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2,3,7,8-TCDD (Dioxin) Risk Assessment and Cancer Risk

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Quantitative Human Health Risk Assessment Methodology for 2,3,7,8-TCDD (Dioxin)

2,3,7,8-TCDD (Dioxin) — CASRN 1746-01-6

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RoseyFields conducts 2,3,7,8-TCDD (Dioxin) risk assessments using the following methodology. This document sets out the toxicological basis, quantitative derivations, and application protocols that RoseyFields’ risk assessors rely upon when TCDD is identified as a chemical of potential concern in a site investigation, exposure reconstruction, or litigation support matter. It is intended to be read alongside the site-specific data evaluation and exposure assessment work products prepared for each engagement, and it will be updated as the underlying peer-reviewed literature and regulatory toxicity values evolve.

1. Chemical Identity, Regulatory Background, and Exposure Sources

1.1 Chemical Identity

2,3,7,8-Tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) is the prototype and most extensively studied member of a family of 75 structurally related chlorinated dibenzo-p-dioxin (CDD) congeners. The CDD family is subdivided into eight homologue groups according to the number of chlorine substituents (mono- through octa-chlorinated), with the position and number of chlorine atoms determining relative toxicity. 2,3,7,8-TCDD carries four chlorine atoms at the 2, 3, 7, and 8 ring positions, the substitution pattern associated with the greatest toxic potency among the CDDs, and it therefore serves as the reference congener against which the toxicity of other “dioxin-like” compounds is expressed using Toxicity Equivalence Factors (TEFs). The chemical formula is C12H4Cl4O2, molecular weight 322 g/mol. TCDD is a colorless, odorless solid with very low vapor pressure (7.4 × 10-10 mm Hg at 25°C) and a high octanol-water partition coefficient (log Kow 6.8–7.58), properties that govern its strong affinity for soil organic carbon, sediment, and lipid-rich biological tissue and its resistance to volatilization, hydrolysis, and biodegradation.

1.2 Regulatory Background

TCDD is not intentionally manufactured for commercial use; its only current legitimate use is as a research standard. It occurs instead as an unintentional trace by-product of combustion and of certain chlorine-based industrial chemistry, and it is regulated as a hazardous waste and priority pollutant rather than as a marketed product. EPA’s Integrated Risk Information System (IRIS) database provides the current chronic oral Reference Dose (RfD) for TCDD, finalized February 17, 2012; no inhalation Reference Concentration (RfC) has been derived. EPA has separately classified TCDD as a probable human carcinogen (Group B2) based on both animal bioassay and epidemiologic evidence, and the Department of Health and Human Services (DHHS) and the International Agency for Research on Cancer (IARC) have similarly concluded that the evidence is sufficient to regard TCDD as carcinogenic. TCDD has been identified at a substantial number of National Priorities List (NPL) sites, and FDA, EPA, and state agencies maintain fish-consumption and drinking-water guidance values discussed in Sections 6–8 below.

1.3 Sources and Exposure Pathways

TCDD is generated predominantly by incomplete combustion. Current environmental releases are dominated by combustion of fossil fuels and wood, and by municipal, medical, and hazardous-waste incineration; historical releases were also associated with the manufacture of 2,4,5-trichlorophenol and its derivatives (including the herbicide 2,4,5-T, a component of Agent Orange), chlorine bleaching in pulp and paper manufacture, and the production or use of certain chlorinated phenols and phenoxy herbicides. Because TCDD binds strongly to organic carbon and bioaccumulates and biomagnifies through the food chain, more than 90% of background dietary intake in the general population comes from food — principally meat, dairy products, and fish — with comparatively minor contributions from inhalation of ambient air (elevated near incinerators and high-traffic areas) and from dermal or incidental-ingestion contact with contaminated soil. Municipal drinking water is not typically a significant pathway because TCDD’s low water solubility causes it to partition to particulates that are removed by conventional filtration; unfiltered groundwater at contaminated sites is the exception. Occupational and accidental exposures — historically documented in phenoxy-herbicide and chlorophenol production workers, Vietnam-era herbicide applicators, and populations affected by industrial releases such as the 1976 Seveso, Italy accident and the Times Beach, Missouri contamination — have produced the serum- and adipose-tissue dose data that anchor most of the quantitative toxicity values discussed below. Because TCDD is eliminated from the human body slowly (elimination half-life on the order of 7–12 years), body burden and serum lipid concentration are used as integrated biomarkers of both past and ongoing exposure in dose reconstruction.

2. Point-of-Departure (POD) Selection

RoseyFields’ noncancer POD selection for TCDD follows the current EPA IRIS assessment (finalized 2012), which departs from the traditional whole-animal-bioassay approach and instead derives the POD from two co-principal human epidemiologic studies of the Seveso, Italy population exposed following the 1976 industrial accident. This approach was adopted because TCDD has an unusually rich human dose-response database — a rarity among IRIS assessments — allowing the RfD to be anchored directly in human data rather than requiring interspecies extrapolation.

2.1 Co-Principal Studies and Endpoints

  • Mocarelli et al. (2008) — reduced sperm concentration and motile sperm count in men who were 1–9 years of age at the time of the Seveso accident, identifying early childhood as a critical window of susceptibility for TCDD-induced reproductive toxicity. The lowest exposed quartile (median lipid-adjusted serum concentration [LASC] of 68 parts per trillion, ppt) was designated the Lowest-Observed-Adverse-Effect-Level (LOAEL), corresponding to a pharmacokinetically modeled adjusted intake of 0.020 ng/kg-day.
  • Baccarelli et al. (2008) — increased thyroid-stimulating hormone (TSH) in neonates born to Seveso mothers, using a maternal lipid-adjusted serum concentration at delivery of 235 ppt as the LOAEL, corresponding to the same PBPK-modeled adjusted intake of 0.020 ng/kg-day. The 5 µU/mL neonatal TSH benchmark used to define the effect level derives from World Health Organization iodine-deficiency surveillance criteria (WHO/UNICEF/ICCIDD, 1994), and the toxicological rationale for its selection as an adverse endpoint reflects the established sensitivity of the developing brain to thyroid hormone disruption (Glinoer and Delange, 2000; Zoeller and Rovet, 2004; Oerbeck et al., 2005).

These two studies were selected as co-critical because they converged, independently, on an identical adjusted LOAEL (0.020 ng/kg-day) for two biologically distinct, sensitive endpoints — male reproductive/endocrine development and neonatal thyroid function — despite differing exposure windows (childhood versus in-utero/perinatal) and differing modeling approaches. This convergence was treated as strengthening confidence in the POD. A Benchmark Dose (BMD) approach was considered but the underlying data were not amenable to BMD modeling, so a LOAEL/NOAEL approach with a corresponding uncertainty factor was used instead.

2.2 Supporting Studies

A broader set of human and animal studies was reviewed to confirm that immune, endocrine, and developmental effects represent the most sensitive toxicological domains for TCDD, consistent with the selected co-principal endpoints:

  • Mocarelli et al. (2011) reported comparable reductions in semen quality among men exposed to TCDD perinatally (in utero and via lactation), corroborating the childhood-exposure findings of the principal study.
  • Rodent reproductive-toxicity data (Simanainen et al., 2004; Latchoumycandane and Mathur, 2002; Ikeda et al., 2005; Ohsako et al., 2001) demonstrate qualitatively similar male reproductive and gonadal effects across species, supporting the biological plausibility of the human findings.
  • Animal thyroid-hormone studies (Seo et al., 1995; Sewall et al., 1995; Van Birgelen et al., 1995a; Van Birgelen et al., 1995b; Crofton et al., 2005) independently establish thyroid hormone disruption as a sensitive, dose-dependent TCDD effect, corroborating the Baccarelli neonatal TSH endpoint.
  • Bell et al. (2007) (developmental delay in the Wistar(Han) rat) and NTP (2006) (liver and lung lesions in female Sprague-Dawley rats, the most comprehensive chronic rodent bioassay of TCDD to date) generated candidate animal-based RfDs, several of which were lower than the human-data-based RfD; these were retained as supporting/comparative values rather than as the principal POD, consistent with IRIS’s preference for adequate human data over interspecies extrapolation.
  • Additional human endpoints reviewed but not used as the principal POD because dose-response could not be adequately quantified include immune-marker changes (Baccarelli et al., 2002; Baccarelli et al., 2004), developmental dental aberrations (Alaluusua et al., 2004), female reproductive endpoints such as endometriosis, age at menarche, age at menopause, uterine leiomyoma, and ovarian function (Eskenazi et al., 2002; Warner et al., 2004; Eskenazi et al., 2005; Eskenazi et al., 2007; Warner et al., 2007), and diabetes (Michalek and Pavuk, 2008).

3. Pharmacokinetic Modeling and Dose Conversion

Because the Seveso serum-concentration data reflect a single high initial exposure event followed by gradual first-order elimination over several years — rather than the constant daily intake assumed by a standard chronic RfD — RoseyFields’ methodology relies on the physiologically based pharmacokinetic (PBPK) modeling performed by EPA to convert these atypical exposure profiles into continuous daily oral intake equivalents. The Emond human PBPK model (Emond et al., 2005) was used for both co-principal studies:

  • For Baccarelli et al. (2008), maternal lipid-adjusted serum concentration at delivery (235 ppt) was treated as approximately constant across the 30-year exposure period preceding delivery, and the Emond model was used to back-calculate a corresponding continuous daily oral intake of 0.020 ng/kg-day.
  • For Mocarelli et al. (2008), the atypical time-course (an acute high-dose spike from the 1976 accident followed by slow multi-year elimination) required a more complex treatment. EPA used the PBPK model to estimate both (i) the peak initial intake associated with the accident (0.032 ng/kg-day) and (ii) the average intake across the full ten-year critical window of susceptibility, defined as the first ten years of life (0.008 ng/kg-day). Because of uncertainty about whether the peak or the time-averaged exposure was more toxicologically relevant to the observed sperm effects, the LOAEL[adjusted] of 0.020 ng/kg-day was calculated as the arithmetic average of these two internal-dose estimates.

This PBPK-based approach allows body-burden and serum-lipid data — the exposure metrics actually measured in the underlying human cohorts — to be translated into the external, continuous daily oral intake terms in which the RfD is ultimately expressed, and is the same general dose-reconstruction logic RoseyFields applies when converting site-specific biomonitoring or intake-modeling results into internal-dose equivalents for comparison to the RfD in individual case work (see Section 7).

4. Uncertainty Factor Derivation

A composite uncertainty factor (UF) of 30 was applied to the co-critical LOAEL[adjusted] of 0.020 ng/kg-day (2.0 × 10-8 mg/kg-day) to derive the RfD. The individual UF components, and RoseyFields’ understanding of the basis for each, are as follows:

Uncertainty Factor Value Basis
LOAEL-to-NOAEL 10 Applied because a No-Observed-Adverse-Effect-Level (NOAEL) could not be identified in either co-principal study; both Mocarelli et al. (2008) and Baccarelli et al. (2008) identified their lowest studied exposure group as a LOAEL.
Interindividual (intraspecies) variability 3 A reduced UF of 3 (rather than the default 10) was applied because the principal cohorts already comprised sensitive human lifestages — infants exposed in utero and children exposed under age 10 — rather than average healthy adults, so the toxicodynamic component of interindividual variability is partly addressed by study design. A full UF of 10 was not used because the sample sizes for these lifestages were relatively small and chronic effects in humans are not fully characterized, leaving residual uncertainty about the full range of human sensitivity.
Interspecies extrapolation 1 No interspecies UF was needed because the POD is derived directly from human epidemiologic data, eliminating the animal-to-human extrapolation uncertainty that would otherwise apply.
Subchronic-to-chronic / study duration 1 No study-duration UF was applied. Although chronic effect levels are not precisely defined in humans, animal bioassay data indicate that developmental and other short-term effects occur at exposure levels comparable to those causing effects in chronic studies, so extending the observed effect level to a lifetime chronic exposure was not judged to require an additional discount.
Database completeness 1 No database-deficiency UF was applied because the TCDD toxicological database is unusually extensive, spanning a comprehensive set of endpoints in both human and animal studies of multiple species and both sexes, with no indication that additional studies would identify a more sensitive effect or lower the derived RfD.
Composite UF 30 10 (LOAEL-to-NOAEL) × 3 (interindividual) × 1 × 1 × 1.

5. Derivation of the Chronic Oral Reference Dose (RfD)

Applying the composite UF of 30 to the co-critical LOAEL[adjusted] of 2.0 × 10-8 mg/kg-day yields:

RfD = 2.0 × 10-8 mg/kg-day ÷ 30 = 7 × 10-10 mg/kg-day

This is the chronic oral RfD currently listed on IRIS for TCDD alone. Overall confidence in the RfD is rated high by EPA: confidence in the co-principal studies is high given their well-conducted design and demonstrated human health effects (tempered by exposure-assessment uncertainty in Mocarelli et al. (2008) due to the atypical acute-then-declining exposure profile, mitigated by PBPK modeling of the exposure window); confidence in the supporting database is high given its breadth across species, sexes, and endpoints with consistent findings between animal and human data. No inhalation RfC has been derived for TCDD; RoseyFields therefore does not apply a chronic inhalation noncancer toxicity value for TCDD unless and until EPA (or another responsible regulatory body) issues one, and instead relies on the oral RfD as the operative noncancer benchmark, adjusted by route-to-route extrapolation only where scientifically defensible and clearly disclosed as such. Where a mixture of dioxin-like compounds (DLCs), polychlorinated dibenzofurans (PCDFs), or dioxin-like PCBs is present, RoseyFields converts the mixture to a TCDD-equivalent dose using the consensus mammalian Toxicity Equivalence Factors developed by the World Health Organization (Van den Berg et al., 2006) before comparing the summed TCDD-equivalent dose to the RfD, consistent with current EPA guidance.

6. Carcinogenicity Classification and Cancer Slope Factor

6.1 Weight-of-Evidence Classification

TCDD has been placed in EPA’s Group B2 (“probable human carcinogen”) classification, and DHHS and IARC have similarly concluded that it is reasonable to regard TCDD as carcinogenic to humans. This weight-of-evidence conclusion draws on both animal bioassay data and human epidemiologic data:

  • Animal bioassay evidence: The foundational chronic bioassay is Kociba et al. (1978), a two-year dietary study in Sprague-Dawley rats that reported increased incidence of hepatocellular tumors and other neoplasms and has served as the principal basis for TCDD cancer risk assessment in North America and Western Europe for several decades. The more recent, larger NTP (2006) bioassay in female Harlan Sprague-Dawley rats corroborated an elevated incidence of liver, lung, and oral-mucosal neoplasms using contemporary histopathological criteria, with the NTP study additionally identifying cholangiocarcinoma of the liver not seen in the earlier Dow-sponsored study. TCDD has been shown to act as a complete, multi-site, trans-species carcinogen, producing tumors of the liver, lung, thyroid, and other tissues across multiple rodent strains and species.
  • Human epidemiologic evidence: Occupational cohort mortality studies — most notably the NIOSH cohort of 5,172 U.S. workers at 12 plants producing TCDD-contaminated chemicals (Fingerhut et al., 1991) and the pooled international phenoxy-herbicide/chlorophenol worker cohort coordinated by IARC (Kogevinas et al., 1997) — reported associations between TCDD exposure and increased mortality from soft-tissue sarcoma, non-Hodgkin lymphoma, and all-cancer mortality, particularly among the more highly exposed workers, together with the Seveso and Vietnam-veteran cohorts summarized in Section 1.3. Consistent with the general epidemiologic literature on TCDD, associations are stronger for aggregate/all-cancer endpoints than for any single cancer site, and the human evidence has been characterized as “probable” rather than definitive largely because of co-exposure to other chlorophenoxy chemicals and chlorophenols in most cohorts.

6.2 Quantitative Cancer Slope Factor and Unit Risk

EPA’s Health Assessment Document for Polychlorinated Dibenzo-p-Dioxins derived the following quantitative cancer potency values for TCDD, calculated by application of a linearized multistage low-dose extrapolation model to the Kociba et al. (1978) rat liver tumor incidence data:

  • Oral Cancer Slope Factor (CSF): 1.5 × 105 (mg/kg-day)-1
  • Oral Unit Risk: 4.5 (µg/L)-1
  • Inhalation Cancer Slope Factor: 1.5 × 105 (mg/kg-day)-1
  • Inhalation Unit Risk (IUR): 3.3 × 10-5 (pg/m3)-1

RoseyFields notes for the record that IRIS’s TCDD cancer assessment (Volume 2 of EPA’s Reanalysis of Key Issues Related to Dioxin Toxicity) had not been finalized as of the most recent IRIS status update available to RoseyFields, and that the legacy CSF/IUR values above originate from EPA’s earlier Health Assessment Document rather than from the same PBPK-modeling framework used for the current noncancer RfD. A pathology re-review of the original Kociba et al. (1978) liver slides using updated NTP histopathological criteria (Keenan et al., 1991) found substantially fewer liver tumors than originally reported and generated revised, generally lower, cancer potency estimates; RoseyFields discloses this reanalysis in any matter where the CSF materially drives risk conclusions and, where appropriate to the jurisdiction and regulatory framework governing the matter, presents both the legacy and reanalyzed potency estimates as a bounded range rather than relying on a single point value without qualification. RoseyFields’ default practice is to apply the current EPA/IRIS-endorsed value applicable in the regulatory jurisdiction governing the specific engagement, and to flag prominently in every work product when a cancer potency value has not been updated by EPA to reflect PBPK-based human-data modeling analogous to that used for the RfD.

7. Application by RoseyFields’ Risk Assessors

7.1 Hazard Quotient and Hazard Index

For noncancer evaluation, RoseyFields calculates a chemical-specific Hazard Quotient (HQ) for each relevant exposure pathway as:

HQ = Estimated Daily Intake (or Absorbed/Internal Dose, mg/kg-day) ÷ RfD (7 × 10-10 mg/kg-day)

Because the RfD is expressed in mg/kg-day and derived from lipid-adjusted serum data via PBPK modeling, RoseyFields’ preferred approach where biomonitoring data are available is to compare a claimant’s or population’s measured or reconstructed lipid-adjusted serum TCDD concentration directly to the serum-concentration benchmarks underlying the co-principal studies (68 ppt for the Mocarelli childhood-exposure endpoint; 235 ppt for the Baccarelli perinatal/neonatal endpoint), in addition to — not instead of — the conventional external-dose HQ calculation; this dual approach is more consistent with how the toxicity value was itself derived and reduces the uncertainty introduced by exposure-modeling assumptions when converting external dose back into an internal-dose equivalent. Where multiple TCDD exposure pathways are present (e.g., soil ingestion, dermal contact, dietary intake, inhalation of particulates), individual pathway HQs are summed to a route-specific Hazard Index (HI), and, where dioxin-like congeners, PCDFs, or dioxin-like PCBs co-occur, the TEF-weighted TCDD-equivalent dose (Section 5) is used as the numerator before HQ/HI calculation. An HI or HQ at or below 1.0 is treated as indicating exposure below a level of health concern for the endpoints on which the RfD is based, consistent with standard EPA risk assessment practice; values exceeding 1.0 prompt further evaluation of exposure assumptions, uncertainty, and — where litigation-relevant — comparison against the specific health endpoints (male reproductive/endocrine effects, neonatal thyroid disruption) that the RfD is designed to protect against.

7.2 Excess Lifetime Cancer Risk

For cancer risk evaluation, RoseyFields calculates excess lifetime cancer risk (ELCR) using the applicable oral or inhalation slope factor/unit risk from Section 6.2:

ELCR (oral) = Lifetime Average Daily Dose (mg/kg-day) × Oral CSF (1.5 × 105 (mg/kg-day)-1)

ELCR (inhalation) = Air Concentration (pg/m3) × IUR (3.3 × 10-5 (pg/m3)-1)

Resulting ELCR values are compared against the applicable regulatory or litigation-context risk range (commonly 1 × 10-6 to 1 × 10-4, depending on jurisdiction and program). RoseyFields’ cancer-risk calculations always disclose which slope factor vintage was used and the Keenan et al. (1991) pathology-reanalysis caveat described in Section 6.2, and, given TCDD’s extraordinarily high potency relative to most regulated chemicals, RoseyFields applies particular care to units (nanogram, picogram, and part-per-trillion conversions) and to whether a given intake or concentration has been correctly TEF-adjusted before it is entered into the ELCR equation.

7.3 Exposure and Dose Reconstruction

Because TCDD is persistent, bioaccumulative, and eliminated slowly from the human body (elimination half-life 7–12 years), RoseyFields treats serum lipid and adipose tissue concentration as the preferred, most defensible exposure metric wherever biomonitoring data are available or obtainable, using PBPK back-calculation (per Section 3) to reconstruct historical intake where dose timing is relevant to causation or damages analysis. Where biomonitoring data are unavailable, RoseyFields reconstructs exposure using site-specific environmental concentration data (soil, sediment, air, food, or drinking water), standard or site-specific intake parameters, and, where appropriate, food-chain biomagnification and bioaccumulation modeling, consistent with the environmental fate behavior described in Section 1.3. Reconstructed external doses are then compared to background serum TCDD levels in the unexposed U.S. population (typically 3–7 ppt lipid-adjusted, rarely exceeding 10 ppt) to distinguish above-background exposure attributable to a specific source from ambient background body burden — a distinction that is frequently central to causation analysis in litigation matters.

7.4 Litigation and Regulatory Support

In litigation support engagements, RoseyFields uses the methodology above to (i) quantify a claimant’s or class’s estimated dose or reconstructed body burden attributable to a specific release or source; (ii) express that dose as an HQ/HI or ELCR against the current toxicity values, clearly distinguishing the noncancer RfD-based analysis (anchored to reproductive/endocrine and thyroid endpoints) from the cancer slope-factor-based analysis; (iii) compare reconstructed exposure to relevant regulatory benchmarks (e.g., FDA’s 50 ppt action level for TCDD in fish and shellfish tissue, EPA’s drinking water guidance of 0.01 ng/L per day for long-term exposure in children and 0.04 ng/L for long-term adult exposure, or state-specific fish and wildlife consumption advisories); and (iv) transparently document every assumption, uncertainty factor, and study selection decision so that the analysis can withstand adversarial scrutiny and, where relevant, a Daubert/Frye-type reliability challenge. In regulatory-facing work (e.g., NPL site risk assessments, permit support, or agency submissions), the same toxicity values and calculation framework are used to support baseline risk assessment, remedial goal development, and fish/wildlife or soil cleanup level derivation.

8. Key Uncertainties and Sensitive Populations

RoseyFields’ risk assessors flag the following uncertainties and sensitive-population considerations whenever TCDD toxicity values are applied to case-specific facts:

  • Co-exposure confounding: Most human epidemiologic cohorts underlying both the RfD and the cancer classification involved concurrent exposure to other chlorophenols, phenoxy herbicides, and dioxin-like compounds, complicating attribution of observed health effects to TCDD alone.
  • Exposure-timing and critical windows: Both co-principal RfD studies identify early-life exposure (childhood for the Mocarelli sperm endpoint; in utero/perinatal for the Baccarelli thyroid endpoint) as the period of greatest sensitivity, meaning that case-specific exposure timing relative to age and developmental stage is often more consequential to the health-effects analysis than lifetime-average dose alone.
  • Sensitive populations: Fetuses and infants (via placental transfer and breast milk), young children (via higher food intake per body weight and hand-to-mouth soil contact), and individuals with high dietary intake of meat, dairy, or self-caught fish and wildlife are all considered more highly exposed or more sensitive subpopulations. Breast milk can contain higher TCDD concentrations than cow’s milk, though the World Health Organization has concluded that the benefits of breastfeeding outweigh this risk at general population background exposure levels.
  • Cancer slope factor vintage: As discussed in Section 6.2, the operative CSF/IUR values predate the PBPK-based human-data modeling approach used for the current RfD, and the underlying Kociba et al. (1978) tumor pathology has since been reanalyzed with materially different results (Keenan et al., 1991); RoseyFields treats this as an active area requiring case-by-case judgment about which value(s) to present.
  • Small human sample sizes: EPA’s own rationale for retaining an interindividual UF of 3 (rather than reducing it to 1) for the RfD reflects the relatively small size of the exposed Seveso cohorts, which limits the studies’ power to capture the full range of human interindividual variability, particularly for chronic, low-dose, real-world exposure scenarios rather than the acute high-dose accidental release from which the data derive.
  • TEF-mixture uncertainty: Where dioxin-like PCBs, PCDFs, or other CDD congeners co-occur with TCDD, the TEF methodology assumes strict dose-additivity of a common mechanism of toxic action; departures from additivity for specific endpoints or mixtures are a recognized source of uncertainty in TEF-weighted TCDD-equivalent dose estimates.

RoseyFields updates this methodology document as EPA finalizes its pending cancer reassessment (IRIS Volume 2), as new peer-reviewed epidemiologic follow-up data become available from the Seveso and NIOSH occupational cohorts, and as other regulatory bodies revise applicable guidance values.

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This risk assessment methodology was developed under the direction of Dr. Paul Rosenfeld, an environmental scientist with extensive experience in human health risk assessment, toxicology, exposure assessment, environmental chemistry, and regulatory science. Through RoseyFields LLC, Dr. Rosenfeld develops scientifically rigorous methodologies for evaluating environmental exposures, characterizing potential health risks, and interpreting toxicological and regulatory benchmarks across a wide range of chemicals and exposure scenarios. His technical oversight helped ensure that the methodology, scientific interpretation, and risk characterization presented in this report are consistent with accepted scientific principles, current regulatory guidance, and established risk assessment practices.