includes a figure showing the percentage loss over 12 days (>95% of initial total) and an explanation of this assumption. As explained in the document, the absolute magnitude of the HTO loss rate does not have a significant effect on the total dose. The use of an hourly time step would increase the number of days required to deplete the HTO soil source by 95%, but would also cause a commensurate decrease in the HTO air concentration such that the total dose would be unchanged.

89. It is stated several times that the "most probable" doses are determined by a Monte Carlo simulation technique. "Most probable" doesn't mean average of the distribution, but it occurs where the distribution peaks. Also, the Monte Carlo is done such that the authors are estimating uncertainty (in a limited way) of the maximum exposure (and they go on to say that the doses are probably greater than anyone actually received). So, these are not "most probable" doses. Likewise, when 95th percentiles (and distributions) are given, those are the 95th percentile estimates of the maximum dose, not 95th percentile estimates of the expected dose. This could be quite confusing to the public.

ATSDR Response: The revised document uses the term "average (or mean) dose" to describe the estimated doses. We have further included a chart of the probability distribution of the estimated doses to show that the mean doses over-estimate the "most probable" doses.

90. Page 21 (top) - only the laws applicable to LLNL should be stated regarding dose limits.

ATSDR Response: The Livermore community is very knowledgeable with regard to regulations and laws pertaining to radiological releases and dose limits. In this case, it is necessary to explain why specific laws are not applicable, as well as those that are applicable.

91. Biological half-life of tritium varies between 1 and 40, with a triangular distribution? This parameter is expected to be lognormally distributed over a population.

ATSDR Response: A triangular distribution was used for the Monte Carlo simulations for the biological half-life of tritium. The conditions for our usage of this type of distribution are as follows. Limited data exist on the biological half-life; however, we have a reported minimum and a reported maximum value. Furthermore, the most likely value is somewhere between these values. The software we used to run the simulations, recommends this type of distribution for these limited data. A lognormal distribution is typically used when the data are skewed toward one end of the distribution. In this case, the assumption of a most likely value of 10 days skews the triangular distribution in a similar manner as a lognormal distribution.

92. Why is the dose and dose rate reduction factor used here? This parameter is normally used to convert high-dose-rate risk to low-dose-rate risk. Is that the intent here?

ATSDR Response: The dose and dose rate effectiveness factor (DDREF) is used to extrapolate from high doses (derived from A-bomb studies) to low dose and dose rates. Typically, the DDREF is applied when the doses are less than 20 rads as recommended by the UNSCEAR 1993 report.

93. Table 3 - "mean" and "95th" doses are reported - but these are the mean and 95th percentile of the maximally exposed person (for which the author's say is an overestimate of dose) … does "mean" and "95th" really describe what is intended? The caption states, "The potential average tritium doses (to the maximally exposed individuals)…" - seems odd to be talking this way. Is the assessment meant to determine the actual mean or most probable dose? or the uncertainty of estimating a mean dose?

ATSDR Response: Several sentences have been added to the introduction to clarify that this evaluation focuses on estimated exposures and the potential for those exposures to produce adverse health effects to those individuals with the highest potential doses. References to the mean and 95th percentile doses have been clarified by indicating that those parameters refer the probability distribution of the estimated doses and that the most frequent doses from the Monte Carlo simulations are less than the mean values.

94. Page 21/22 - how soon after the 1970 release did Meyers take his measurements of urine in exposed individuals? Timing might be the reason for not seeing tritium in the samples.

ATSDR Response: Most urine samples were taken from off-site individuals on the day following the release (one each were taken 2,3, and 4 days following release). About one half of the on-site employee samples were collected on both the day of and the day following the release. Others were collected up to six days following the release. We agree that the timing of the sampling could reduce the potential for detecting any tritium exposures. However, based on our estimated doses, even exposure for 1 day should produce a dose well above the stated detection limit (5,000 pCi/L; or a 0.025 mrem dose).

95. Page 17 (top line) - states that the 95th percentile dose is 0.14 mrem, but then the corresponding dose (or so it seems) in Table 2 shows 0.2 mrem.

ATSDR Response: This was probably due to multiple runs of the Crystal Ball simulations and rounding error in the results. The revised tables and text have been checked and corrected for consistency.

96. Page 19 (bottom) - again, the "most probable" of the distribuiton is NOT the arithmetic average of those distributions - the most probable or "most likely" is at the peak.

ATSDR Response: References to the mean and 95th percentile doses have been clarified by indicating that those parameters refer the probability distribution of the estimated doses and that the most frequent doses are less than the mean values.

97. Page 40 - it is stated that in order to adequately represent the range of all parameters, thousands of iterations are needed … that's not true … Latin hypercube methods solve this problem and have been shown to require only a hundred or so samplings to adequately represent input distributions and describe an output distribution.

ATSDR Response: The phrase "in a Monte Carlo analysis" has been added to that statement.

98. There are errors in the calculation of dose, if the DDREF variable was used. Also, the "weighting factor" is mandated by 10 CFR 20 as unity (1) for beta emitters. That parameter probably shouldn't be varied in an assessment such as this.

ATSDR Response: See response to comment 16 regarding DDREF. The radiation weighting factor of 1 is recommended by the Nuclear Regulatory Commission as well as the ICRP and the NCRP. The expert panel convened by ATSDR recommended a higher value for tritium and that is the value used by ATSDR. As referenced in the Expert Panel Report (ATSDR 2002), there are many references in the literature reporting a tritium radiation weighting factor in excess of 1.0.

99. The technical form of the document is still quite high in places.

ATSDR Response: We agree that this document presents information in a technical form. However, the Livermore community is quite sophisticated with regard to their concerns about radiological exposures and has requested that we do not omit information from our analyses. In addition to distributing the PHA documents, we have made several presentations of this information to the community and have made every effort to address and answer any questions the community may have about this material.

100. Page 13 (1st paragraph, next-to-last sentence) - the phrase "from weather years 1991 and 1993" would seem to belong in parentheses.

ATSDR Response: This section (and sentence) have been re-written in response to newly available weather information.

101. Page 14 (2nd sentence) - according to other statements in the document, the time of the accident was 6:14 am, not 6:30 am.

ATSDR Response: Different reports and news articles have reported slightly different times for the release. Although the specific time of the release may seem unimportant, dispersion of the HT plume is highly dependent on the atmospheric stability classification, which in turn, is dependent on whether the release occurred before or after sunrise. A newly available accident report unequivocally states that the release occurred at 5:49 AM. All references to the time of release have been so corrected.

102. Page 15 (top line) - what is a "flow vector"? It's meaning probably should be explained.

ATSDR Response: This section has been re-written in accordance with newly available 1965 weather data.

103. Page 20 (bottom) - "360 mrem" should be "360 mrem/yr".

ATSDR Response: This change has been made.

104. Page 40 (equation) - the tritium concentration parameter is shown to have units of Bq, but the dimensions of concentration should be something like Bq/L or Bq/m3.

ATSDR Response: See the response to comment 51 (page 54).

105. Page 40 - is the weighting factor "wt" the same as a radiation weight factor (quality factor) for beta emitters? That should be made more clear.

ATSDR Response: Yes, the term "radiation weight factor" has been inserted in the bulleted description. The assumption label is an abbreviation.

106. Page 40 - again, the DDREF is used incorrectly. It should not be in this assessment.

ATSDR Response: See the response to comment 16 (page 62).

107. Page 45 (sensitivity chart) - what parameters do cells D18 and K14 represent?

ATSDR Response: Cell D18 represents an inhalation dose from a light work breathing rate and K14 represents the strenuous work breathing rate. The Crystal Ball report and sensitivity chart have been updated and all of the significant assumptions and forecasts labeled.

108. Many of the Crystal Ball assumptions are overly biased by making them normal when there may not be sufficient data to support such assumptions.

ATSDR Response: As previously stated, the assumption of any probability distribution presents some bias into the evaluation. However, an assumption of normality is the least biased of the probability distributions. For example, a lognormal distribution assumes that most occurrences of a variable will be significantly lower than the mean value. Conversely, a normal distribution assumes those values are equally likely to be lower or higher than the mean. It is our belief that most parameter distributions should be assumed normal, unless there is adequate information to support a non-normal distribution. There is sufficient data to indicate that body masses and breathing rates should be lognormally distributed. These parameters and the associated text and figures have been so changed.

109. Page 47 - vegetable and milk concentrations given in different units. This is confusing.

ATSDR Response: Milk and vegetable concentrations units have been revised to similar units (Ci/L).

110. Page 48 - the inhalation rate distributions are given for adults or children? Or both?

ATSDR Response: The revised Crystal Ball report (Appendix 4) contains both child and adult breathing rates.

111. Page 49 - normal distributions for body mass? It's been shown that lognormal is a better distribution to use for body mass (and other physiological and environmental parameters).

ATSDR Response: The revised dose calculations use lognormal distributions for body weights and breathing rates.

112. In several places it is stated that there is "no public health concern" regarding tritium doses. But, hasn't concern been expressed by the public? Maybe the wording should be altered to state that "ATSDR has no public health concern". Also, the second bullet on page 27 and at the end of the first paragraph on page 28, the author's state that "incidence was elevated" but it was "not statistically significant"? It would seem more reasonable to conclude simply that "there was no statistical increase in incidence", and leave it at that.

ATSDR Response: The phrase "no public health concern" has been replaced with the phrase "below levels of public health concern". Further, "levels of public health concern" are clearly defined by addition of a subsection on "Tritium Doses of Public Health Concern" to the Public Health Implications Section. This subsection clearly identifies the health comparison values and standards that ATSR uses in making its health determinations. The sentences describing the expected or elevated incidences of disease frequency accurately convey the information that those rates were higher than expected, but that the increase was not significant. Some members of the community might find it misleading if we did not state that there was a higher than expected incidence rate.

Reviewer 3 Comments

113. The public health assessment provides a complete description of the potential pathways of human exposure. The figures are helpful in presenting these pathways.

An additional statement on the significance, or lack of significance, of the HTO inhalation pathway would be helpful. It is stated (page 12) that it has been found that "there is little direct atmospheric conversion of HT to HTO (compared with the conversion rate from microbes)". This is a comparison of atmospheric conversion of HT with the conversion of HT deposited onto the soil, and only a fraction of the atmospheric HT is deposited onto the soil. Thus, this comparison does not provide complete information on the size of the HTO inhalation dose relative to the doses from other pathways considered. It would be helpful to add a statement on the DOSE from HTO inhalation relative to the DOSES from other pathways considered.

ATSDR Response: The RASCAL model estimates a maximum dose of 103 mrem (at 1 mile) based on the assumption that 100% of the tritium plume was present as HTO. Noguchi et al. (1989) and Noguchi (1995) conclude that the conversion of HT to HTO in air will be less than 0.5% over the time period of a few hours (according to the RASCAL model the HT plume has completely dispersed in 30 minutes). Using the RASCAL dose estimates, if 1% of the HT plume was present in an HTO form, it would add about 1 mrem to the total tritium dose at the point of maximum exposure. A paragraph indicating this source of uncertainty will be added to the section on "Total Tritium Doses and Uncertainties in Dose Modeling Procedures."

114. The pathways and models for tritium fate and transport are appropriate.

115. The RASCAL computer model developed by the Nuclear Regulatory Commission and the ISC model developed for the Environmental Protection Agency are very appropriate for the exposure analyses. Also, considering uncertainties in model parameters, it was appropriate to compute probability distributions, and the Crystal Ball model used is a well-established model.

116. The literature on the relationship between human exposure to ionizing radiation (all sources) and the potential health effects associated with that exposure is vast. The public health assessment presents data from only a small fraction of these references, but the information presented is adequate to allow conclusions and recommendations to be formulated. Including material from the report of the United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR) was particularly appropriate since the work reported was based on a review of many studies. Similar reviews of many studies have been conducted by the National Academy of Sciences' Committee on the Biological Effects of Ionizing Radiation (BEIR). The report of this work would also be an appropriate reference.

117. The public health assessment accurately and clearly communicates the absence of a health threat posed by the accidental tritium releases from this site. I think that the Summary of the report is excellent and very clearly presents the results of this assessment.

118. The conclusions and recommendations are appropriate in view of the potential doses as described in the public health assessment.

Many conservative assumptions were made in this assessment, which will result in an overestimation of the calculated doses. Even with these conservative assumptions, the calculated doses are small. Furthermore, it is noted (page 21) that following the 1970 accident, analyses were made of urine samples from potentially exposed people, including both LLNL workers and potentially affected community workers. The detection limit for these analyses corresponded to a very low dose of 0.025 mrem, and no elevated tritium body burdens were detected. This is quite significant, and it is appropriately noted in the report that these data indicate that the modeled exposures over-estimate the actual human exposures.

The modeled exposures and the actual data collected following the 1970 accident provide strong support for the conclusion that the accidental HT releases are "no apparent public health hazard."

ATSDR Response: The above comments are noted, no responses are necessary.

119. In the first paragraph on page 10 it is stated that the hypothetical maximally exposed individual for the 1965 accident is assumed to be 0.5 miles from the LLNL tritium facility. In the Summary it is stated that this distance is about 1 mile. Also, in Appendix A it is stated those distances less than 0.7 miles are on site. Thus, the stated distance on page 10 needs to be clarified.

ATSDR Response: The revised dispersion calculations and dose estimates use newly available weather data for the 1965 release. Dispersion calculations for both the 1965 and 1970 releases indicate that maximum HT and HTO air concentrations occurred at a distance of 1 mile from the release source. All references to the locations of maximum exposure have been revised.

120. In the fourth paragraph on page 11 it is stated that soil HT concentrations are represented by probability distributions. However, in Table 1 on page 15 soil HTO emission rates are presented as point values. The deposition rate used in the computation of these values of the soil HTO emission rates should be explained.

ATSDR Response: The ISC model can only use point values for HTO emission rates. The revised calculations accommodate this limitation by running the ISC model iteratively using the mean and 95th% HTO emission rates. The resulting mean and 95th% HTO air concentrations are used to develop a lognormal HTO air concentration probability distribution which is used in subsequent dose calculations. Descriptions of this procedure (and the HT deposition velocity assumptions) have been added to the document.

121. In the second paragraph on page 15 there is reference to computing soil HTO emission rates using "the cumulative soil HT loading values from Table A-2". I think that this should refer to the "cumulative HT air concentrations" rather than to the "cumulative soil HT loading values". This issue occurs in other places within the report also, including the caption of Table 1 on page 15.

ATSDR Response: These suggested changes have been made in the revised document.

122. Many conservative assumptions are made in this assessment, which will result in an overestimation of the calculated doses. This fact is appropriately noted in the report. Assumptions in the ingestion exposure pathway are especially conservative. These include the assumption that all fruits and vegetables eaten were grown at the point of maximum exposure. The probability distribution for tritium in vegetation, which was based on actual measurements along the centerline of the 1970 plume, was censored so that no values less than the detection limit were used. I am not aware of a technical justification for censoring the data. Tritium concentrations in milk were also based on actual measurements. Only three out of thirteen milk samples had detectable tritium concentrations, and only these three measurements were used for the probability distribution. This method of analysis produces a positive bias. Even with these overly conservative assumptions, the doses from the ingestion pathway are negligible. Thus, these assumptions did not effect the conclusions in this assessment. It is recommended that this degree of conservatism not be a precedent for all assessments.

ATSDR Response: The revised vegetation tritium concentration parameter is not censored at the detection limit. Also, the section on "Total Tritium Doses and Uncertainties in Dose Modeling Procedures" includes a paragraph on uncertainties associated with the ingestion dose component.

123. In the fourth paragraph on page 23, one sentence states that "the dose-response relationship appears to have a threshold at about 25 rads" and another sentence in this paragraph states that " these data suggest a linear dose-response curve without a threshold". The same reference is given for both statements. Are these different conclusions from different aspects of the same study? It would be helpful to add some statement clarifying these statements or just acknowledging the differences.

ATSDR Response: Although both health effects relate to central nervous system development, the referenced study concludes that the collective data may support either a threshold or no threshold dose response. More important, is that, overall, these health effects occurred at doses that are 500 to a 1000 times higher than the doses estimated from the LLNL tritium releases. A statement clarifying the uncertainty of the dose response has been added.

124. Table A-4 appears to be incomplete - at least, it does not contain all of the data presented in Table A-5.

ATSDR Response: Table A-4 (re-numbered as A-1) has been modified to include all of the output contained in Table A-5 (A-2).

125. In Appendix 3 on page 40 an equation used in the Monte Carlo simulation is presented. The equation contains a term for the "tritium concentration in curies". The location of the tritium concentration should be stated. Is it in the body? Also, concentrations have units of activity per unit volume or mass, which does not seem to work in this equation. Is the term actually an "activity" rather than a "concentration"? This should be clarified.

ATSDR Response: The term "tritium concentration" refers to the total amount of tritium in the body. The equation as it is expressed is calculating the amount of energy deposited inside the body. This is shown in the equation by the "wt factor/body mass" term. In the equation, the tritium concentration is dependent on the total amount of tritium in the body and the water content of the body to yield the concentration in picocuries/liter. However, we are calculating to dose to the entire body (hence the body mass term) so only the total amount of tritium is important. For the purpose of calculating the radiation dose to the entire body, the total amount of tritium in the body is used. Typically, this equation is used after one has calculated the intake of a substance using concentration factors, intake amounts, and time frame of intake to give the total intake.

126. The questions asked of peer reviewers are appropriate ones, and the opportunity is provided to submit additional comments, which is good.

18 Note that all references for the appendices are listed in the preceding References section.

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