March 26, 2010 

NOTE TO: PETER PRUESS

Thank you for the opportunity to participate in the interagency science consultation on 

the response to the NRC’s evaluation of EPA’s dioxin reassessment. The document was 

reviewed by several components of the Department of Health and Human Services, 

including the Food and Drug Administration, the National Center for Environmental 

Health/Agency for Toxic Substances Disease Registry, the National Institute of 

Environmental Health Sciences/National Toxicology Program, the National Institute for 

Occupational Safety and Health, and the Office of Public Health and Science. Their 

comments have been consolidated and are attached. Let me know if you have any 

questions about them. I look forward to hearing about next steps in the reassessment of 

dioxin. 

Sandra Howard 

Senior Policy Analyst 

Office of Science and Data Policy 

U.S. Department of Health 

 and Human Services 

 

 
1 

3/26/2010 

HHS Comments on 

“EPA’s Response to ‘Health Risks from Dioxin and Related Compounds: 

Evaluation of the EPA Reassessment’ Published by the National Research Council 

of the National Academies” 

BACKGROUND 

Since 1991, the United States Environmental Protection Agency (EPA) has been working 

to complete a scientific reassessment of the health risks of exposure to Dioxin-Like 

Compounds (DLCs). In 2004, EPA sent a 2003 draft document titled “Exposure and 

Human Health Reassessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and 

Related Compounds” to the National Academy of Sciences (NAS) for review. 

In 2006, the NAS published a report titled “Health Risks from Dioxin and Related 

Compounds: Evaluation of the EPA Reassessment.” The NAS report recommended 

improvement in the quantitative approaches used by EPA to characterize risks. Three 

areas that were specifically identified are: 

(1) Justification of approaches to dose-response modeling for cancer and noncancer end points, 

(2) Transparency and clarity in selection of key data sets for analysis, and 

(3) Transparency, thoroughness, and clarity in quantitative uncertainty analysis. 

In response to these three areas of concern, EPA drafted a document titled “EPA’s 

Response to ‘Health Risks from Dioxin and Related Compounds: Evaluation of the EPA 

Reassessment’ Published by the National Research Council of the National Academies.” 

Following are the compiled comments from five HHS components on this document, 

hereafter referred to as “the EPA Response” or “the Response”. Participating agencies or 

offices include the Food and Drug Administration (FDA), the National Institute of 

Environmental Health Sciences/National Toxicology Program (NIEHS/NTP), the 

National Institute for Occupational Safety and Health (NIOSH), the National Center for 

Environmental Health/Agency for Toxic Substances Disease Registry (NCEH/ATSDR) 

and the Office of Public Health and Science (OPHS). 

COMMENTS 

Overarching 

Overall, EPA’s Response is well written, comprehensive in the points it addressed and 

scientifically robust. The EPA document provides an understandable review of the 

epidemiological and toxicological literature concerning the potential health hazards 
2 

associated with DLCs. Likewise, the Science Plan for developing the Response is well 

designed and articulated. 

In their review of the 2003 Reassessment, the NAS also provided a number of 

recommendations to address key concerns. Each of those recommendations is included 

on Page 7 of the NAS review. In general, the EPA Response to the key NAS Findings 

considered the recommendations of the NAS. 

It would be helpful if there were a better linkage between the previous analyses from the 

2003 document to the present Response to NAS. Such linkage is lacking particularly for 

cancer slope factors between the two assessments. For example, in the 2003 document 

there were several hypothesized precursor events for cancer. These endpoints are 

discussed in Appendix H as cancer precursors, but are not used in either dose response 

modeling for cancer or for derivation of an RfD. The reason for this exclusion should be 

better described. In addition, it is not clear whether this is an extension of the dose 

response chapter or an entirely different document. This should be clarified. 

Transparency/Study Selection 

The Ranch Hand cohort results were excluded from the EPA qualitative cancer 

comparisons due to inability to distinguish between exposures to 2,4,5-T, and its 

contaminant TCDD, and exposure to 2,4-D, the other herbicide component of Agent 

Orange, as stated in chapter 2 (2-65). Inability to control for exposure to other chemicals 

including 2,4-D is a problem of a majority of occupational cohorts included this 

assessment, most apparently in the NIOSH cohort that included workers from multiple 

plants. However, TCDD is the main concern, and evidence of carcinogenicity of 2,4-D in 

humans is sparse. Production data from the Dutch cohorts (Boers et al., 2009) and 

Hamburg cohorts (Flesch-Janys et al., 1998) illustrate the range of potential exposures 

among chemical workers. The lack of dose-response, and elevation in “all cancer” 

mortality in Ranch Hand (Akhtar et al., 2004), should not be a reason for the exclusion of 

results generated by this cohort from qualitative cancer assessment. 

 

The reasons some studies were excluded for the development of an RfD and cancer slope 

factor are not clear. Some studies were excluded that appear to be consistent with the 

inclusion criteria. For example, EPA did not include DEN initiated models or the 

transgenic models (Wyde et al). These are 26-week bioassays that would provide useful 

information. 

Why was a meta analysis not attempted – for example for CVD or diabetes – by 

combining some of the industrial cohorts this could be accomplished (see Humblet et al 

2009)? 

Endpoint Exclusion Criteria 

The endpoint exclusion criteria are poorly articulated. In some accepted studies not all 

endpoints made it into the candidate RfD array and it is not clear why. The relevant 
3 

section (4.2.1) is not clear about the distinction between relevant and non-relevant 

endpoints. 

Endpoint Selection in Bioassay 

There is an inconsistency in how EPA picked which tumor data to model from the NTP 

reports. EPA does not use the NTP carcinogenic calls consistently. From the 1982 NTP 

report, EPA uses statistical significance in choosing the tumor types to model which 

results in more cancer types than were called in the NTP report. In contrast, for the 2006 

NTP report, they do not use uterus, even though, it contributed to the evidence for 

carcinogenicity and was a tumor site that was statistically significant. 

EPA could better describe why they did not use the poly 3 adjustment for early deaths 

when analyzing the NTP data. 

Human Epidemiologic Studies 

The description of the epidemiologic studies considered in this evaluation is very 

thorough and accurate. This level of description is important to ensure transparency in 

justifying the final studies selected for risk assessment modeling. 

The preferred risk assessment model is derived from epidemiologic data, namely, a 

NIOSH cohort of dioxin-exposed U.S. workers. This seems quite appropriate, given the 

comprehensive nature of the cohort (e.g., the wide range of exposures and large number 

of plants contributing information) and the attention given to dose modeling in this 

cohort. The cancer slope factor is based on the Cheng et al. (2006) reanalysis of doseresponse modeling conducted by Steenland et al. (2001). EPA used the all-cancer model 

lagged 15 years, and removed the highest 5% of exposed persons as a means of 

minimizing the problem with cancer risk attenuation at high doses. This appears to be an 

acceptable way of addressing this problem, which is common in occupational 

epidemiology studies. 

Notably, however, there are other approaches that could have been used. For example, a 

piecewise loglinear model (lagged 15 years) might provide a better fit to the data than 

deleting the high-dose individuals. Using the estimated risk at the lower ranges of dose 

(which are most relevant for risk assessment) is straightforward with piecewise models. 

This is a minor issue, however, as the choice of the most appropriate dose-response 

model to employ must consider a wide variety of factors in addition to model fit, and 

EPA may not have had the data necessary to identify the best-fitting model. 

Dose-Response 

One key area in which the NAS requested a response from EPA was justification of 

approaches to dose-response modeling for cancer and noncancer endpoints. The 

following comments focus on that aspect of the EPA dioxin risk assessment, particularly 

in regard to pharmacokinetic modeling for the NIOSH dioxin cohort. 
4 

As noted in the EPA Response to NAS, an earlier analysis of the NIOSH dioxin cohort 

(Steenland et al, 2001) relied on a fixed half-life model of TCDD kinetics. Although 

such a model was consistent with the observed kinetics of humans exposed to low levels 

of TCDD, more recent data from Seveso indicate that a fixed half-life model is not 

consistent with the observations of TCDD kinetics in highly-exposed individuals. The 

updated EPA risk assessment for dioxin makes use of an analysis by Cheng et al (2006), 

which used a non-linear model of TCDD kinetics (Aylward et al, 2005a) to estimate the 

internal doses of the subjects of the NIOSH cohort. The Aylward et al (2005a) model 

includes concentration-dependent elimination kinetics, which the EPA characterizes as 

consistent with our current physiological understanding of TCDD kinetics. EPA’s 

decision to use a concentration-dependent model of TCDD kinetics, as a fixed half-life 

model is incompatible with the observed kinetics of TCDD in highly-exposed 

individuals. 

Throughout most of the EPA Response, TCDD kinetics are modeled using the 

physiologically-based pharmacokinetic model of Emond et al (2006). From the 

standpoint of internal consistency, it would be desirable to use the Emond model rather 

than the Aylward model for internal dose estimation for the NIOSH dioxin cohort. 

However, we are unaware of any such analysis having been published, to date. 

Choice to Ignore TEQ in Dose Response Modeling 

EPA chose not to include background TEQ in the dose response modeling of either the 

human or experimental animal data. EPA pointed out the uncertainties of the use of 

TEQs in these efforts; however, we feel that analyzing the data with TEQ will provide a 

quantitative estimate of this uncertainty. In addition, it will also provide insight into 

whether the exclusion of the TEQ in the dose response analysis is a conservative 

assumption. It is thought that if EPA used the TEQ analysis from the NTP cancer studies 

in rats, this would support the robustness of their cancer slope factors and the use of the 

TEQ methodology. We would support dose response modeling of mixture studies that 

include TCDD and other dioxins. Dose response analysis should not be done on mixtures 

containing the mono-ortho PCBs and chemicals not part of the WHO TEF methodology. 

PBPK Modeling 

The Response presents a comprehensive and voluminous review of peer-reviewed 

literature to support the pharmacokinetic/toxicokinetic dose-response modeling 

approaches it has taken to estimate body burden from 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD) in human (observational) and non-human (experimental) studies. 

EPA agreed with the recommendation of the NAS committee to use biologically-based 

pharmacokinetic models instead of using a simple one-compartment kinetic model to 

estimate past exposure to Dioxin-Like Compounds (DLCs) on the basis of more recent 

serum levels. Specifically, the rationale and scientific justification for the selection and 

application of the Emond human physiologically-based pharmacokinetic (PBPK) model 

developed by Emond et al. (2004, 2005, 2006) appears justified. EPA assumed that the 
5 

same blood TCDD levels leading to effects in animals would also lead to effects in 

humans; this seems like a reasonable assumption. 

The use of pharmacokinetic models in the development of the RfD and cancer slope 

factors is a clear advance and is supported by the data indicating large differences in the 

PK between humans and animals and the clear non-linearity in the pharmacokinetics. The 

uncertainty in the use of these models is well described qualitatively. There are some 

minor suggestions on the use of these models. 

The comparison between the Emond PBPK model and the CADM model are done at 

dose levels that would either kill the animals or are at least a 1000 times higher than 

human background. Model comparisons for humans should be at dose levels within the 

exposure range of interest (0.1-100 pg/kg/d) and the animal dose levels should be at 0.1-

100 ng/kg/d. 

Using the CADM model to estimate cancer risk from the human epidemiology studies is 

a significant advance from previous analyses of this data, it would be helpful to examine 

the same data using the Emond model to better understand the uncertainty in the use of 

the CADM model. 

Non-Cancer Effects 

Per the advice of the NAS, EPA calculated a reference dose (RfD) for TCDD. A 

reference dose is an estimate (with uncertainty spanning about an order of magnitude) of 

a daily oral exposure to the human population (including sensitive subgroups) that is 

likely to be without an appreciable risk of harmful effects during a lifetime. The decision 

to use this approach seems reasonable and as explained below is aligned with how DLCs 

are managed in foods internationally. 

The EPA document provides an understandable basis for EPA’s determination of hazard 

identification and dose-response analysis. EPA’s selection of human studies (Baccarelli 

et al. 2008, Mocarelli et al. 2008) for derivation of the draft TCDD Reference Dose (RfD) 

appears to be appropriate, as these studies represent exposures over different life stages 

including gestation, childhood, and young adulthood. Hence, use of these data to assess 

the potential risk from exposure to TCDD and other dioxin-like compounds would be 

relevant to vulnerable members of the population, such as children, pregnant and nursing 

women, and fetuses. Moreover, use of epidemiology studies decreases the uncertainties 

that are associated with extrapolating toxicity from animals to humans. 

The Baccarelli and Mocarelli studies evaluated health endpoints from people in the 

Seveso cohort, whose members were exposed in 1976 to high peak concentrations of 

TCDD, a potent DLC with a TEF of 1.0, as a consequence of an industrial accident in 

Seveso, Italy. This incident resulted in the highest known exposure TCDD in residential 

populations. While this approach complicated the estimation of average daily doses 

associated with specific endpoints, EPA was able to calculate candidate points of 

departure (PODs) for derivation of an RfD from each of the studies. Given that the 
6 

source of the exposure to TCDD was an industrial accident, it is possible that other DLCs 

also were present. Also, it is possible that non-DLC exposures could have resulted from 

exposure to emissions from the accident; this is possibly a source of uncertainty in these 

studies. One limitation to this approach is that, in human populations it is not possible to 

identify a No Adverse Effect Level (or NOAEL) since the reference groups in human 

studies have background exposures to DLCs. A second limitation is that both of these 

study populations are small and fairly homogeneous. EPA addressed these latter two 

uncertainties via application of “uncertainty factors (UFs)”, a UF of 10 to account for the 

lack of a NOAEL and a UF of 3 to account for human interindividual variability that 

might be unobserved in a small study, for a total composite UF of 30. 

With regard to the Mocarelli study, men who had been exposed to elevated TCDD levels 

when they were children between the ages of one and nine had reduced semen quality 22 

years later. From this study, it is difficult to be certain about the relevant time interval 

over which TCDD dose should be considered. Investigators measured TCDD lipidadjusted serum concentrations within approximately one year of the initial exposure 

event. Because effects were only observed in men who were under 10 years of age at the 

time of exposure, EPA has assumed a maximum 10-year critical exposure window for 

elicitation of these effects. However, because dioxins are stored in fat tissue and slowly 

released over many years, it is difficult to be certain whether the observed effects are a 

consequence of the initial high exposure between one and nine years of age or a function 

of the cumulative exposure that would have occurred during the entire 22 years between 

exposure and semen studies, or something in between. The differences between these 

two dose estimates (the initial high exposure versus the cumulative exposure for the 9 

year window) is within an order of magnitude, which is not considered to be a large 

difference in this context. The RfD derived from the Mocarelli study was 0.7 pg/kg/day. 

For the Baccarelli study, on the other hand, both the TCDD and the endpoint (thyrotropin 

or thyroid-stimulating hormone) levels were measured in neonatal plasma so that there is 

not uncertainty about the relevant exposure window. Despite these differences in 

exposure measurement and endpoint, the RfD derived from the Baccarelli study was 

similar to the one for the Mocarelli study, 0.8 pg/kg/day. Rather than present a single 

RfD EPA could have presented the RfD as a range between these two very close 

estimates. Of note is that RfDs that would have been derived from animal studies are 

higher or lower and mostly within an order of magnitude of these RfDs, so that the data 

are fairly consistent. 

Although different approaches have been taken, EPA’s Draft Reference Dose (RfD) for 

TCDD (0.7 pg/kg-day) is slightly lower than, but similar to, the WHO Tolerable Daily 

Intake (TDI) (1 – 4 pg/kg-day), the ATSDR Maximum Residue Limit (MRL) (1 pg/kgday) and the JECFA Provisional Tolerable Monthly Intake (PTMI) (~2.3 pg/kg-day). 

Although these dose guidelines have different names, the methods for computing them 

and their interpretations are very similar. Thus, despite the uncertainties that are inherent 

in trying to draw conclusions from these kinds of data, and that the EPA assessment is 

more recent, these estimates are close together. 
7 

Carcinogenicity 

The weight of evidence for TCDD carcinogenic hazard potential is objectively analyzed 

in the EPA document. For the cancer dose response assessment, EPA derived an Oral 

Slope Factor (OSF). This approach seems to be appropriate and reasonable. EPA 

selected Cheng et al. (2006) as the study from which to derive the OSF and the methods 

used also appear to be appropriate. Considering the possibility that key events in the 

modes of action (MOA) for TCDD may interact to affect multiple pathways leading to 

carcinogenesis, EPA’s consideration of nonlinear as well as a linear cancer dose-response 

models is reasonable. 

Mode of Action for Cancer 

In figures 5.4 and 5.5 EPA describes an MOA for some cancers. As presented, it can be 

implied from the figures that the adenomas and carcinomas are due to cytotoxicity, which 

is unlikely to be a key event in the hepatocarcinogenic response. This figure also 

excludes cholangiocarcinoma, which were the main hepatocarcinogenic event. 

Recommendation: Since we do not know MOA beyond Ah receptor activation, it may be 

better to exclude the proposed MOAs in this section. 

Cancer Slope Factor 

Agreement in RfDs from 2 human studies is encouraging, as is fact that the OSFs from 

animal [3.2 × 105 to 9.4 × 106 (per mg/kg-day)] and human data [3.75 × 105 to 2.5 × 106 

(per mg/kg-day).] fall within a small range. If anything, the humans may be MORE 

sensitive than the animals. 

It would be helpful to present the distribution of cancer slope factors in a figure similar to 

the Candidate RfD array. 

While EPA considered the Seveso cancer study of Warner et al (2002) to be less relevant 

because it was associated with a single high dose exposure, this is the only cohort in 

which the peak TCDD concentrations is actually measured. The occupational cohorts 

require assumptions about the peak concentration. 

The cited NTP 1982 report says it is from adenomas and carcinomas but they only use the 

carcinoma data (Table 5-8). 

The use of the Kociba study needs clarification. For example, the Kociba data analyzed 

here appears to have fewer animals than the published Kociba study. 

Quality Assurance 

Is there an independent QA plan to cross check the data used in the dose response 

analysis compared to the paper to evaluate the data for transcription errors, as well as 

taking info from the appendix into the full chapters? 
8 

Specific Comments 

Pages 2-1: “chloracne is considered a clinical sign of exposure to dioxin.” 

NOTE: chloracne indicates HIGH dose exposure. 

Page 2-196 lines 27: “As discussed above, female Sprague-Dawley rats (81 control; 82 

treatment group)” 

NOTE: 30 out of the 80 were scheduled sacrifices during the study so that 50 were dosed 

for the entire study. 

Pages 3-5 lines 25-26: “Recent efforts of pharmacokinetic modeling have supported 

the concentration dependent elimination of TCDD in animals and humans (Aylward et 

al., 2005b; Emond et al., 2006).” 

NOTE: The Andersen models always had inducible metabolism starting in about 1991. 

Pages 3-9 lines 19-25: “The redistribution of TCDD tissue levels from liver to fat with 

increasing time suggests that binding of the chemical in the liver (including via 

induction of CYP1A2) is an important kinetic consideration at early exposure points 

with relatively high applied doses. At steady state levels (longer than 35 days, and low 

applied doses), there seems to be a tendency for TCDD to redistribute to fat tissue.” 

NOTE: The redistribution from liver to fat is not a steady-state vs. time issue. It is really 

related to dose and CYP1A2 induction. Following a single exposure, CYP1A2 takes 1-3 

days to peak in the liver, at this time concentrations in liver and fat are dynamic and once 

the peak cYP1A2 is attained, the peak liver:fat ratio is attained. Following a single dose, 

overtime, tissue concentrations decrease. As liver concentrations decrease so does 

CYP1A2 induction and liver:fat TCDD concentrations. Thus, high steady-state 

exposures result in high liver:fat ratios and low steady-state exposure will result in lower 

liver:fat ratios. 

Pages 3-15 lines 1-6: “Bell et al. (2007b) reported that the disposition of TCDD into 

the fetus shows dose dependency, with a greater proportion of the dose reaching the 

fetus at lower doses of TCDD. Further, both CYP1A1 and CYP1A2 are highly 

inducible (~103−fold) in fetal liver, whereas CYP1A2 shows much lower induction 

(10−fold) in maternal liver. It has been speculated that this is due to the lower basal 

levels of CYP1A2 in fetal liver, as compared to maternal liver (Bell et al., 2007b).” 

NOTE: The greater relative disposition to the fetus at low doses may be that less is 

bioavailable from the dam at high doses due to greater hepatic sequestration in mom’s 

liver and greater elimination. 

Table 4-4: These are the PODs and should be mentioned in this table 

Page 429 text box: “NTP (2006), however, found virtually no TCDD in the tissues of 

untreated animals or in the feed stock. In all of these studies, except the 28-day 

exposure in Bell et al. (2007a), control animals were gavaged with corn oil vehicle. 

TCDD concentrations in corn oil were not reported in any of the studies. 
9 

NOTE: Note the lack of TCDD in control animals was due to incorrect method- with too 

high a LOD. In the TEF mixture, we could detect TCDD in controls-we saw 13 pg/g fat 

at 2 yrs and 8 pg/g in liver. The bigger issue here is background TEQ total 

Figure 6-3: States it is from Hattis 2009. EPA should check this citation. It does not 

appear correct. 

Page 20 R-9 Line 5: Wrong Reference 

Correct Reference: 

NTP TECHNICAL REPORT ON THE TOXICOLOGY AND CARCINOGENESIS 

STUDIES OF 2,3,7,8-TETRACHLORODIBENZO-p-DIOXIN (TCDD) (CAS NO. 

1746-01-6) IN FEMALE HARLAN SPRAGUE-DAWLEY RATS (GAVAGE 

STUDIES) NATIONAL TOXICOLOGY PROGRAM April 2006 NTP TR 521 NIH 

Publication No. 06-4468